Conservation biology and applied zooarchaeology 9780816521135, 0816521131

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Conservation Biology and Applied Zooarchaeology

Conservation Biology and Applied Zooarchaeology Edited by Steve Wolverton and R. Lee Lyman

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© 2012 The Arizona Board of Regents All rights reserved www.uapress.arizona.edu Library of Congress Cataloging-in-Publication Data Conservation biology and applied zooarchaeology / edited by Steve Wolverton and R. Lee Lyman.   p. cm.  Includes bibliographical references and index.  ISBN 978-0-8165-2113-5 (cloth : alk. paper)  1.  Conservation biology.  2.  Archaeology.  3.  Animal remains (Archaeology) 4. Animal ecology.  I. Wolverton, Steve. II. Lyman, R. Lee.   QH75.C6614 2012  930.1’0285--dc23 2012015236

Manufactured in the United States of America on acid-free, archival-quality paper containing a minimum of 30% postconsumer waste and processed chlorine free. 17 16 15 14 13 12  6 5 4 3 2 1

Contents

Preface

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  1 Introduction to Applied Zooarchaeology steve wolverton and r. lee lyman

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  2 Zooarchaeological Evidence for Sandhill Crane (Grus canadensis) Breeding in Northwestern Washington State kristine m. bovy

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  3 Archaeological Freshwater Mussel Remains and Their Use in the Conservation of an Imperiled Fauna evan peacock

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  4 Prehistoric Biogeography and Conservation Status of Threatened Freshwater Mussels (Mollusca: Unionidae) in the Upper Trinity River Drainage, Texas charles r. randklev and benjamin j. lundeen   5 Ancient Actions Predict Modern Consequences: Prehistoric Lessons in Marine Shellfish Exploitation heather b. thakar   6 The Overkill Hypothesis and Conservation Biology lisa nagaoka

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  7 Paleozoological Stable Isotope Data for Modern Management of Historically Extirpated Missouri Black Bears (Ursus americanus) 139 corinne n. rosania   8 Rockfish in the Long View: Applied Zooarchaeology and Conservation of Pacific Red Snapper (Genus Sebastes) in Southern California todd j. braje, torben c. rick, and jon m. erlandson

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vi Contents   9 The Past, Present, and Future of Small Terrestrial Mammals in Human Diets karen gust schollmeyer and jonathan c. driver

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10 Applied Zooarchaeology: History, Value, and Use r. lee lyman

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About the Contributors

Index

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Preface

I have wanted to compile and edit a book on applied zooarchaeology for several years, and in 2010 the time was ripe. Charles Randklev and I decided to organize a conference session on applied zooarchaeology, and we chose the Society of Ethnobiology conference as a venue. A growing frustration that I share with several chapter authors is that applied zooarchaeologists seem to increasingly preach to the choir about the merits of deep temporal data in environmental management, and despite a few calls from ecologists, very little attention has been paid to our work in conservation science. An obvious exception is the conservation agenda to “rewild” North America with species analogous to those that became extinct at the end of the Pleistocene, which I find to be a troubling example of applied zooarchaeology. A session at the Society of Ethnobiology was a step toward ecology, but it hardly represents a full journey into conservation science, ecology, and environmental management. Charles and I asked Lee Lyman to serve as discussant for that session, and he later joined me as coeditor of this volume. We agree that applied zooarchaeology has been limited by the reluctance of its practitioners to present and publish in venues that appeal to conservation biologists and ecologists. However, through the process of editing this volume it has become transparent that applied zooarchaeologists face a host of other problems. First, the examples of deep temporal research we provide are often not contextualized very well in conservation science. That is, we rarely explicate the management implications of our research, overlooking that what is of obvious merit to the applied zooarchaeologist may not be so to other conservation scientists. Second, and more important, are the barriers to conservation in general characterized under the realm of political ecology. All conservation actions occur in economic, social, and political contexts. One would think that archaeologists trained as anthropologists would easily recognize this fact, but we tend to be idealistic about the deep temporal perspectives we offer to conservation. A final problem we face, one that is tough to acknowledge and that does not receive much attention, is

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viii Preface that archaeologists are often characterized as pseudoscientists by members of other scientific disciplines. That is, the results of our research are not often taken seriously. In this book, we try to overcome these barriers. Within the volume are chapters that touch on issues of political and social ecology, and a host of case studies for which management implications of zooarchaeological research are clearly stated. Adding a deep temporal perspective to conservation science (here in terms of animal ecology) is not a silver bullet. However, as Lee Lyman has said to me many times, “in light of the current environmental crisis can humans really afford to ignore any source of relevant data?” Lee challenged us in his classic 1996 article published in World Archaeology to integrate zooarchaeology into environmental management. For several years now, zooarchaeologists have tried to do so. This book is a product of those efforts. Steve Wolverton, University of North Texas

Conservation Biology and Applied Zooarchaeology

CHAPTER ONE

Introduction to Applied Zooarchaeology Steve Wolverton and R. Lee Lyman

Applied zooarchaeology is the study of zooarchaeological data sets to provide temporal information on biological and cultural changes and conditions relevant to conservation science (Frazier 2007a; Hadly and Barnosky 2009; Lyman 1996, 2006a; Lyman and Cannon 2004a). Typically, zooarchaeological collections are conceived as sets of ancient animal remains that have associated artifacts or at least owe part of their existence as a collection to human or hominid activity; paleontological collections of ancient animal remains have no associated artifacts and are not the result of human behaviors but instead were created by natural processes of accumulation, deposition, and modification (Lyman 1994). We use the term applied zooarchaeology broadly here to include studies of the paleontological record (e.g., Dietl and Flessa 2009) as well as the zooarchaeological record (Lyman and Cannon 2004b). The term conservation “denote[s] policies and programmes for the long-­­­term retention of natural [biological] communities under conditions which provide the potential for continuing evolution” (Frankel and Soulé 1981:4). Applied zoo­ archaeologists investigate questions such as: What species are native and occurred in an area in the past (Grayson 2006; Lyman 2006b)? What species are exotic, or have been introduced by human agency, to a particular area (Harpole 2004)? What kinds of changes to ecological communities have been caused by humans (Peacock et al. 2005; Randklev et al. 2010; Stahl 1996)? Applied zooarchaeologists are uniquely poised

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to investigate the evolutionary and ecological implications of human-­­ environment interactions (Lyman 2006a; Russell 2003). Applied zooarchaeology offers a distinct perspective on the evolutionary trajectories of ecosystems, both terrestrial (e.g., Graham 1988; Landres 1992; Lyman and Cannon 2004a) and aquatic (e.g., Randklev et al. 2010; Rick and Erlandson 2009). An example of the application of zooarchaeological data to problems in conservation science involves detection of Holocene changes in sea mammal populations (members of Muste­lidae and Pinnipedia) along the North American northwest coast (Etnier 2004; Lyman 1988; Newsome et al. 2007) and the relevance of those changes to the conservation and restoration of particular species. This example and others (Butler and Delacorte 2004; Graham 1988; Humphries and Winemiller 2009; Lepofsky and Lertzman 2008; chapters in volumes edited by Dean 2010; Dietl and Flessa 2009; Frazier 2007b; Lyman and Cannon 2004b; Penn and Mysterud 2007; Rick and Erlandson 2008) highlight the importance of temporal scale in conservation biology and restoration ecology. It is precisely this temporal scale that we perceive as the critical variable that zooarchaeology brings to conservation biology.

Why Is Applied Zooarchaeology Intriguing (among Archaeologists)? Applied zooarchaeology is intriguing to archaeologists because it offers conservation biology what philosopher Albert Borgmann terms a perspective of disclosure. Disclosure entails uncovering knowledge that is not transparent through scientific research without a scale shift beyond humans’ common spatial and temporal experiences (Borgmann 2000). An undisclosed (unrecognized) aspect of the world that each human being experiences daily is the spatial and temporal vastness of the earth. As paleobiologist Stephen Jay Gould (1990:24) eloquently wrote, “Phenomena unfold on their own appropriate scales of space and time and may be invisible in our myopic world of dimensions assessed by comparison with human height and times metered by human life spans.” The scale of a single human lifetime is trivial when compared with “the realm of geological [history]” because that realm is not easily “inhabitable by human imagination” (Borgmann 2000:103; see also Oel­ schlaeger 2000:109–111). The profound time depths of geological and archaeological perspectives are disclosive because these disciplines study temporally and spatially extensive records of historical earth processes (e.g., Gould 1986; Jablonski and Sepkoski 1996; Simpson 1963). These geological, ecological, and biological processes are “historically contin-



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gent”—they have particular tempos, modes, and outcomes as a result of their particular spatiotemporal context. This does not mean we cannot generalize about earth processes such as erosion, but rather that the Grand Canyon is not the same in terms of the historical particulars of its creation as the ditch in a neighbor’s backyard that expands a bit every time it rains. Borgmann (2000) is quite clear that a disclosive perspective is not only about science (which he values) but also about reverence. Applied zooarchaeology, as a discipline that can provide a deep temporal perspective to conservation science, has much to offer conservation biology and restoration ecology in terms of disclosure. Given that it is widely recognized that the contemporary environmental crisis cannot be solved unless values change (Frazier 2010; Lepofsky 2009; Rozzi 1999), it may be that the greatest contribution of applied zooarchaeology transcends the scientific validity of its results, important in their own right, into the realm of reverence for deep temporal contingency in conservation science. That is, the scale shift that applied zooarchaeology provides might prompt a change in values through empirical disclosure of processes that are outside the realm of typical human experience. It is precisely this scale shift that the studies in this book imply is necessary to the future evolutionary and ecological well-­­being of humanity and our coresidents on the big blue marble we call earth. A disclosive perspective is conspicuous in conservationist Aldo Leo­ pold’s essay “Thinking Like a Mountain,” which portrays a value shift that occurred for Leopold when he witnessed a wolf (Canis lupus) die. In his own words, upon encountering a pack of wolves, “in those days we never heard of passing up a chance to kill a wolf. In a second we were pumping lead into the pack, but more with excitement than accuracy” (Leopold 1966). Upon emptying their firearms, the group witnessed the alpha female of the pack dying, and Leopold experienced an awakening that was geologically disclosive as he realized the deep, temporal implications of his and other humans’ anthrocentrism (see discussion in Rolston 1988:62–93). From this direct encounter, Leopold gained what Callicott (1989:3–4) terms an “ecocentric value system,” changing from valuing human individuals exclusively to valuing ecosystems that include humans. That a disclosive perspective was instrumental in his personal change is clearly stated by Leopold: A deep chesty bawl echoes from rimrock to rimrock, rolls down the mountain, and fades into the far blackness of the night. It is an outburst of wild defiant sorrow, and of contempt for all of the adversities of the world. Every living thing (and perhaps many a dead one as well) pays heed to that call. To the deer it is a reminder of the way of all flesh,

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Wolverton and Lyman to the pine a forecast of midnight scuffles and blood upon the snow, to the coyote a promise of gleanings to come, to the cowman a threat of red ink at the bank, to the hunter a challenge of fang against bullet. Yet behind these obvious and immediate hopes and fears there lies a deeper meaning, known only to the mountain itself. Only the mountain has lived long enough to listen objectively to the howl of a wolf. (Leopold 1966:137, emphasis added)

Leopold’s narrative contrasts the short-­­term “obvious and immediate” challenges of humanity to the “objectivity” of a deep temporal or disclosive perspective transparent to him only through geology—that of the mountain. “Aldo Leopold’s land (environmental) ethic is firmly rooted in natural history—in evolution and ecology” (Callicott 1989:11, emphasis added). The adoption of Leopold’s land ethic by modern US citizens has been slow. Adoption of ecocentric values, many of which scale to centuries and millennia, is challenging because national and international economic and political concerns scale to years (Frankham and Brook 2004). Further, most conservation programs are pitted financially against more immediate and for some more pressing economic agendas (Wilhere 2008). These “tensions between science-­­dominated environmental research and management and cultural understanding” occur between science and society and among scholarly disciplines, and make conservation difficult (Head et al. 2005:253). Into this context of political ecology enters the applied zooarchaeologist; the nature of interdisciplinary research is such that communication across boundaries of traditional academic disciplines is required. In a sense this requires researchers to learn the language of new disciplines to communicate effectively; applied zooarchaeologists are still learning that language and also how to play the game (see chapter 10, this volume). The value of crossing disciplinary boundaries in applied zooarchaeological research is clear. Contrasting the short-­term “quick fix” of modernity with a deep temporal record of human-­environment interactions, as does applied zooarchaeology (Grayson 2001; Lyman 1996, 2006a), might encourage values to shift toward ecocentrism because human impacts (including those that occur today) are explicitly acknowledged as geological and evolutionary in scale. Unfortunately, the utility of this disclosive perspective is obscured by a more proximate issue. Though we (applied zooarchaeologists) may be enamored with the temporal perspective we offer, we have not fully demonstrated the merit of that perspective in ways that are practical and meaningful. Applied zooarchaeology must offer results that promote practical solutions to challenges facing conservation biologists.



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Practically Speaking, Whither Applied Zooarchaeology? There is much ado about benchmarks (sometimes referred to as ecological reference conditions) in conservation biology (e.g., Egan and Howell 2001; Hunter 1996; Landres 1992). A benchmark represents a set of environmental conditions that are the target for conservation, but no single benchmark is realistic in terms of how the world actually works. Stream ecologists, for example, argue that there is no particular reference stream that can serve as a model for conservation or restoration (Chessman and Royal 2004). A condition cannot be pinpointed as “natural” or “pristine” for two basic reasons. First, environments fluctuate (perhaps seasonally) around a median over time and they also evolve, and thus represent moving targets for conservation (Landres 1992). It is misguided to set the environmental conditions in 1492 when Europeans arrived in the New World as a conservation benchmark for North and South America. For example, Peacock et al. (2005) demonstrate extensive agricultural impacts on late Holocene pre-­1492 streams in the southern Midwest by prehistoric Native Americans. It is also misguided to select late Pleistocene North America as an ecological benchmark (e.g., Donlan et al. 2005, 2006). North American ecosystems have evolved considerably during subsequent millennia (Callicott 2002; Rubenstein et al. 2006). Further, a terminal Pleistocene benchmark is so coarse in resolution as to invite ecological disaster. The second reason natural pristine conditions are identified with difficulty is made clear by National Park Service historian Paul Schullery (1997). He notes that we are searching for Yellowstone (or whatever our favorite ecosystem or landscape is), and the search is never ending because we define what Yellowstone is or should be (Joyce 2012). The concept of “nature” is what we decide it is; the first European colonists of North America thought they were entering a pristine wilderness—that is, a landscape and ecosystem unmodified by human hands (Grove 1992; Wilson 1992). Relative to what they had emigrated from, North America was indeed undespoiled Eden. Early Euro-­American society sought to utilize that landscape, and this operationalized the originally biblically ordained distinction of humans as separate from, and having dominion over, nature (Bowler 1993). Social and behavioral sciences that emerged in the late nineteenth and early twentieth centuries perpetuated the distinction with the nature versus nurture question. Over the last four or five decades, our society has learned three important things. First, we have learned that there really are few if any intrinsically natural, pristine ecosystems in the sense of a biological, geological, environmental context not somehow influenced to a greater or lesser de-

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gree by humans. Second, we have learned that as our society’s values have changed, so has our conception of nature and how we might attain something that is, somehow, natural and pristine, even if artificially created. Third and finally, we have learned that the concepts of nature, naturalness, pristine, wilderness, and the like are all value laden; they are equal parts of what is out there on the landscape for us to perceive, and what we conceive them to be. The latter might come from a biological or ecological context, an economic context, a political context, or some other context (e.g., Escobar 1999; Greenberg and Park 1994). None of these are in any sense scientifically objective or “right,” but rather each of them is uniquely value laden. As paleozoologists what we can do is provide scientifically, that is empirically, based suggestions to attain and perpetuate whatever chosen value-­laden benchmark is desired. We can do so because the record we work with is temporally both extensive and empirical. The existence of conservation biology and restoration ecology implies that there is some condition that should be conserved or restored (Egan and Howell 2001). Zooarchaeology can play a practical role here because its subject matter tells us something about past biota (Grayson 2001, 2006). The zooarchaeological record is empirical and of greater time depth than the historical record; its study enables conservation scientists to contrast modern environmental conditions with a variety of past ones. Further, it is a record of environmental variability over time and space, the study of which conveys expectations about long-­term rates and magnitudes of environmental change, either free of or including human influences. As such, it reveals a normal range of baseline conditions over the long term (Hadly and Barnosky 2009). This is advantageous because the short-­term observations of neoecologists provide a temporally limited perspective of flux in ecosystems, which inhibits distinction of normal from abnormal flux. Zooarchaeologists provide unique long-­term observations that have a much higher probability of capturing the full normal range of ecosystem states and dynamics. Such will help us to tease out the causes of changes, and whether they are normal and natural, or anthropogenic. Whether or not conservation recommendations can be based on zooarchaeological data will depend on the nature of the zooarchaeological record of an area. Zooarchaeologists are trained to assess such records to determine just what research questions can be examined in a valid manner given preservation and sampling constraints (Jablonski and Sepkoski 1996; Lyman 1994). Zooarchaeological data will not be available for all areas of conservation or restoration interest, but assessing the availability and potential of such data should be a standard order of business in conservation science.



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The recommendations of the applied zooarchaeologist, even with a hefty weight of empirical evidence backing a particular conservation practice, may confront resistance from the general public or from conservation scientists. One of us (SW) has demonstrated, for example, that white-­tailed deer (Odocoileus virginianus) in central Texas today are significantly smaller than during the rest of the Holocene or the last 10,000 years (Wolverton et al. 2007). This relates to the high population density of deer today, which likely relates to extermination of large carnivores (larger than coyotes [Canis latrans]), the displacement of Native American subsistence hunters within the last couple hundred years, and wildlife management practices geared to maximize deer numbers (Côté et al. 2004). SW argues that zooarchaeological data provide disclosure concerning the long-­term effects of Euro-­American land use. This provides a warrant for culling white-­tailed deer populations because in the absence of predation, a high deer population density results in overbrowsing and ecosystem decay (Wolverton 2007; Wolverton et al. 2011). Although wildlife biologists support culling of deer in central Texas, some members of the general public adopt the perspective that deer should not be killed (see also Levy 2006). Resistance to the implications of applied zooarchaeological research may also come from environmental managers. Randklev et al. (2010) demonstrate that the modern freshwater unionid community in the upper Trinity River of Texas comprises species that are tolerant of lentic conditions in impoundment reservoirs and/or to intermittent stream flow from impoundment release. The late Holocene (last 2000 years) unionid fauna was different than that of today, including many species that are today absent. Historic records documenting mussel distributions in the upper Trinity date to the postimpoundment period; the zooarchaeological record is the only record of mussel community composition prior to the creation of modern reservoirs. Successful restoration of the zooarchaeologically documented late Holocene mussel community would require large tracts of unimpounded river. Politically, socially, and economically this target is impossible to achieve in many areas of the upper Trinity. But zooarchaeological data demonstrate what has been lost due to impoundments and may help ecologists identify extant refugia that should be the focus of conservation efforts. Some resource managers have been prompted to act on the basis of zooarchaeological data. Lyman (1998) pointed out that the US National Park Service should review the zooarchaeological record for mountain goats (Oreamnos americanus) in Olympic National Park prior to eradicating the population, which is thought to have originated from individuals that were transplanted to the area in the 1920s. Park officials hired a number of independent researchers to evaluate zooarchaeological and

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other (e.g., historical, biogeographic) evidence that park officials provided. Park officials have thus far not published the results of the review completed in 2000 (the final report is available on the Conservation Biology Institute’s Web site, www.consbio.org, accessed March 11, 2010), nor have they initiated action to eliminate the goat population. Importantly, the Park Service took note and did respond to Lyman’s queries and suggestions. In other cases, biologists may doubt the pertinence of zooarchaeological evidence to conservation issues. For example, some years ago Lyman submitted a manuscript in which he documented that prehistoric Oregon sea otter (Enhydra lutris) teeth were morphometrically intermediate between those of modern Southern California sea otters and Alaskan sea otters. On that basis he suggested that individual sea otters translocated from the latter deme to the Oregon coast may have failed to survive and establish a population because they were phenotypically (and by implication genetically) maladapted to lower-­latitude habitats. Biologists and editors demanded that the suggestion be removed from Lyman’s (1988) manuscript, and it was. Recent study of ancient DNA (Valentine et al. 2008) of prehistoric Oregon sea otters and modern sea otters from more northern and more southern latitudes has shown that Lyman may well have been correct. Prehistoric Oregon sea otters are genetically more like California sea otters and dissimilar to Alaskan sea otters from which transplants were derived. It remains to be seen if this zooarchaeological result will influence future efforts to reestablish midlatitude populations of this charismatic animal.

Political Ecology and Applied Zooarchaeology Early efforts to apply zooarchaeological findings to conservation issues had minimal influence on conservation biology or on its attendant policies. As noted above, not only did Lyman’s (1988) findings on sea otters not sway biologists, those findings were thought to be invalid by some. Other similar experiences are known to us (e.g., D. Gifford-­Gonzalez, University of California-­Santa Cruz, personal communication, April 2011). What the early applied zooarchaeologists failed to recognize was that their efforts were being played out in an arena known as political ecology (Greenburg and Park 1994; Ludwig et al. 2001). Although the term and concept have been in the anthropological literature for some time (e.g., selected references in Vayda and Walters 1999), and in fact the origins of the term are often traced to anthropologist Eric Wolf (1972), it has not been used much if at all in the literature of applied zooarchaeology. It is



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important in the context of this volume to devote a bit of space to the topic. Applied zooarchaeology and political ecology are loosely integrated in contributions in Kay and Simmons’s (2002a) volume. The major focus of that volume is demonstration that the conception of North America as a pristine natural wilderness in 1491 AD is false. Kay and Simmons (2002b:xv) noted that they regularly found that when they presented this fact it was ignored or said to be invalid by ecologists who seemed to use ecological data “to support preordained philosophical values or political agendas”; hence the title of their book Wilderness and Political Ecology. The fact that anthropologists have long known that pre-­1491 North American ecology was to at least some degree anthropogenic (e.g., Adler 1969; Day 1953; Elder 1965; Heizer 1955) underscores the failure of anthropologists to effectively communicate with ecologists. Perhaps the failure is benign and comes from a lack of forcible lobbying for conceptual change, or perhaps it is rooted in fear that comments questioning the traditional concept of ecologically noble primitive peoples are too often perceived as politically incorrect (e.g., Hames 2007; Krech 1999, 2005). Those with serious commitments to applied zooarchaeology can ill afford to neglect, and must be wary of, the political ecology context in which they practice. Simplistically, political ecology concerns the social, political, economic, ecological, and any other sort of human-­interest context in which conservation biology, restoration ecology, and landscape management occurs (Greenburg and Park 1994; Walker 2006). Do not be misled here. Notice that the fourth context we identified is “ecological.” That is, political ecology does not just concern human values and variables and processes, but also ecological variables and processes, despite suggestions that biophysical ecology is often ignored (Walker 2005). Political ecology involves the dynamics and evolution of the relations and dialectic between human society and the environment (Nygren and Rikoon 2008; Walker 2005). Those relations are constantly changing over time, and they vary across geographic space at any one time as well. Those with interests in applied zooarchaeology cannot, therefore, blithely present their data and analytical results and expect conservation biologists or policy makers to react immediately, or to react at all. Rather, applied zooarchaeologists must be cognizant of the ongoing and evolving political and economic struggles over natural resources (animal, vegetable, mineral), the indigenous cultural meanings and significance of those resources (which is unabashedly anthropological), and the fact that environments (e.g., climatically or anthropogenically driven) change and have attendant cascade effects through ecosystems, societies, economics, and politics.

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Although awareness of particular political ecological contexts is advisable, deep and intimate knowledge of such contexts is, we think, unnecessary among applied zooarchaeologists. We say this because despite years of effort, even conservation biologists are still grappling with influencing policy makers (Robinson 2006). Similarly, political ecologists seem unable as yet to successfully and frequently engage and influence policy makers (Walker 2006). And while this does not excuse conservation biologists, political ecologists, or applied zooarchaeologists from working within particular political ecology contexts, it does indicate the slippery slope represented by those contexts. None of the case studies making up the chapters in this volume delve deeply into the particular political ecological contexts to which a set of zooarchaeological remains pertains, yet all of the authors are quite aware of such contexts. It is the simple hope of the authors of those chapters, and our hope as well, that the weight of the evidence from a combination of fields—ecology, biology, history, zooarchaeology, paleontology—will eventually gain the attention of policy makers. That, it seems to us, is the critical first step; engagement with and influence on policy are steps for the future.

The Chapters Regardless of how the public or conservationists respond to us, how do we get their collective attention? We contend that it is critical that we publish our research in biological and ecological scholarly journals. Many of the authors of chapters in this book have published in conservation biology, ecology, and biology journals. This volume of essays does not contradict this suggestion. Rather, we believe that, because of its breadth of taxonomic and geographic coverage, the set of case studies here will be more convincing of the value of applied zooarchaeology than any single isolated case study in a journal. We also believe that it is important that applied zooarchaeologists regularly present their research at regional conference venues that environmental managers attend. All of these activities will activate, build, and exercise our voice in environmental management and conversation. With those goals in mind, each case study presented in this volume concludes with a section on management implications and precise recommendations that relate to wildlife biology and conservation science. What, then, are the chapters about? As Lyman points out in chapter 10, applied zooarchaeology has a temporally deep if not particularly extensive history. Nevertheless, those who have done applied zooarchaeology have addressed a number of topics. These include determining which taxa are native to an area and which are exotic (e.g., Langemann 2004), and the related deciphering of which taxa



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are currently invading an area and which are simply recolonizing it (e.g., Lyman et al. 2002). Because zooarchaeological data provide multiple results of long-­term variously natural and anthropogenic experiments, they can be used in an applied context to choose a management action that can be shown to produce a desired result (e.g., Lyman 2006b). They can be used to test the validity of benchmarks suggested by written historic documents (e.g., Etnier 2002). Some zooarchaeological data reveal unanticipated effects of early conservation efforts (e.g., Lyman 2006c). Finally, these data help identify what may happen in the future should recent trends in climate change such as global warming continue (e.g., Barnosky 2009; Blois et al. 2010; Grayson and Delpech 2005). The value of zooarchaeological data seems limited only by the thoughtfulness of those who seek to use it in the service of conservation biology. The chapters in this volume exemplify many of these uses, and more. The volume is organized into two sections. The first presents four case studies that relate to geographic ranges of animal species today and in the past. Kris Bovy discusses the late Holocene breeding range of the sandhill crane (Grus canadensis) in the Pacific Northwest of North America in chapter 2. Bovy’s study highlights historic-­period range reductions and local extirpations; an important management implication for threatened species and subspecies is to locate suitable habitat for range expansion that can be protected. Evan Peacock highlights the conservation implications of paleobiogeographic studies of freshwater mussels in chapter 3. He argues that applied zooarchaeology involves methodological challenges and epistemological differences that exist across traditional scholarly boundaries. Peacock demonstrates his position with a case study focusing on stream ecology in Mississippi. Charles Randklev and Benjamin Lundeen (chapter 4) demonstrate that during the late Holocene the community composition of stream fauna and the biogeography of several species was much different than today. Many of the mussel species they discuss are of conservation concern in Texas and other areas of the United States. Their discussion provides an important context for conservation of those species. Heather Thakar’s case study (chapter 5) on the marine Pismo clam (Tivela stultorum) along the western coast of North America highlights the importance of metapopulation ecology and habitat connectivity for species of concern across temporal and spatial continua. In particular, species with limited breeding opportunities across disconnected local populations are vulnerable to overexploitation by humans, particularly in isolated (e.g., island) situations. The second set of chapters includes four case studies that focus on temporal variability in behavioral ecology or population ecology of animals. Lisa Nagaoka (chapter 6) discusses differences in behavioral ecology in island and continental terrestrial species with reference to the

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conservation implications of the Pleistocene extinctions and rewilding. Pleistocene rewilding, the proposal to introduce to North America the extant species most closely related to those that became extinct at the end of the Pleistocene, is a controversial proposal that is a commonly cited example of applied paleozoology. Proponents of rewilding argue that because humans caused the extinctions through overkill, there is merit to the rewilding proposal. Overkill is based on several assumptions that relate to the ecology of island extinctions, and Nagaoka critically examines those assumptions, assesses the validity of rewilding, and addresses why Pleistocene overkill and rewilding exemplify applied zooarchaeology. Corinne Rosania (chapter 7) focuses on the behavioral ecology of North American black bears (Ursus americanus) in Missouri. Local black bear populations were dramatically reduced and perhaps extirpated about 200 years ago. Very little is known about the historical population, and a transplanted population in northern Arkansas has expanded into Missouri. Rosania uses isotope chemistry to determine whether or not late Holocene bears sampled from paleontological contexts differ from modern bears in terms of diet. She is able to make recommendations about habitat availability and conservation of the modern bear population in the state. Todd Braje, Torben Rick, and Jon Erlandson (chapter 8) provide a detailed perspective on the rockfish fishery of coastal California. Through biometric analysis of size changes during the Holocene, Braje et al. assess long-­term harvest pressure and make comparisons to the modern fishery. Their assessment of long-­term human-­rockfish relations provides a template that informs attempts at sustainable fisheries management. Karen Schollmeyer and Jonathan Driver also discuss records of long-­ term human-­prey relationships in chapter 9. They demonstrate that in distinctive areas of the world during the Holocene, humans consistently harvested a variety of terrestrial small game species, even during periods of large game animal decline. They argue that the zooarchaeological record provides evidence of a sustainable harvest strategy that is relevant to modern societies in many parts of the world that supplement their diets with garden hunting. In the final chapter, R. Lee Lyman discusses the current state of applied zooarchaeology and identifies the broader implications of each chapter.

The Applied Zooarchaeologist The summary of chapters provides hints as to the nature of an applied zooarchaeologist as a scholar. An applied zooarchaeologist is a scientist who straddles a variety of disciplines (Lyman 1996). She or he must em-



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brace the discomfort of presenting and publishing research outside of her or his own discipline (e.g., Lyman 2006b, 2006c). Straddling a number of disciplines is also financially burdensome. The zooarchaeologist might receive institutional funding to attend one conference per year; to attend a biology conference entails missing one in a home discipline with colleagues and friends. In addition, publishing outside of one’s home discipline is not always recognized and rewarded in academia, thus jeopardizing the possibility of winning promotion and tenure. Finally, an applied zooarchaeologist can find it difficult to publish in ecological journals because those venues represent an unfamiliar style of writing and an audience unfamiliar with paleozoology. It is these discomforts that have led to many examples of applied zooarchaeology being presented and published in archaeological or paleontological venues (e.g., Dietl and Flessa 2009; Lyman 1996, 2006a). Conservation biology, with few exceptions (e.g., Frazier 2007b), has not called upon zooarchaeology for recommendations about conservation practice. Further, we have presented case studies in applied zooarchaeology primarily to ourselves instead of to conservation biologists, though this is starting to change. We are encouraged that recently applied zooarchaeological case studies have been published in ecological or biological journals (Campbell and Butler 2010; Lyman 2006b, 2006c; McKechnie 2007; Murray 2008; Randklev et al. 2010; Wolverton et al. 2007). The applied zooarchaeologist brings an empirical record of particular cases that conservationists only encounter conceptually (see discussion by Humphries and Winemiller 2009). Many conservationists and restorationists have found zooarchaeological data to be too unlike the neoecological data they are familiar with and thus discount the value of the former; fortunately, that perception is changing (e.g., Morrison 2001), though slowly. Zooarchaeologists are aware of the utility of paleofaunal data for addressing questions of spatial and temporal variability in taxonomic composition, body size, demography, and other variables, as well as the difficulties and limitations of rendering paleoecological data relevant and useful to conservation and restoration. Zooarchaeologists are also aware of the kinds of qualitative and quantitative analyses that can be validly applied to their data (Jablonski and Sepkoski 1996). Zooarchaeological quantitative data are at best ordinal scale, meaning that many of our studies and recommendations will be at the nominal scale (Grayson 1984; Lyman 2008). It is the responsibility of the applied zooarchaeologist to communicate both the utility and limits of such data for conservation. The ecocentric ethic that Aldo Leopold proposed is a natural fit for applied zooarchaeology. But Leopold’s perspective is distant from the reality of conservation practice for the same reason that applied zooarchae-

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ology is. Among philosophers the ecocentric ethic is intriguing (Callicott 1989), but it does not offer a useful product given economic, political, and social reality. Applied zooarchaeology has a distinct advantage in that it is empirically founded. Leopold’s perspective can thus be offered to conservation with tangible research products independent of or alongside ethical positions. Applied zooarchaeologists have made the mistake of maintaining that our research products are of merit without adequately addressing practical conservation challenges such as the realities of political ecology.

Articulating Zooarchaeology and Conservation Biology Lindenmayer and Hunter (2010) have laid out key goals, strategies, and limitations for conservation biologists (table 1.1). We find zooarchaeological research to be readily articulated (some might say integrated) with those goals and strategies, and able to overcome some of the limitations. The two general goals are that explicit conservation targets be stated and that conservation not be for species richness alone but for biodiversity in general, thus accounting for richness, evenness, and sensitive rare species. The first goal relates to political ecology and is important for applied zooarchaeologists because they must be aware that their ideas and recommendations occur in the same sociopolitical contexts as all conservation actions. This goal relates closely to the “constraint” (table 1.1, point 10) that “human values are diverse” and they shape effective conservation. The second goal relates directly to what applied zooarchaeology can offer to conservation science: it can provide information on long-­term variability in biodiversity, species richness, and evenness and on rare and common species in many areas of the world (Hadly and Barnosky 2009; Lyman 1988, 2006a). In light of these two general goals, Lindenmayer and Hunter (2010) suggest a number of strategies (table 1.1). Points 3 to 5, “a holistic approach is needed to solve conservation problems”; “diverse approaches to management”; and the limits of “using nature’s template,” provide warrants for the use of zooarchaeological data in conservation science, but their holistic inclusiveness mentions only recent historical examples. Disturbance regimes are construed mainly in terms of classic ecological examples, and temporal perspectives are paid minimal lip service. Point 6 relates directly to applied zooarchaeology; causation implies a temporal perspective that should be critical in terms of assessing the “symptoms of poor environmental management.” Much of the discussion of point 6



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Table 1.1.  Guiding Concepts for Conservation Biology Category Goals 1 2  Strategies 3 4 5 6 Constraints and Considerations 7 8 9 10

Title Successful conservation management requires achievement of consensus on explicit goals and perspectives The overall goal of biodiversity management will usually be to maintain or restore biodiversity, not to maximize species richness A holistic approach is needed to solve conservation problems Diverse approaches to management can provide diverse environmental conditions and mitigate risk Using nature’s template is important for guiding conservation management, but is not a panacea Focusing on causes, not symptoms, enhances efficacy and efficiency

Every species and ecosystem is unique, to some degree Threshold responses are important, but not ubiquitous Multiple stressors often exert critical effects on species and ecosystems Human values are diverse and dynamic and significantly shape conservation efforts

Adapted from Lindenmayer and Hunter (2010).

criticizes examples of poor ecological restoration that occurred because the “underlying degrading processes [were] not addressed” (Lindenmayer and Hunter 2010:1463, emphasis added). Like causation, process implies time and thus implicates zooarchaeology. Applied zooarchaeologists might also offer important perspectives for “constraints and considerations” discussed in points 7 to 10. The “uniqueness” of species and ecosystems (not to mention communities) would be better understood through consideration of changes (or lack thereof) through time. Identifying changes and their causes necessitates a deep temporal approach such as the National Science Foundation–funded Long Term Ecological Research Program (Hobbie 2003; Kaiser 2001). The temporal perspective offered by applied zooarchaeology is critical for understanding long-­term contingency and continuity of ecological thresholds for species and ecosystems (point 8), which are “changes in an ecological driver that generate large responses” (Lindenmayer and Hunter

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2010:1464). Similarly, the uniqueness of species and ecosystems (point 7), the nature of “threshold responses” and whether or not they exist within an ecosystem (point 8), and the “cumulative effects” that make conservation a multivariate problem (point 9) all require consideration of many forms of ecological data including those spanning long temporal durations. The popularity of archaeology in society (point 10) renders it an as yet little-­used tool for introduction of conservation ethics and agendas in the context of public outreach and communication. We suspect many conservation biologists, including Lindenmayer and Hunter, have not read applied zooarchaeological literature and seldom consider a truly deep (disclosive) temporal perspective. For example, authors of an article on conservation literacy touch on “impacts of human colonization in ancient times” (Lusk et al. 2004:1185) but they cite an outdated, controversial example (the Pleistocene extinctions [Martin and Klein 1984]) that does not represent the systematic approach that modern applied zooarchaeology has to offer. Applied zooarchaeologists have not yet convinced the general conservation community of the merit of our approach for a holistic conservation biology, even though a few conservation biologists call for inclusion of applied zooarchaeological and other forms of temporal research in conservation (Bjorkman and Vellend 2010; Frazier 2007a; Humphries and Winemiller 2009). We see relevant applications of applied zoarchaeological research related to all 10 of the points identified by Lindenmayer and Hunter (2010) as central to conservation biology. Given this nuanced articulation of these two otherwise disparate disciplines, we challenge conservation managers to employ applied zooarchaeologists to understand past characteristics of ecosystems, communities, and species as a regular part of conservation biology. Applied zooarchaeologists, because of the records we study and the skills we possess, are in a unique position to provide pragmatic recommendations relevant to many aspects of environmental management to help achieve the holistic approach that Lindenmayer and Hunter (2010) advocate. Whether or not those recommendations will be heard and/or adopted depends on a host of factors that can be only partially controlled by the zooarchaeologist, such as where we present and publish our research. We cannot control whether conservationists or restorationists acknowledge our contributions to their fields, but we are obligated to give them the opportunity to acknowledge what we can bring to their table. That opportunity is created by presenting our data and conclusions in their venues. Our goal here has been to produce a volume that extends well into the conservation world and beyond that of typical archaeology. We are confident that paleozoologists will find much of interest in the pages that follow, and sincerely hope that conservation biologists and restoration ecologists will find much of value.



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CHAPTER TWO

Zooarchaeological Evidence for Sandhill Crane (Grus canadensis) Breeding in Northwestern Washington State Kristine M. Bovy

Some years ago British paleobiologist John Stewart (2004:42) suggested that zooarchaeological data for birds might help “reconstruct the so-­ called ‘natural’ condition of [an area’s] avifauna of the past”; thus those data should “be reported to organizations which determine conservation policy.” Stewart was too understated; in my view, paleo-­­­ornithological data are critically important for purposes of modern avifauna restoration and conservation. In this chapter I present an example of exactly how such data can inform modern management of our avifauna. Sandhill crane (Grus canadensis) populations were significantly reduced over much of the United States during the twentieth century as a result of overhunting and loss of habitat, and are still an endangered species in Washington State. There is a lack of historical baseline data on breeding sandhill crane populations for much of Washington, and therefore management and recovery efforts have targeted the small breeding populations in the south-­central portion of the state where the historical distribution is better understood. Recently analyzed zooarchae­ ological remains of sandhill cranes from Watmough Bay (45-­SJ-­280), a

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1500-­year-­old shell midden located on Lopez Island in the San Juan Islands, reveals that this species may have once also bred in the northwestern portion of the state. This new finding demonstrates the value of zooarchaeological data for informing management actions regarding species whose continued existence may be in question.

Historical Information on Sandhill Cranes in the Pacific Northwestern United States In 1859, Lieutenant George Suckley, an army doctor stationed in Washington State, commented, “Sand-­hill Cranes are very abundant on the Nisqually Plains, Puget Sound, in autumn. They there commence to arrive from their summer breeding grounds about the last week in September, from which time until the 10th of November, they are quite plen­ tiful” (Cooper and Suckley 1859:227). Suckley’s colleague, Dr. James Cooper, added that cranes were also a “common summer resident, arriving at the Straits of De Fuca in large flocks in April, and then dispersing in pairs over the interior prairies to build their nests” (Cooper and Suckley 1859:227). Historical documents such as these reveal that sandhill cranes were once much more abundant than today (Jewett et al. 1953; Lewis and Sharpe 1987). Sandhill crane populations faced “near extermination” during the nineteenth and early twentieth centuries, with declines throughout their North American range (Littlefield and Ivey 2002:18). In 1925, the greater sandhill crane was given the subspecies name tabida, meaning “wasting away,” in reference to their declining habitat and numbers (Cooper 1996:6). In Washington, breeding populations of sandhill cranes were extirpated from 1942 through 1971 (Littlefield and Ivey 2002). Likewise, many local breeding populations were reduced or disappeared in southern British Columbia during this period (Cooper 1996). For example, in his ethnography of the Katzie First Nation, Jenness (1955:9) commented that “thousands” of sandhill cranes once arrived in the meadows near Pitt Lake in the Fraser Valley to feed and breed; today only a handful of cranes breed there (Marks et al. 2008). Loss of habitat, disturbance, and hunting were primary causes of population decline (Leach 1979). Pioneers referred to sandhill cranes as “wild turkeys” and considered them to be “good eating” (Leach 1979:4). In the words of Lieutenant Suckley, the cranes were “excellent for the table” (Cooper and Suckley 1859:227). In the mid-­1800s, market hunting for sandhill cranes began in California, an important breeding area, where they were even touted as a substitute for the Christmas turkey (Littlefield and Ivey 2002:17). Market hunting increased from 1880 until



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the passage of the federal Migratory Bird Treaty Act in 1916, which ended the hunting of cranes (Leach 1979; Littlefield and Ivey 2002).

The Modern Situation Although it is clear sandhill cranes were once more abundant and widespread, their exact historical distribution in Washington is unknown, due to both ambiguities in historical accounts and confusion over the taxonomic status of the subspecies (Ivey et al. 2005; Littlefield and Ivey 2002). Three subspecies of sandhill cranes occur in the state. The large greater sandhill crane (G. c. tabida) is widespread across North America; all the breeding cranes found in Washington State today are part of the Central Valley population of greaters, which winter in the Central Valley of California. The smaller lesser sandhill crane (G. c. canadensis), formerly the “little brown crane” (Jewett et al. 1953), breeds in the Arctic and migrates through the Pacific Northwest on the way to their wintering grounds in California. The status and distribution of the Canadian sandhill crane (G. c. rowani), which migrates through western Washington en route to breeding sites along coastal British Columbia and Alaska, is the least well known (Cooper 1996; Littlefield and Ivey 2002). While mitochondrial DNA research has suggested the Canadian subspecies is not significantly different than the greater subspecies, tracking studies indicate they have different annual migrations and should be managed separately (Ivey et al. 2005). Historical records indicate that sandhill cranes once bred in both western and eastern Washington (Jewett et al. 1953). There are numerous known breeding locations east of the Cascade Mountains (figure 2.1), but the only reliable nesting account for western Washington is at Fort Steilacoom in southern Puget Sound. Today, greater sandhill cranes breed only in the south-­central portion of the state; in 2002, there were 19 breeding pairs, all of which were located in Klickitat and Yakima counties (Littlefield and Ivey 2002). There is also a small breeding population of sandhill cranes near Vancouver, British Columbia (figure 2.1). While the sandhill crane was once designated as a “blue-­listed” species (“of special concern”) in British Columbia, their status has been changed to “yellow” (“not at risk for extinction”) (Marks et al. 2008). The sandhill crane remains an endangered species in Washington State (WDFW 2010). Various measures have been proposed and enacted to help stabilize sandhill crane populations in Washington and British Columbia in the past 30 years, including habitat protection, reintroductions, reduction of human disturbance and hazards, and public education (Cooper 1996; Gebauer 2004; Leach 1979; Littlefield and Ivey 2002). In Washington,

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Figure 2.1.  Known historic and current sandhill crane breeding locations in Washington State and southwestern British Columbia (adapted from Littlefield and Ivey 2002). Place names mentioned in text are also shown.

the effort to manage breeding birds has focused on the south-­central portion of the state, especially the Conboy Lake National Wildlife Refuge (figure 2.1).

Materials and Methods The Watmough Bay site (45-­SJ-­280) is located on a narrow bay surrounded by steep cliffs and a freshwater marsh (figure 2.2). A University of Washington field school excavated the site in 1968, but recovered materials were never fully analyzed. The original notes and maps, as well as the artifacts, are curated at the Burke Museum of Natural History and Culture on the University of Washington campus. The bird bone analysis was part of a larger analytic effort to investigate the effects of climate change, tectonic events, and human hunting on waterbird populations along the Pacific Northwest coast (Bovy 2005, 2007a, 2007b). I identified over 12,600 bird bones from three archaeological sites, and here focus only on those of the sandhill cranes from Watmough Bay. The site was excavated in 2 × 1 m units and 20 cm arbitrary levels. A total area of 37 m2 (~67 m3) was excavated. All excavated matrix was screened through ¼" mesh (6.35 mm), and artifacts, faunal remains, and



Zooarchaeological Evidence for Sandhill Crane Breeding

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Figure 2.2. Location of the Watmough Bay site on southeastern Lopez Island.

other items of interest were collected. I analyzed the bird bones from ten 1 × 2 m units and two 1 × 1 m units, with depths varying from 160 cm to 210 cm below the surface. I made no attempt to identify vertebrae or ribs below the class level. I quantified the bird bones using the number of identified specimens (NISP). Although it might seem more appropriate to determine the number of individual cranes in the collection given a desire to know how many sandhill cranes were in the site area, there are robust theoretical and empirical reasons to rely on NISP (Grayson 1984; Lyman 2008). In brief, the number of individuals in a paleozoological collection is a function of NISP and thus the two quantitative units are redundant with one another. Further, there is no valid technique to determine if, say, a right humerus is from the same individual animal as a left tarsometatarsus unless they are found articulated during excavation. Thus I rely on NISP here. While NISP values may be inflated due to fragmentation, those values may also be reduced by extensive fragmentation (Marshall and Pilgram 1993). Fragmentation is not a concern for the sandhill crane assemblage as many of the specimens are nearly complete, and fragmented specimens recovered from the same excavation levels were refitted to control for specimen interdependence. Taxonomic identifications were made using the extensive comparative collections of the zoology department of the Burke Museum. The use of

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Kristine M. Bovy

comparative collections is critical to producing valid taxonomic identifications regardless of whether birds, mammals, mollusks, or some other group is the subject of study (e.g., Driver 1992; Lyman 2010); taxonomic identification of zooarchaeological bird remains in particular has been the subject of some concern (e.g., Bochenski 2008; Stewart 2005). The sandhill crane is an old species, with fossils dating to 2.5 million years ago (Tacha et al. 1992), and has no close relatives in the Pacific Northwest. Adult sandhill crane bones are therefore quite distinctive and are unlikely to be mistaken for other species, even the great blue heron (Ardea herodias), which, in life, are often confused with sandhill cranes (Gebauer 2004; Lewis and Sharpe 1987). Immature specimens are more challenging to identify, however, because they often lack taxonomically diagnostic features and suitable comparative specimens of juvenile animals are rare. Since I did not have access to any comparative specimens of immature sandhill cranes, careful analysis was required to distinguish the subadult crane bones from those of the great blue heron. Fortunately, there were 108 subadult heron bones in various stages of development in the Watmough Bay assemblage. Thus I was able to trace the development of heron bones from chick to adult, and could use them for comparison with the immature sandhill crane specimens. While most skeletal elements are taxonomically distinct morphologically and/or metrically, even among the subadult specimens, the radius and ulna are difficult to distinguish, especially when fragmented. Criteria for distinguishing these two elements are described in Bovy (2005:338). Given the lack of adequate comparative specimens of immature sandhill cranes, I made no attempt to distinguish between the different subspecies of cranes and I conservatively identified some of the archaeological specimens as “cf. Grus canadensis.” I assigned an age class to each specimen following Broughton (2004:8). “Chick” bones are porous and much smaller than those of adults, and lack adult cortical bone and muscle attachments. “Juveniles” are similar in size to adults but lack completely developed cortical bone. Unfortunately, it was not possible to securely assign specific age ranges to these age categories, given the lack of immature comparative specimens of known ontogenetic age.

Results Nineteen sandhill crane bones (table 2.1) were identified from the large Watmough Bay bird assemblage (site total NISP = 7597), which is dominated by diving ducks and cormorants (Bovy 2005, 2007b). Three of the

29

Ulna (distal) Ulna (distal) Radius (complete)

Femur (complete)

Tibiotarsus (distal) Tarsometatarsus (proximal)

Tarsometatarsus (distal)

Tarsometatarsus (complete)

Phalanx (complete) Phalanx (complete)

Chick Chick Chick

Chick

Chick Chick

Chick

Chick

Chick Chick

0N6W, 100–120 0N6W, 100–120

Baulk C, 100–120

0N6W, 100–120

0N6W, 100–120 0N6W, 100–120

1N9W, 80–100

0N3W, 80–100 0N6W, 80–100 0N6W, 80–100

0N24W, 100–120 0N3W, 120–140 1.5N0E, 100–120 0N18W, 100–120

0N9W, 30–50 0N15W, 90–110

0N9W, 30–50 0N15W, 90–110 0N0E, 80–100

Provenience1

2

1

Provenience: excavation unit, depth (centimeters below surface). LB: level bag.

R

L

L L

L

R R R

R R

Tarsometatarsus (distal) Tarsometatarsus (distal) Phalanx (proximal) Phalanx (complete)

Juvenile Juvenile Juvenile Juvenile

R R L

Side

L L

Humerus (proximal) Humerus (distal) Femur (proximal)

Element (Segment)

Juvenile Tarsometatarsus (distal) Juvenile Tarsometatarsus (distal)

Adult Adult Adult

Age

22 22

202

23

22 23

90

115 21 21

5 117 99 126

131 31

131 31 12

LB No.2

Distal condyles not well developed; 3 fragments Proximal and distal not well developed; 3 fragments Proximal and distal only partially formed; 2 fragments Distal not fused Proximal not well developed; may fit with distal in same level bag Trochlea partially developed; may fit with proximal in same level bag Morphologically like crane, but very short (small chick?); 2 fragments

Size of adult; trochlea fairly well developed Distal trochlea broken except trochlea for metatarsal 2 Size of adult; distal trochlea broken Size of adult; distal trochlea broken

Shaft is broken into many fragments Shaft is modified (“grooved and snapped”)

Comments

Table 2.1. Sandhill Crane (Grus canadensis) Bones Recovered from the Watmough Bay Site (45-SJ-280)

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Kristine M. Bovy

sandhill crane specimens are from adult birds. One of the adult bones is a distal humerus fragment that was cut from the shaft using the “groove-­ and-­snap” technique. Parmalee (1977, 1980) has argued that while the main shaft of wing bones was often used for tools and/or ornaments, these groove-­and-­snap proximal or distal end-­of-­the-­skeletal-­element specimens are the discarded portions of the bones. Of the 16 subadult sandhill crane bones, 6 are classified as juvenile and 10 as chick. The most abundant element (NISP = 7) is the tarsometatarsus. Figure 2.3 shows the age progression of these tarsometatarsi from oldest (a modern adult comparative specimen) to youngest. The juvenile specimens are similar in size to those of adults and have relatively well-­developed trochlea (when present), while the chick specimens are much smaller and have poorly developed proximal and distal ends. The chick ulnae, radius, femur, and tibiotarsus from the assemblage are shown in figure 2.4, along with modern adult specimens for size comparison (note: although the subspecies designations were not indicated on the comparative specimens, both were from Harney County in southeastern Oregon and are therefore most likely the larger G. c. tabida). None of the subadult specimens exhibit cultural modifications such as burning or cut marks; four of these bones were fragmented, and all but one (the femur) appear to be recent breaks that occurred during or after excavation. The subadult assemblage contains three distal right tarsometatarsi, indicating that at least three individuals are represented. There is a cluster of seven chick bones in excavation unit 0N6W (level bags 21–23); these seven bones are likely all from one individual. The remaining subadult bones are spread over numerous excavation units and layers, and so the actual number of immature birds may be greater. For example, all six of the juvenile bones were recovered from different units (table 2.1). All but one of the subadult specimens (a distal tarsometatarsus from 0N9W) are from lower layers of the site (below 80 or 90 cm). Extensive radiocarbon dating indicates that the bulk of the shell midden (30–120 cm below surface) was deposited AD 300–600 (Bovy 2005; Deo et al. 2004; Stein et al. 2003). There is a smaller older component at the site (950–550 BC), but it contained only 95 identifiable bird bones, none of which were sandhill cranes.

Discussion Knowledge of sandhill crane breeding ecology, along with close examination of the ethnographic and archaeological records, is necessary to better understand when and where the immature sandhill cranes from Wat-



Zooarchaeological Evidence for Sandhill Crane Breeding

Figure 2.3.  Subadult sandhill crane tarsometatarsi from Watmough Bay arranged from oldest (an adult comparative specimen, left) to youngest (chick, right). (a) Left and right tarsometatarsi from an adult female sandhill crane from Malheur National Wildlife Refuge, Oregon (UWBM #59275); (b) juvenile, distal left (LB 131); (c) juvenile, distal right (LB 5); (d) juvenile, distal right (LB 117); (e) chick, proximal left (LB 23); (f) chick, distal left (LB 23); (g) chick, right (LB 202). The juvenile left distal tarsometatarsus from LB 31 is not shown because the specimen has been repatriated. Photo courtesy of the Burke Museum of Natural History and Culture, Catalog #45-SJ280/LB 5, 23, 117, 131, 202.

31

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Kristine M. Bovy

Figure 2.4.  Sandhill crane chick bones compared to modern adult specimens (left of each photo). The comparative specimen is an adult male from Burns, Oregon (UWBM #81894). (a) Distal right ulnae (LB 21 and LB 115); (b) right radius (LB 21); (c) left femur (LB 90); (d) distal left tibiotarsus (LB 22). Photo courtesy of the Burke Museum of Natural History and Culture, Catalog #45-SJ-280/LB 21, 22, 90, 115.



Zooarchaeological Evidence for Sandhill Crane Breeding

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mough Bay may have been hunted. Spring migration of sandhill cranes occurs along coastal British Columbia during March and April (Cooper 1996). Sandhill cranes nest in a variety of habitats throughout their range, but usually choose relatively inaccessible locations isolated from human activity (Cooper 1996, 2006; Littlefield and Ivey 2002). Breeding cranes in the Fraser Valley typically build mounded nests in shallow wetlands with emergent vegetation such as sedges and bulrush stems (Cooper 1996). Sandhill cranes usually lay two eggs, although often only one survives to fledge (Tacha et al. 1992). Young hatch between mid-­May and late June and are able to leave the nest within 24 hours after hatching (Tacha et al. 1992). During the first few weeks of their lives chicks stay within a few hundred feet of the nest, and then gradually expand their ranges (Tacha et al. 1992). Average home range size for breeding cranes in the Pacific Northwest varies from 25.3 ha at the Malheur National Wildlife Refuge in Oregon to 140 ha at Conboy Lake National Wildlife Refuge (Littlefield and Ivey 2002); breeding territory sizes in British Columbia are not known (Cooper 1996, 2006). After 20 to 30 days, the wings and legs of sandhill crane young are about half the size of adults’; after 40 days the legs are almost full grown, but wings do not approach adult size until cranes are 60 days old (Tacha et al. 1992). Growth is not complete, however, until about 10 months of age (Tacha et al. 1992). Sandhill cranes fledge when about 65 to 75 days old and are strong fliers within a few weeks thereafter (Tacha et al. 1992). The young cranes depart with their parents for their wintering grounds between September and November, and family units stay together through the winter until the following March, when the breeding cycle begins again (Tacha et al. 1992). In order to determine the precise age of death for the individual subadult cranes in the Watmough Bay assemblage, numerous comparative specimens of different known ages would be needed. In addition, since the legs of birds develop earlier than the wings (Tacha et al. 1992) it is possible that the leg bones of one individual may be classified as a juvenile and the wing bones of the same individual could be classified as a chick. However, given that those specimens labeled as chick are significantly smaller than adults (figures 2.3 and 2.4), it is very likely they were caught in the first few months of life, prior to fledging. It is not possible to know for sure, of course, where the people living at the Watmough Bay site hunted the immature cranes. Bird bones were sometimes transported significant distances by prehistoric human hunters (Parmalee 1977, 1980) and bird carcasses may drift ashore from distant places of origin (Schäfer 1972). However, there are a number of indicators that help determine whether or not a specific bird species was hunted locally, including carcass completeness, the overall species com-

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Kristine M. Bovy

position of the assemblage, the presence of eggshells, ethnographic accounts, and the availability of suitable breeding habitat. I consider each in turn. When animals are transported from far away, rather than obtained locally, they are more likely to have incomplete carcasses in the archaeological record (Lyman 1994). In particular, since sandhill cranes are large birds, we might expect differential transport of the meatier elements and/ or elements curated for tools or ornaments. Although the sandhill crane assemblage is small, it contains both wing (n = 5) and leg elements (n = 14). Thirteen out of the 16 subadult crane bones are leg elements; cranes, of course, have long, thin legs, which contain little meat. In addition, wing bones, rather than leg bones, are more often selected for bird bone tool or ornament manufacture (e.g., Morejohn and Galloway 1983). For example, of 17 modified bird bones at the Minard site at Grays Harbor, Washington, which included both awls and groove-­and-­snap pieces, all were wing elements (Bovy 2005). In short, the skeletal part representation of the sandhill cranes does not suggest long-­distance transport. The mammal remains from the 1968 excavation at Watmough Bay have not yet been analyzed, so it is not possible to comment on whether or not nonlocal mammals were found at the site. However, at least 60 other avian taxa are present in the Watmough Bay assemblage, including waterfowl, seabirds, shorebirds, and terrestrial birds (Bovy 2005), all of which are found in the San Juan Islands today. Many other immature bird specimens were recovered from Watmough Bay, especially cormorants (Phalacrocoracidae [n = 2366]), herons (Ardeidae [n = 107]), and gulls (Laridae [n = 51]), suggesting the people living at the site were highly adept at procuring young birds and were perhaps harvesting eggs. All skeletal elements (axial, pectoral, wings, legs) of these immature birds were recovered at the site, thus indicating they were hunted nearby and that complete carcasses were deposited in the midden (Bovy 2005, 2007b, 2011). The lack of nonlocal birds in the large avian assemblage, as well as skeletal part analysis of other immature taxa, suggest the cranes were also hunted locally. Although many immature birds were recovered from Watmough Bay and ethnographic and historic evidence suggests native peoples in the region collected eggs (e.g., Barnett 1955; Elmendorf 1960), no eggshells were recovered from Watmough Bay. This is not necessarily surprising, however, given that pieces of eggshell do not preserve as well as bone and are usually recovered only through wet screening with mesh sizes of 2 mm or less (Serjeantson 2009). If bird eggs were brought back to Watmough Bay, they were simply not recovered during the excavation. Cranes are mentioned briefly in the ethnographic literature in the Pacific Northwest, both as sources of food (Elmendorf 1960; Friedman



Zooarchaeological Evidence for Sandhill Crane Breeding

35

1976; Olson 1936; Underhill 1944) and for their mythological importance in some cultures (Jenness 1955). For example, the Twana of southern Puget Sound caught sandhill cranes at night while hunting deer along the shore (Elmendorf 1960:93). The crane was a guardian spirit for the Katzie First Nation of the lower Fraser River Valley, who referred to the month of March, when the cranes arrived to breed, as the “sandhill-­crane month” (Jenness 1955:9; Leach 1979). In The Natural History of Washington Territory, Cooper reported that young cranes “are often raised from the nest by the Indians for food” (Cooper and Suckley 1859:227), although he does not specify to which native groups he is referring. There is no mention of either raising young cranes for food or obtaining them from distant locations in any of the extensive ethnographies of the Straits Coast Salish (Barnett 1955; Stern 1934; Suttles 1951) and there is no archaeological evidence to suggest such practices occurred in the deeper past. Sandhill cranes have rarely been identified in archaeological sites in the Pacific Northwest (e.g., Butler and Campbell 2004; Friedman 1976; Hanson 1991). Small numbers of sandhill crane bones have been identified from the following archaeological sites: Minard (45-­GH-­15) at Grays Harbor, Washington (NISP = 3; Bovy 2005), Hoko River Rockshelter on the Olympic Peninsula of Washington (NISP = 2; Wigen and Stucki 1988), Shoemaker Bay (DhSe2) on western Vancouver Island (NISP = 1; Calvert and Crockford 1982), and Umpqua/Eden (35-­DO-­83) in the south-­central Oregon coast (NISP = 1; Bovy 2005). Sandhill cranes were also identified at the nearby Cattle Point site (45-­SJ-­1) on San Juan Island, though no comparative specimens were consulted for the analysis (Hanson 1991; King 1950) and thus these identifications must be viewed with caution. The Watmough Bay assemblage appears to be unique in terms of having remains of immature cranes, although the lack of such remains at other sites could be due partly to the difficulty in identifying juvenile birds. In terms of suitable breeding habitat of sandhill cranes, there are wetlands on Lopez Island today, including one adjacent to the site (figure 2.2). However, recent charcoal and pollen analysis indicates that the regional climate was warmer and drier when the site was occupied (Hallett et al. 2003; Lepofsky et al. 2005; Sugimura et al. 2008), so there may have been fewer freshwater marshes in the past. Further environmental reconstruction is necessary to determine suitable past breeding habitat. In sum, there is no evidence to suggest that the sandhill crane remains originated from a geographic location remote from Watmough Bay, and every reason to conclude that they originated near the site, say, within a few tens of kilometers. The 16 subadult sandhill crane bones from the Watmough Bay assemblage, therefore, provide strong evidence that the species bred in the San Juan Islands in the past.

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Kristine M. Bovy

The breeding population closest to Lopez Island today is in the Fraser Lowlands at Burns Bog and Pitt Polder, east of Vancouver, British Columbia (figure 2.1). While sandhill crane populations in British Columbia as a whole have increased in recent decades, the number of cranes breeding in the Fraser Lowlands remains small and vulnerable (Gebauer 2004). In 1945, both the Burns Bog and Pitt Polder had maintained about eight breeding pairs, but today there are only two to five pairs at each location (Cooper 1996; Marks et al. 2008). As elsewhere, loss of habitat has been one of the main factors for this decline, as well as human disturbance in this heavily developed region (Cooper 1996; Leach 1979). Plant succession in the bogs has quickened due to drainage and development, as well as cessation of burning by native peoples, which may be further reducing crane habitat (Leach 1979). In 1981, 17 greater sandhill cranes were released into the Fraser Lowland in an effort to boost their populations, but the success of this project is unknown (Cooper 1996; Ivey et al. 2005). It is clear that the sandhill cranes breeding in the vicinity of the Fraser Lowlands are a remnant population; historical accounts indicate that local breeding populations have been extirpated from Lulu Island and nearby Vancouver Island (Cooper 1996; but see Cooper 2006 for a discussion of the recent “rediscovery” of small numbers of breeding pairs in remote locations on northern Vancouver Island). The immature birds recovered from the Watmough Bay site were likely part of this once more-­ widespread Fraser Lowlands population. The subspecies status of the breeding birds in the Fraser Lowlands and those occurring elsewhere in British Columbia is unclear (Cooper 1996). This confusion is due in part to the fact that the Canadian subspecies (G. c. rowani), which is intermediate in size between the greater and lesser sandhill crane, was not designated a subspecies until the mid-­1960s and is therefore absent from earlier accounts (Ivey et al. 2005). Recent research and observations, however, indicate that the Fraser Valley population may be the Canadian subspecies (Ivey et al. 2005). Ivey et al. (2005) captured and tracked eight sandhill cranes from a staging area located along the lower Columbia River in southwestern Washington and northwestern Oregon in order to better understand the subspecies composition and migration of cranes in the Pacific flyway. The birds in their study migrated west of the Cascades and bred in southern Alaska and coastal British Columbia. On the basis of these migration patterns, as well as physical characteristics, Ivey et al. (2005) conclude that the birds are distinct from the greater sandhill crane breeding in eastern Washington and are likely the Canadian subspecies. Ivey et al. (2005) also argue that these birds are associated with the breeding population



Zooarchaeological Evidence for Sandhill Crane Breeding

37

in the Fraser Lowlands and that G. c. rowani may have once bred from Alaska to Oregon. The immature specimens from Watmough Bay align with this latter assertion. The possibility that G. c. rowani in fact bred across much of the Pacific Northwest could be tested with metric studies or ancient DNA analysis of prehistoric remains to confirm subspecies affiliation.

Management Implications Data presented here on prehistoric sandhill crane rookeries are similar to those from other studies that document past ranges of locally extirpated species (Etnier 2004; Lyman 1988; Nagaoka et al. 2008; Randklev et al. 2010; Smith 1985; Valentine et al. 2008). The key to management of sandhill cranes is habitat availability; shallow marshes and wet meadows for feeding and nesting are essential and must be protected (Cooper 1996; Littlefield and Ivey 2002; Tacha et al. 1992). Currently, conservation efforts for sandhill cranes are focused on those breeding populations in south-­central Washington (Littlefield and Ivey 2002). However, Ivey et al. (2005) argue that the Canadian subspecies should be managed separately from the greater, given their different annual migrations. Specifically, they suggest that the lower Columbia River provides an important staging area for G. c. rowani and this area should also be protected (Ivey et al. 2005). If the immature Watmough Bay specimens are also the Canadian subspecies, this may point to another region of potential importance for habitat conservation and sandhill crane management. A broader distribution of breeding cranes in the state is necessary to avoid catastrophic losses due to natural disasters and other forms of disturbance (Littlefield and Ivey 2002). A more complete understanding of the historical breeding distribution of sandhill cranes should help inform management of the species in Washington State. Bird remains from the Watmough Bay site include the remains of multiple immature sandhill cranes. These remains provide a much greater time depth (about 1500 years) than historical accounts for the breeding loci of sandhill cranes and are the first probable evidence for historical breeding by the species in northwestern Washington State. The presence of chicks and subadults indicates that rookeries occurred in the area in the prehistoric late Holocene. This finding clarifies the historical breeding distribution of the sandhill crane subspecies and suggests another possible focus for conservation efforts. In particular, these data suggest that restoration of breeding populations of sandhill cranes in the northwestern portion of Washington is warranted.

38

Kristine M. Bovy Acknowledgments

The research has been supported by a grant from the US Environmental Protection Agency’s Science to Achieve Results (STAR) program (grant #U-­91576301); it has not been subject to EPA review, however, and therefore does not necessarily reflect the view of the agency. The radiocarbon dates were funded by a National Science Foundation Dissertation Improvement Grant (#BCS-­ 0242632). Thanks to Burke Museum archaeology staff (Peter Lape, Laura Phillips, Megon Noble) for permission to analyze the Watmough Bay assemblage, obtain radiocarbon dates and photograph the bones, and for help facilitating the loan. Rob Faucett, Sievert Rohwer, and Chris Wood generously provided access to bird skeleton specimens at the Burke Museum. Stephanie Jolivette (Burke Museum) deserves special thanks for photographing the archaeological and comparative specimens for me. Valuable advice was also provided by Don Grayson, Steve Wolverton, and Lee Lyman. References Barnett, H. G. 1955. The Coast Salish Indians of British Columbia. University of Oregon Press, Eugene. Bochenski, Z. M. 2008. Identification of Skeletal Remains of Closely Related Species: The Pitfalls and Solutions. Journal of Archaeological Science 35: 1247–1250. Bovy, K. M. 2005. Effects of Human Hunting, Climate Change and Tectonic Events on Waterbirds along the Pacific Northwest Coast during the Late Holocene. Unpublished PhD dissertation, Department of Anthropology, University of Washington, Seattle. ———. 2007a. Global Human Impacts or Climate Change? Explaining the Sooty Shearwater Decline at the Minard Site, Washington, USA. Journal of Archaeological Science 34:1087–1097. ———. 2007b. Prehistoric Human Impacts on Waterbirds at Watmough Bay, Washington, USA. Journal of Island and Coastal Archaeology 2:210–230. ———. 2011. Archaeological Evidence for a Double-­Crested Cormorant (Phalacrocorax auritus) Colony in the Pacific Northwest, USA. Waterbirds 34:89–95. Broughton, J. M. 2004. Prehistoric Human Impacts on California Birds: Evidence from the Emeryville Shellmound Avifauna. Ornithological Monographs No. 56, American Ornithologists’ Union, Washington, DC. Butler, V. L., and S. K. Campbell. 2004. Resource Intensification and Resource Depression in the Pacific Northwest of North America: A Zooarchaeological Review. Journal of World Prehistory 18:327–405. Calvert, G., and S. Crockford. 1982. Appendix IV: Analysis of Faunal Remains from the Shoemaker Bay Site (DhSe 2). In Alberni Prehistory: Archaeological



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and Ethnographic Investigations on Western Vancouver Island, edited by A. D. McMillan and D. E. St. Claire, pp. 174–219. Theytus Books, Penticton, BC. Cooper, J. G., and G. Suckley. 1859. The Natural History of Washington Territory. Ballière Brothers, New York. Cooper, J. M. 1996. Status of the Sandhill Crane in British Columbia. Wildlife Bulletin No. B-­83. Ministry of Environment, Lands and Parks, Wildlife Branch, Victoria, BC. ———. 2006. Sandhill Cranes Breeding on Northern Vancouver Island, British Columbia. Northwestern Naturalist 87:146–149. Deo, J. N., J. O. Stone, and J. K. Stein. 2004. Building Confidence in Shell: Variations in the Marine Radiocarbon Reservoir Correction for the Northwest Coast over the Past 3,000 Years. American Antiquity 69:771–786. Driver, J. C. 1992. Identification, Classification and Zooarchaeology. Circaea 9:35–47. Elmendorf, W. W. 1960. The Structure of Twana Culture. Research Studies: A Quarterly Publication of Washington State University (Monograph Supplement No. 2) 28:1–576. Etnier, M. A. 2004. The Potential of Zooarchaeological Data to Guide Pinniped Management Decisions in the Eastern North Pacific. In Zooarchaeology and Conservation Biology, edited by R. L. Lyman and K. P. Cannon, pp. 88–102. University of Utah Press, Salt Lake City. Friedman, E. 1976. An Archaeological Survey of Makah Territory: A Study in Resource Utilization. Unpublished PhD dissertation, Department of Anthropology, Washington State University, Pullman. Gebauer, M. 2004. Sandhill Crane (Grus canadensis). In Accounts and Measures for Managing Identified Wildlife (Version 2004). Province of British Columbia, Ministry of Water, Land and Air Protection, Victoria, BC. Grayson, D. K. 1984. Quantitative Zooarchaeology: Topics in the Analysis of Archaeological Faunas. Academic Press, Orlando, FL. Hallett, D. J., D. S. Lepofsky, R. W. Mathewes, and K. P. Lertzman. 2003. 11000 Years of Fire History and Climate in the Mountain Hemlock Rain Forests of Southwestern British Columbia Based on Sedimentary Charcoal. Canadian Journal of Forest Research 33:292–313. Hanson, D. K. 1991. Late Prehistoric Subsistence in the Strait of Georgia Region of the Northwest Coast. Unpublished PhD dissertation, Department of Archaeology, Simon Fraser University, Burnaby, BC. Ivey, G. L., C. P. Herziger, and T. J. Hoffmann. 2005. Annual Movements of Pacific Coast Sandhill Cranes. Proceedings of the North American Crane Workshop 9:25–35. Jenness, D. 1955. The Faith of a Coast Salish Indian. Anthropology in British Columbia. Memoir No. 3. Provincial Museum, Victoria, BC. Jewett, S. A., W. P. Taylor, W. T. Shaw, and J. W. Aldrich. 1953. Birds of Washington State. University of Washington Press, Seattle.

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King, A. R. 1950. Cattle Point: A Stratified Site in the Southern Northwest Coast Region. Memoirs of the Society for American Archaeology No. 7. Leach, B. A. 1979. The Sandhill Crane in the Lower Fraser Valley. Information Booklet No. 31. Institute of Environmental Studies, Douglas College, New Westminster, BC. Lepofsky, D., K. Lertzman, D. Hallett, and R. Mathewes. 2005. Climate Change and Culture Change on the Southern Coast of British Columbia 2400-­1200 Cal. B.P.: A Hypothesis. American Antiquity 70:267–293. Lewis, M. G., and F. A. Sharpe. 1987. Birding in the San Juan Islands. The Mountaineers, Seattle. Littlefield, C. D., and G. L. Ivey. 2002. Washington State Recovery Plan for the Sandhill Crane. Washington Department of Fish and Wildlife, Olympia. Lyman, R. L. 1988. Zoogeography of Oregon Coast Mammals: The Last 3000 Years. Marine Mammal Science 4:247–264. ———. 1994. Vertebrate Taphonomy. Cambridge University Press, Cambridge. ———. 2008. Quantitative Paleozoology. Cambridge University Press, Cambridge. ———. 2010. Paleozoology’s Dependence on Natural History Collections. Journal of Ethnobiology 30:126–136. Marks, D. J., A. Teucher, and L. Ramsay. 2008. Conservation Status Report: Grus canadensis. BC Conservation Data Centre. Electronic document, http://a100 .gov.bc.ca/pub/eswp/esr.do?id=18288, accessed July 7, 2010. Marshall, F., and T. Pilgram. 1993. NISP vs. MNI in Quantification of Body-­Part Representation. American Antiquity 58:261–269. Morejohn, G. V., and J. P. Galloway. 1983. Identification of Avian and Mammalian Species Used in the Manufacture of Bone Whistles Recovered from a San Francisco Bay Area Archaeological Site. Journal of California and Great Basin Anthropology 5:87–97. Nagaoka, L., S. Wolverton, and B. Fullerton. 2008. Taphonomic Analysis of the Twilight Beach Seals. In Islands of Inquiry: Colonisation, Seafaring, and the Archaeology of Maritime Landscapes, edited by G. Clark, F. Leach, and S. O’Connor, pp. 475–498. Terra Australis 29, Australian National University Press, Canberra. Olson, R. L. 1936. The Quinault Indians. University of Washington Publications in Anthropology VI(1), Seattle. Parmalee, P. W. 1977. The Avifauna from Prehistoric Arikara Sites in South Dakota. Plains Anthropologist 22:189–222. ———. 1980. Utilization of Birds by the Archaic and Fremont Cultural Groups of Utah. Los Angeles County Museum Contributions in Science 333:237–250. Randklev, C. R., S. Wolverton, B. Lundeen, and J. H. Kennedy. 2010. A Paleozoological Perspective on Unionid (Mollusca: Unionidae) Zoogeography in the Upper Trinity River Basin, Texas. Ecological Applications 20:2359–2368.



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Schäfer, W. 1972. Ecology and Palaeoecology of Marine Environments. University of Chicago Press, Chicago. Serjeantson, D. 2009. Birds. Cambridge University Press, Cambridge. Smith, I. W. G. 1985. Sea Mammal Hunting and Prehistoric Subsistence in New Zealand. Unpublished PhD dissertation, University of Otago, Dunedin, NZ. Stein, J. K., J. N. Deo, and L. S. Phillips. 2003. Big Sites—Short Time: Accumulation Rates in Archaeological Sites. Journal of Archaeological Science 30:297– 316. Stern, B. J. 1934. The Lummi Indians of Northwest Washington. Columbia University Press, New York. Stewart, J. R. 2004. Wetland Birds in the Recent Fossil Record of Britain and Northwest Europe. British Birds 97:33–43. ———. 2005. The Use of Modern Geographical Ranges in the Identification of Archaeological Bird Remains. In Feathers, Grit, and Symbolisms: Birds and Archaeology in the Old and New Worlds, edited by G. Grupe and J. Peters, pp. 43–54. Proceedings of the ICAZ Bird Remains Working Group Meeting in Munich 2004. Documenta Archaeobiologiae 3. Rahden, Westfalen. Sugimura, W. Y., D. G. Sprugel, L. B. Brubaker, and P. E. Higuera. 2008. Millennial-­Scale Changes in Local Vegetation and Fire Regimes on Mt. Constitution, Orcas Island, Washington, USA, Using Small Hollow Sediments. Canadian Journal of Forest Research 38:539–552. Suttles, W. P. 1951. Economic Life of the Coast Salish of Haro and Rosario Straits. Unpublished PhD dissertation, Department of Anthropology, University of Washington, Seattle. Tacha, T. C., S. A. Nesbitt, and P. A. Vohs. 1992. Sandhill Crane. In The Birds of North America No. 31, edited by A. Poole, P. Stettenheim, and F. Gill. Academy of Natural Sciences, Philadelphia, American Ornithologists’ Union, Washington, DC. Underhill, R. M. 1944. Indians of the Pacific Northwest. Education Division of the US Office of Indian Affairs, Washington, DC. Valentine, K., D. A. Duffield, L. E. Patrick, D. R. Hatch, V. L. Butler, R. L. Hall, and N. Lehman. 2008. Ancient DNA Reveals Genotypic Relationships among Oregon Populations of the Sea Otter (Enhydra lutris). Conservation Genetics 9:933–938. Washington Department of Fish and Wildlife (WDFW). 2010. Species of Concern. http://wdfw.wa.gov/wildlife/management/endangered.html, accessed July 8, 2010. Wigen, R. J., and B. R. Stucki. 1988. Taphonomy and Stratigraphy in the Interpretation of Economic Patterns at Hoko River Rockshelter. In Prehistoric Economies of the Pacific Northwest Coast, edited by B. L. Isaac, pp. 87–146. Research in Economic Anthropology Supplement 3. JAI Press, Greenwich, CT.

CHAPTER THREE

Archaeological Freshwater Mussel Remains and Their Use in the Conservation of an Imperiled Fauna Evan Peacock

Freshwater mussels (Bivalvia: Unionoida) are one of the most endangered faunas in the world. As sessile gill breathers and filter feeders, they are sensitive to siltation and other forms of water pollution. Their life cycle includes a parasitic stage on host fishes; larvae (glochidia) attach to fish for a period of metamorphosis, after which juvenile mussels drop to the substrate where they are vulnerable to smothering due to their small size. A downside to the parasite stage is that anything impacting host fish automatically affects mussel species as well. A plethora of impacts has altered mussel populations in North America, where “202 of the nearly 300 unionid species known are . . . presumed extinct, possibly extinct, critically imperiled, imperiled, or vulnerable. . . . In the United States alone, 37 species are presumed extinct or possibly extinct” (Lydeard et al. 2004:325). The Unionoida have one of the highest extinction rates of any group of organisms over the last century (Bogan 1993, 2006, 2008; Haag 2009a; Neves et al. 1997; Strayer et al. 2004; Williams et al. 1993). As some species have become extirpated, others have expanded into water bodies altered in such a way as to favor

42



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their (or their hosts’) habitat requirements (Haag and Warren 1998; Starnes and Bogan 1988; Strayer et al. 2004). Such changes have been taking place since the inception of Euro-­ American settlement (Haag 2009a; Hughes and Parmalee 1999) and perhaps earlier (Peacock et al. 2005). Recording of North American unionids by natural scientists began in the nineteenth century (Bogan 1998) when extirpations related to anthropogenic impacts already were common (Bogan 2006). Quantitative data accurately reflecting taxonomic abundances were not forthcoming until the early twentieth century (Bogan 1998, 2006). Many watercourses, especially smaller ones, have not yet been surveyed (Backlund 2000; Theler 1990), and many reports provide only presence or absence data. Thus, knowledge of biogeography and ecology is limited, based as it is upon an incomplete historical record that may not represent America’s waterways prior to modern impacts even though such data are “critical to understanding the current status of many mussel populations” (National Native Mussel Conservation Committee 1998:1422). Although a call for use of long-­term data in mussel conservation has repeatedly been made (e.g., Bogan 1993; Cummings and Bogan 2006; Lydeard et al. 2004; Strayer 2008), recommendations are limited to consultation of museum collections or survey data provided by early naturalists. Such collections can be biased by the aims or abilities of individuals dealing with what were largely unknown faunas (Hughes and Parmalee 1999), and such deficiencies are not always reparable due to collection loss, poor cataloging, or other curation problems (Haag 2009b; Hoke 2000). Archaeological shell assemblages (figure 3.1) offer a solution, as shell-­ bearing sites are common on waterways large and small, and large collections relative to modern ones have been, or may easily be, obtained (Haag 2009a). For example, “the original [mussel] fauna prior to major twentieth century stream degradation is poorly known” for the Des Moines and Monongahela rivers (Haag 2009b:4). Yet of about 1,850 sites in the Monongahela subbasin, 893 are attributable to specific prehistoric time periods (Pennsylvania Historical and Museum Commission 2009), and many of these sites contain shell. There is no reason why the freshwater molluscan fauna should remain “poorly known.” Starnes and Bogan (1988:table 4) tabulate 42 cases from the middle and lower Tennessee River where species are represented only by zooarchaeological records, while Lyons et al. (2007:9–10) note 10 mussel species from prehistoric middens along the Black River, Ohio, which were not found in modern surveys. Faunas can be considered more complete if zooarchaeological data are incorporated because modern streams have been impacted to unknown degrees. Because shell remains are common in archaeological

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Evan Peacock

Figure 3.1. Left valves of Obliquaria reflexa (exterior) and right valves of Plectomerus dombeyanus (interior) showing range of sizes recovered from the Kinlock site (22SU526), Sunflower County, Mississippi. Scale in centimeters.



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deposits, they can indicate baseline stream conditions prior to modernity (Hughes and Parmalee 1999). Zoologists have identified archaeological molluscan remains for decades (Baker 1923, 1979; Call 1992; Call and Robinson 1983; Hartfield 1986, 1990; Matteson 1953, 1958, 1959, 1960; Morrison 1942; Murray 1981; Ortmann 1909; Stansbery 1966a, 1966b, 1974). Work on species ranges as determined from zooarchaeological specimens has occasionally appeared in the biological literature (Barber 1982; Gordon 1983). A familiar use of zooarchaeological data has been to emphasize the loss of diversity in modern times (Neves et al. 1997; Taylor and Spurlock 1982). However, such endeavors have occupied the fringes of malacological work in North America. Fortunately, due largely to the efforts of a few archaeologists-­cum-­ malacologists, the applied research value of archaeological shell is being realized. Studies have grown in number and sophistication (see the references cited at the end of this chapter), with many cross-­disciplinary collaborations. The use of zooarchaeological data in mussel conservation cannot yet be called mainstream, however (e.g., the lack of mention in an overview by Strayer [2008]), for two reasons. The first is a widespread belief that archaeological assemblages are too biased by prehistoric human collection preferences—for example, for particular sizes or species— to provide representative samples (see figure 3.1). The second involves preservation biases, but these are often overstated or can be compensated for (Peacock 2000). The difference between the death assemblages studied by zooarchaeologists and the life assemblages studied by zoologists is often perceived as too great to bridge, but this perception reveals a lack of understanding of how zooarchaeology works. There is no valid reason why archaeological shell should not be recognized as important in management of freshwater molluscan faunas. A detailed defense of the merits of zooarchaeological data is beyond the scope of this chapter. Suffice it to say that cultural influences, such as long-­distance transport of shell, as well as taphonomic effects that influence shell preservation would be examined as hypotheses by the archaeologist and would not merely be assumed to be biases.

Goals of Applied Archaeomalacology Application of zooarchaeological data to conservation issues depends on choosing measured variables appropriate for the chosen target conditions (Lyman 2008:11–16). For many conservation purposes, nominal-­scale (presence/absence) data are appropriate (e.g., for the generation of range

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maps). Many neoecological studies of mussels provide only lists of species represented at sampling stations (e.g., McGregor and Haag 2004). In other cases, ordinal-­scale relative frequencies of taxa may be the appropriate measured variable, as when the target is to track the status of endangered species over time (e.g., Jenkinson and Ahlstedt 1987). Many species of freshwater mussel are extinct, so “reconstructing” complete communities is in many cases impossible. Zooarchaeological data indicate that exceptions exist. Relatively little change has taken place in lower Gulf coastal plain faunas, for example, presumably due to adaptation to naturally turbid conditions and hence less sensitivity to the accelerated erosion of historic times (Peacock and Mistak 2008). At present, what matters is the preservation of taxonomic diversity; genomic diversity may soon also matter. Maintaining sufficient habitat is the surest way of meeting these goals, but captive breeding, habitat reclamation, species reintroductions, and radical steps such as cloning may be required (Dunn and Layzer 1997). Biologists and archaeomalacologists alike argue that reintroduction of species into former habitats should be undertaken (e.g., Bogan 1998; Haag 2009a; Humphries and Winemiller 2009), but attempts to do so have met with varying results (e.g., Dunn and Sietman 1997; Havlik 1997; Morgan et al. 1997). To determine which species should be transplanted to a location, species lists compiled from zooarchaeological data are a primary resource. The choice of benchmarks for ecological reconstruction can be tricky (see chapter 1, this volume). But it can be an informed choice, determined by the goals of the researcher and the quality of the data available for characterizing the desired target state.

A Case Study from the Tombigbee River Drainage The Tombigbee River drains large portions of eastern Mississippi and western Alabama in the southeastern United States (figure 3.2). Although modern surveys have been carried out in tributary streams (e.g., Jones 1991; Jones and Majure 1999; Jones et al. 1996; Miller 2001), little systematic work was done in the main river channel prior to completion of the Tennessee-­Tombigbee Waterway in 1985. Based on available modern (historic-­era) data, canalization and dam construction severely impacted main-­channel mussel populations (Hartfield and Jones 1989a, 1989b; McGregor 2000; Miller and Hartfield 1988; van der Schalie 1939; Williams et al. 1992; Williams et al. 2008). Some tributaries have diverse mussel populations (e.g., the Sipsey River [McCullagh et al. 2002]), while others have become impoverished (e.g., Tibbee Creek [Jones et al. 1997]; see also Haag and Warren [1998] for discussion). Without re-



Mussel Remains’ Use in Conservation of Imperiled Fauna

47

Figure 3.2. Tombigbee River drainage, Mississippi and Alabama, with archaeological sites from which quantified data on freshwater mussel shell are available.

course to zooarchaeological data, the cause of spatial variability in mussel communities may be assumed to be historic-­period impacts, either to the mussels themselves and/or to their host fish populations. While such assumptions are warranted, the pre-­modern-­impact fauna of the Tombigbee River basin cannot be determined using modern data. Due to changing environmental conditions and different sampling strategies, results from modern surveys in the same streams can differ over as little as a decade (Hartfield and Jones 1990). As noted above, many modern studies give only presence or absence data for sampling stations, and sample sizes vary. These factors underscore the need for zooarchaeological data, which in many cases are more robust for addressing conservation goals (e.g., knowledge of species ranges and pre-­modern-­impact community ecology) than are modern data. A great deal of archaeology was done in conjunction with the Tenn-­ Tom Waterway and other “improvement” projects, and many large assemblages of archaeological shell have been analyzed (Peacock 1998; Robison 1983). Small samples from sites on tributaries also have begun to be analyzed. Data on 71,074 identifiable valves from 21 sites in Mississippi

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and Alabama are presented in table 3.1. With the exception of the assemblage from the Vaughn Mound site (22LO538 [Peacock and Seltzer 2008]), which was recovered without screening, all shell was recovered using quarter-­inch mesh at largest.1 The assemblages represent shell obtained from a variety of contexts (Peacock 1998; Peacock et al. 2011). Zooarchaeological sample sizes vary, from single specimens at interior sites to tens of thousands of shells from those along larger waterways (table 3.1). This reflects the scale of archaeological work undertaken, the size of mussel beds that existed in the past, and the intensity and duration of exploitation of those beds. A way to assess whether or not species represent past community composition is to analyze whether or not taxonomic composition of faunal assemblages varies in clinal fashion along river courses. While this does not unequivocally address the issue of sample representativeness, samples that represent local biological signatures should partition locationally in ways that reflect actual stream geography. Ordination via correspondence analysis was carried out using PC-­ORD (McCune and Mefford 1997). An arbitrary requirement of 100 identifiable valves was used to minimize noise related to sample size differences. Detrending follows Peacock (1998, 2002), employing a chi-­square-­based distance measurement. No downweighting of rare species or other manipulation of the data was done. Geography is the dominant structuring factor in the distribution of the assemblages (figures 3.2 and 3.3). Those that exhibit similar taxonomic composition tend to be located close to one another, which is expected if taxonomic composition varies along a cline. Similarly, the assemblages from Alabama separate as being more positive along Axis 2; these are downstream from the Mississippi sites, with the most positive assemblage, 1CK56, being located on the lower Tombigbee River in Clarke County, Alabama (Peacock 2009). There is one exception to the clinal pattern, and it represents an anomaly in terms of taxonomic composition and geographic location. Site 22OK520, the Lyon’s Bluff site, is strongly positive along Axis 1; this site is situated on Line Creek, a tributary of Tibbee Creek, itself a tributary of the Tombigbee River. The other assemblages from Mississippi are located on or near the main stem of the Tombigbee River within a few kilometers of one another. A species-­area curve (e.g., Lyman and Ames 2007) was produced for the 11 sites with more than 100 identified valves (figure 3.4). That curve indicates that the Lyon’s Bluff molluscan fauna has more taxa for the number of identified valves it produced than the other 10 assemblages; those other 10 assemblages all have considerably more identified valves but only the three with an order of magnitude more identified valves than Lyon’s Bluff have more identified taxa than it does. The Lyon’s Bluff assemblage represents a different habitat and does not belong in



Mussel Remains’ Use in Conservation of Imperiled Fauna

49

Table 3.1. Mussel Data from the Tombigbee River Drainage Used in Correspondence Analysis Species

22CS503 22LE504 22LE827 22OK520

Amblema plicata Arcidens confragosus Ellipsaria lineolata Elliptio arca Elliptio arctata Elliptio crassidens Epioblasma penita Fusconaia cerina Fusconaia ebena Glebula rotundata Hamiota perovalis Lampsilis ornata Lampsilis straminea claibornensis Lampsilis straminea straminea Lampsilis teres Lampsilis complanata alabamensis Leptodea fragilis Ligumia recta Medionidus acutissimus Megalonaias nervosa Obliquaria reflexa Obovaria jacksoniana/unicolor Plectomerus dombeyanus Pleurobema decisum Pleurobema marshalli Pleurobema perovatum Pleurobema taitianum Potamilus purpuratus Quadrula apiculata/rumphiana Quadrula asperata Quadrula metanevra Quadrula nobilis Quadrula stapes Quadrula verrucosa Rangia cuneata Strophitus subvexus Toxolasma parvum Toxolasma texasensis Truncilla donaciformis Truncilla truncata Uniomerus declivus Uniomerus tetralasmus Villosa lienosa Villosa vibex

2 0 0 0 0 14 0 43 0 0 0 0 0 0 0 0 0 0 0 0 0 16 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

1 0 0 0 0 0 0 1 0 0 0 0 0 4 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0

6 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Total valves:

75

7

6

124 0 0 7 0 37 0 180 0 0 1 4 1 302 3 1 0 2 6 6 1 49 0 3 0 42 0 1 4 1 0 0 0 5 0 10 54 0 0 0 0 2 14 23 883 (continued)

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Evan Peacock

Table 3.1. Continued Species

22OK534 22OK578 22OK595 22OK793

Amblema plicata Arcidens confragosus Ellipsaria lineolata Elliptio arca Elliptio arctata Elliptio crassidens Epioblasma penita Fusconaia cerina Fusconaia ebena Glebula rotundata Hamiota perovalis Lampsilis complanata alabamensis Lampsilis ornata Lampsilis straminea claibornensis Lampsilis straminea straminea Lampsilis teres Leptodea fragilis Ligumia recta Medionidus acutissimus Megalonaias nervosa Obliquaria reflexa Obovaria jacksoniana/unicolor Plectomerus dombeyanus Pleurobema decisum Pleurobema marshalli Pleurobema perovatum Pleurobema taitianum Potamilus purpuratus Quadrula apiculata/rumphiana Quadrula asperata Quadrula metanevra Quadrula nobilis Quadrula stapes Quadrula verrucosa Rangia cuneata Strophitus subvexus Toxolasma parvum Toxolasma texasensis Truncilla donaciformis Truncilla truncata Uniomerus declivus Uniomerus tetralasmus Villosa lienosa Villosa vibex

0 0 0 0 0 1 0 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0

0 0 0 1 0 17 1 1 0 0 0 0 1 1 0 2 0 0 0 0 0 0 0 3 0 0 0 0 0 1 0 0 0 2 0 0 0 1 0 0 0 0 0 0

0 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Total valves:

3

31

1

0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 2 (continued)



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Table 3.1. Continued Species

22OK904 22OK905 22OK912 22CL527

Amblema plicata Arcidens confragosus Ellipsaria lineolata Elliptio arca Elliptio arctata Elliptio crassidens Epioblasma penita Fusconaia cerina Fusconaia ebena Glebula rotundata Hamiota perovalis Lampsilis complanata alabamensis Lampsilis ornata Lampsilis straminea claibornensis Lampsilis straminea straminea Lampsilis teres Leptodea fragilis Ligumia recta Medionidus acutissimus Megalonaias nervosa Obliquaria reflexa Obovaria jacksoniana/unicolor Plectomerus dombeyanus Pleurobema decisum Pleurobema marshalli Pleurobema perovatum Pleurobema taitianum Potamilus purpuratus Quadrula apiculata/rumphiana Quadrula asperata Quadrula metanevra Quadrula nobilis Quadrula stapes Quadrula verrucosa Rangia cuneata Strophitus subvexus Toxolasma parvum Toxolasma texasensis Truncilla donaciformis Truncilla truncata Uniomerus declivus Uniomerus tetralasmus Villosa lienosa Villosa vibex

0 0 0 0 0 1 0 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

0 0 0 0 0 1 0 0 0 0 0 0 0 0 2 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0 0

0 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0

Total valves:

3

4

2

61 0 5 192 2 270 179 62 37 0 0 0 10 11 0 0 0 0 0 0 19 27 0 1571 0 32 0 0 4 369 0 0 11 13 0 0 0 0 1 0 0 0 15 0 2891 (continued)

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Evan Peacock

Table 3.1. Continued Species

22CL814 22CL917 22LO530 22LO538

Amblema plicata Arcidens confragosus Ellipsaria lineolata Elliptio arca Elliptio arctata Elliptio crassidens Epioblasma penita Fusconaia cerina Fusconaia ebena Glebula rotundata Hamiota perovalis Lampsilis complanata alabamensis Lampsilis ornata Lampsilis straminea claibornensis Lampsilis straminea straminea Lampsilis teres Leptodea fragilis Ligumia recta Medionidus acutissimus Megalonaias nervosa Obliquaria reflexa Obovaria jacksoniana/unicolor Plectomerus dombeyanus Pleurobema decisum Pleurobema marshalli Pleurobema perovatum Pleurobema taitianum Potamilus purpuratus Quadrula apiculata/rumphiana Quadrula asperata Quadrula metanevra Quadrula nobilis Quadrula stapes Quadrula verrucosa Rangia cuneata Strophitus subvexus Toxolasma parvum Toxolasma texasensis Truncilla donaciformis Truncilla truncata Uniomerus declivus Uniomerus tetralasmus Villosa lienosa Villosa vibex

562 0 11 250 11 586 62 138 22 0 0 0 10 64 0 29 1 11 0 7 38 138 0 621 0 0 0 9 39 65 3 0 11 75 0 1 0 0 5 0 0 0 0 0

192 0 36 982 0 2314 426 247 152 0 0 0 6 37 0 0 0 6 0 7 34 53 1 3181 0 8 9 0 53 1262 8 0 1 31 0 0 0 0 9 0 1 1 0 0

159 0 133 939 0 1421 1991 572 475 0 13 0 250 26 0 3 1 4 0 1 284 703 0 11473 0 84 191 10 90 2362 41 0 157 118 0 2 0 0 35 0 0 0 0 0

82 0 20 21 0 26 27 16 147 0 4 0 2 5 0 11 0 1 0 1 18 52 2 421 0 50 46 17 17 150 0 0 7 13 0 0 0 0 0 0 0 0 0 0

Total valves:

2769

9057

21538

1156 (continued)



Mussel Remains’ Use in Conservation of Imperiled Fauna

53

Table 3.1. Continued Species

22LO600

1CK56

Lubbub

1GR1X1

1GR2

Amblema plicata Arcidens confragosus Ellipsaria lineolata Elliptio arca Elliptio arctata Elliptio crassidens Epioblasma penita Fusconaia cerina Fusconaia ebena Glebula rotundata Hamiota perovalis Lampsilis complanata alabamensis Lampsilis ornata Lampsilis straminea claibornensis Lampsilis straminea straminea Lampsilis teres Leptodea fragilis Ligumia recta Medionidus acutissimus Megalonaias nervosa Obliquaria reflexa Obovaria jacksoniana/unicolor Plectomerus dombeyanus Pleurobema decisum Pleurobema marshalli Pleurobema perovatum Pleurobema taitianum Potamilus purpuratus Quadrula apiculata/rumphiana Quadrula asperata Quadrula metanevra Quadrula nobilis Quadrula stapes Quadrula verrucosa Rangia cuneata Strophitus subvexus Toxolasma parvum Toxolasma texasensis Truncilla donaciformis Truncilla truncata Uniomerus declivus Uniomerus tetralasmus Villosa lienosa Villosa vibex

207 0 60 821 0 1184 397 245 218 0 8 0 101 6 0 0 3 13 0 5 108 220 0 3947 2 93 73 4 29 704 7 0 38 41 0 1 4 0 17 0 0 0 1 0

892 8 77 53 3 73 68 344 7021 482 3 0 5 647 0 49 0 5 0 16 358 452 1007 2019 52 109 40 78 353 2190 22 54 21 9 6 0 1 0 0 0 0 0 27 9

140 0 62 144 1 847 30 160 517 0 0 0 56 26 0 0 0 4 0 3 95 319 1 1518 0 0 0 9 13 894 116 0 0 23 0 0 0 0 0 0 0 0 2 0

23 0 49 55 0 61 0 4 640 0 3 0 2 17 0 8 0 2 0 0 30 32 0 15 0 0 0 5 6 484 50 0 0 1 0 0 0 0 2 0 0 0 0 0

8 0 19 12 0 54 0 3 559 0 0 0 6 5 0 2 0 0 0 0 42 29 0 98 0 0 3 4 4 200 18 0 0 1 0 0 0 0 0 0 0 0 0 0

Total valves:

8557

16,553

4980

1489

1067

Mussels not identifiable to species not included. Pleurobema curtum has been subsumed under P. decisum, and Q. rumphiana complex has been subsumed under Q. apiculata. For original data, descriptions of contexts, sampling and recovery methods, taxonomic revisions, and number of unidentifiable valves, see Hanley (1982), Hartfield (1990), Peacock (1998, 2009), Peacock and Seltzer (2008), Peacock et al. (2011), Reitz (1987), Rummel (1980), and Woodrick (1981, 1983).

54

Figure 3.3. Ordination diagram showing results of detrended correspondence analysis of mussel data from Tombigbee River basin sites.



Mussel Remains’ Use in Conservation of Imperiled Fauna

55

Figure 3.4.  Species-area curve of mussel data from 11 sites on the Tombigbee River. Simple best-fit regression line (solid) and 95% confidence interval (dashed lines) shown for reference. Data from table 3.1.

the same cline as do the other sites. Correspondence analysis in this case has produced a pattern that highlights the difference of the Lyon’s Bluff assemblage, a difference that is unlikely to be the result of prey choice and does not appear to be the result of taphonomic bias. Archaeological shell data from the Tombigbee River drainage can be used to address target variables in mussel conservation because samples are large, are time/space averaged (Peacock 1998), and appear to represent past community composition. Using such data, range maps for species can be produced using nominal-­scale (presence/absence) data, providing clear targets for reintroductions (Peacock et al. 2011). For some of the larger samples, ordinal-­scale data (taxonomic abundances) can be considered representative of mussel communities as they existed prior to modern impacts, providing realistic targets for the management of endangered species. Zooarchaeological assemblages from more sites, especially on more tributary streams, would reveal the structure of pre-­ modern-­ impact mussel communities and changes related to modern environmental impacts at the drainage-­basin scale, providing information useful for habit reconstruction. Funding was obtained from the US Fish and Wildlife Service via the Mississippi Museum of Natural Science to produce range maps for the state using archaeological data (Peacock et al. 2011). This was done by compiling data from archaeological excavation reports and publications

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and by analyzing previously unexamined collections. Several new records for Mississippi streams have been produced using these data, including Cyprogenia aberti, the Western fanshell, and Plethobasus cyphyus, the sheepnose (Bogan 1987; Hartfield 1993; Peacock and James 2002; Peacock et al. 2011). These discoveries emphasize the power of zooarchaeological data for biogeographic and management purposes. For example, Roe (2004:4), in a conservation assessment of C. aberti, states, “The historical range of C. aberti includes the Arkansas, White, Black, and St. Francis river basins in Arkansas, Missouri, Kansas and Oklahoma,” a statement that is in error, as it has now been demonstrated zooarchaeologically that the species was present in a number of eastern tributaries in the Mississippi River drainage. A modern survey conducted partly as a result of these findings discovered a remnant population of P. cyphyus in the Big Sunflower River of the Mississippi Delta (Jones et al. 2005). Documenting this species, which has a state listing of critically imperiled and candidate status for federal listing as a threatened and endangered species, is a particularly important discovery in a part of the state where stream dredging is being proposed. The successful partnership between Mississippi State University and the Mississippi Museum of Natural Science provides a model for the kind of interdisciplinary collaboration that has repeatedly been called for but seldom achieved in the conservation of America’s wildlife.

Management Implications Scientists working to conserve mussel species face the stark reality that tributary streams have become isolated refugia (Haag 2009a) vulnerable to impacts from various land use activities, erosional episodes, and the like (Hartfield 1993; Warren and Haag 2005). This conclusion is based in part on the disparity between prehistoric and modern species biogeographies. The best way to ensure the preservation of species, and thus the genetic diversity they represent, is to reclaim habitat suitable for those species and their fish hosts (Haag and Warren 1998). The main channels of rivers such as the Tombigbee will not be undammed anytime soon, and efforts to create artificial habitat have had only limited effects (e.g., Miller 2006). Therefore, states should exercise eminent domain or other authority to obtain as many free-­flowing tributary streams as possible, preferably up the first order of such waterways (Bean 2000; Humphries and Winemiller 2009). Tax credits and conservation easements are potentially less contentious alternatives (Bean 2000). Stream courses should include buffer strips that are allowed to revegetate for filtration purposes (Brown and Krygier 1970). This would allow gradual desiltation, restoring habitat



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over time and allowing mussels and their host fish to replenish and repopulate lower-­order branches currently depauperate or entirely lacking in unionids. Reintroductions should be carried out concurrently (Bogan 1998; Haag 2009a; Humphries and Winemiller 2009). Strict controls on effluents, bridge emplacements, and other impacts must be maintained for protected streams (Bean 2000; Warren and Haag 2005). Expanding populations of filter-­feeding mussels would help decrease turbidity in target waterways. Over time, cleaner water emerging from these feeder streams would improve main channel habitat as well, at least near the mouth of the streams; an ideal goal would be to link populations between adjacent tributaries. Because reliable historical data are lacking, especially for smaller streams, archaeological shell should be used to determine which mussel species should be restocked in which streams. Existing archaeological shell data should be compiled nationally, as has been done for Mississippi (Peacock et al. 2011). For applied purposes, a “zooarchaeological present” can be assumed because changes brought about by prehistoric human actions were extremely modest compared to later impacts (Peacock et al. 2005). Some variation in populations might have occurred due to environmental changes during the Holocene (Peacock and Seltzer 2008), but these also appear to have been modest (Bogan 1990). Variation from both sources can be minimized by employing data from sites of various ages. Existing collections of unanalyzed archaeological shell should be a priority for analysis. Scientists from different disciplines should serve on graduate committees to foster consideration of the various complicating factors involved. Previously reported assemblages with suspect identifications should be reanalyzed. For all new archaeological projects, state regulatory agencies should not allow shell to be discarded (see also Hirst 2000:25–26) but should require expert analysis as a standard part of CRM (cultural resource management) projects in every case where shell is present.2 Because of their value for contemporary resource management, any archaeological sites containing shell should automatically be considered significant, at least until multisite taxonomic redundancy can be demonstrated along particular stream segments. While this requires a creative reading of 36 CFR 60.4(Code of Federal Regulations, Title 36, Part 60, section 4), criterion d, under which a site may be significant if it “has yielded, or may be likely to yield, information important in prehistory or history,” as argued in this chapter mussel data are of indisputable importance, even if the “history” involved is landscape history. Contrary to conventional thinking, a site that has been completely disturbed can be considered significant as long as shell is present. In fact, highly disturbed sites containing shell can be considered more significant in applied value than undisturbed sites as the shell can be recovered rap-

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idly (e.g., via mass removal of sediment and bulk washing through screens) without the painstaking methods usually required in archaeological excavation. Ceramics or other diagnostic artifacts so retrieved could be used to assess the degree of time averaging, or both time and space averaging (see Peacock 2000) could be assumed if large samples of shell were recovered. Formal channels of communication should be opened between state historic preservation offices and state wildlife agencies so that any mussel data generated by CRM are automatically shared. (This same recommendation should hold for any faunal data.) Taxpayers receive more bang for their buck when multiple agencies benefit from archaeology, and lawmakers who currently see CRM as a nuisance may come to understand its applied importance in our contemporary world if multiple needs are being served (Peacock and Rafferty 2007). Given the extinction of some species (and/or host fish), as well as the magnitude of habitat alteration, the reestablishment of original communities cannot be considered a viable goal in many cases (Humphries and Winemiller 2009). Instead, a species-­based approach is most appropriate, albeit at a watershed scale. Efforts should concentrate first on those areas where reduction of biodiversity has been greatest, as determined from zooarchaeological data in conjunction with historical and modern surveys. Areas where zooarchaeological and historical data indicate that changes have been relatively minor, such as the lower Gulf coastal plain (Peacock and Mistak 2008) or the Great Plains (Warren 2000), should receive attention later rather than sooner, honing the use of the limited money available for mussel conservation (see Lydeard et al. 2004). This plan doubtless will be perceived as a radical one. However, any measures short of what is proposed here are bound to result in the continued loss of species through competition with silt-­tolerant taxa or introduced species, chemical spills, and massive erosion episodes. The plan can be sold as beneficial to other wildlife and native plants as well, and the buffer areas could be open to public activities (e.g., hunting) that would not negatively affect the desired outcomes in terms of mussel restoration. In addition, the public allure of archaeology should be used as a selling point, via a “lessons from the past” approach for which American Indian tribes potentially can be enlisted as allies. Bogan (1998), Lydeard et al. (2004), Warren and Haag (2005), and others have noted the need for education and public outreach to help reduce the loss of mussels. Archaeology is an untapped resource in this regard. Given the importance of archaeological shell for modern resource management, funding for research should not be limited to contracts generated by CRM but should be made available by natural resource management agencies. In general, as Haag (2009a:709) has noted, combined historical and archaeological mussel data provide “an extraordinarily comprehensive re-



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cord of distribution and in some cases abundance throughout the Holocene that exists for few organisms in general and is unprecedented for invertebrates.” Without zooarchaeological data, even the best modern surveys present only the “after” picture, which includes both loss and additions to the original faunas. Although this chapter has focused on unionids, the arguments pertain to aquatic and terrestrial gastropods as well. A paradigm shift is needed among natural scientists in that invocations of the nonrepresentativeness of zooarchaeological collections must give way to recognition that bias is a manageable factor in the analysis of archaeological shell assemblages (e.g., Wolverton et al. 2010). Zooarchaeologists must publish the results of their studies, with their biogeographical implications, in nonarchaeological journals. As is evident in the references, it is probable that zooarchaeologists working with shellfish remains have published more in the biological, zoological, and conservation literature than have other applied zooarchaeologists. This meets the challenge set out by Lyman (1996, 2006; chapter 10, this volume) and others to make our data relevant to today’s world. The growing field of ethnobiology (Anderson et al. 2011) is contributing to more collaborative applied work than has been usual between archaeologists and members of other disciplines. While there is still a long way to go, the materials, the methods, and the rationales all are at hand to put the analysis of archaeological shell where it belongs—in the mainstream. Acknowledgments Thanks to Bob Jones, Paul Hartfield, Wendell Haag, Art Bogan, Mary Stevens, Scott Peyton, and others who have helped me analyze shell, helped me compile information, or otherwise supported my shell research over the years, with special remembrance of Paul Parmalee in this regard. Thanks to Bradley Carlock for allowing me to use shells from the Kinlock site, and to Paul Jacobs for the photograph used in figure 3.1. I am particularly indebted to Lee Lyman, Steve Wolverton, and Charles Randklev for their insightful comments on a draft of this work. Notes 1. The Vaughn Mound material seems to be comparable to the other main-­ river assemblages in taxonomic makeup, and the presence of very small shells and other artifacts suggests that it is not overly biased by the different recovery method employed (Peacock and Seltzer 2008), but this has not been formally tested. 2. CRM concerns conservation and preservation of historical and archaeo-

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logical resources. In a narrow sense, this includes initial discovery and documentation of sites through artifact recovery and recording, more in-­depth field testing of sites considered to be potentially eligible for listing on the National Register of Historic Places, and mitigation (avoidance or full-­scale excavation) of sites ultimately deemed to be eligible for listing.

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Williams, J. D., S. L. H. Fuller, and R. Grace. 1992. Effects of Impoundments on Freshwater Mussels (Mollusca: Bivalvia: Unionidae) in the Main Channel of the Black Warrior and Tombigbee Rivers in Western Alabama. Bulletin of the Alabama Museum of Natural History 13:1–10. Williams, J. D., M. L. Warren, K. S. Cummings, J. L. Harris, and R. J. Neves. 1993. Conservation Status of Freshwater Mussels of the United States and Canada. Fisheries 18:6–22. Wolverton, S., C. Randklev, and J. H. Kennedy. 2010. A Conceptual Model for Freshwater Mussel (Family: Unionidae) Remain Preservation in Zooarchaeological Assemblages. Journal of Archaeological Science 37:164–173. Woodrick, A. 1981. An Analysis of the Faunal Remains from the Gainesville Lake Area. In Biocultural Studies in the Gainesville Lake Area, by G. Caddell, A. Woodrick, and M. C. Hill, pp. 91–168. Report of Investigations 14, Office of Archaeological Research, University of Alabama, Tuscaloosa. ———. 1983. Molluscan Remains and Shell Artifacts. In Prehistoric Agricultural Communities in West Central Alabama, vol. 2: Studies of Material Remains from the Lubbub Creek Archaeological Locality, edited by C. S. Peebles, pp. 391–429. Report submitted to the US Army Corps of Engineers, Mobile District, by the University of Michigan, Ann Arbor.

CHAPTER FOUR

Prehistoric Biogeography and Conservation Status of Threatened Freshwater Mussels (Mollusca: Unionidae) in the Upper Trinity River Drainage, Texas Charles R. Randklev and Benjamin J. Lundeen

Historically, North America contained the most diverse and abundant population of freshwater mussels in the world, with nearly 300 species (Neves 1993). Unfortunately, habitat destruction stemming from sedimentation, impoundment of streams and rivers, release of environmental contaminants, and the introduction of invasive species has reduced this number (Lydeard et al. 2004; Neck 1982a; Strayer 1999). Current estimates suggest that 12% of the mussel species endemic to North America are now extinct and 23% are threatened or endangered (Galbraith et al. 2008 and references therein). The 52 species described in Texas have also been impacted, and many local streams and rivers are unable to support mussel populations at levels that existed in the past (Howells et al. 1996, 1997). As a consequence, 15 Texas species have recently been listed as threatened, and 11 of these are now the subjects of petitions for

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protection under the Endangered Species Act (ESA) (Texas Parks and Wildlife Department [TPWD] 2009). Listing a species under the ESA requires that decisions be made using the “best scientific and commercial (trade) data available” (Nicholopoulos 1999:8). For these species, “substantial information” using biological and biogeographic (past and present) data must demonstrate one of the following: 1. The destruction, modification, or curtailment of habitat or range 2. Overutilization for commercial, recreational, scientific, or educational purposes 3. Population decline related to disease or predation 4. Inadequacy of existing regulatory mechanisms for protecting existing populations 5. Natural or anthropogenic factors affecting a species’ continued existence (US Fish and Wildlife Service 2009) Presumably, this would also be the case for conservation listings at state or local levels. Unfortunately, for both rare and common species, modern and historical data regarding ecological preferences and biogeographic distributions are incomplete at best (Brown and Lomolino 1998; National Native Mussel Conservation Committee 1998). For unionids, the absence of basic biological data stymies conservation efforts. As a result, a national strategy was established in 1997 to help organizations identify tasks needed for the long-­term conservation of mussels (NNMCC 1998). Included in this framework was a call for an increase in sampling effort as well as for the gathering and dissemination of historical records to better understand the current status of mussel populations. However, the call did not mention the potential of paleozoological data sets for examination of the long-­term history of unionids. The potential value of such data is unprecedented because historic and modern data sets are often limited to some degree or biased temporally and spatially. It is therefore questionable whether modern data sets provide adequate baselines from which to infer biogeographic distributions and to measure species declines for the purposes of conservation and restoration. This is not to say that modern and historic accounts are not important but rather that they are insufficient to determine the long-­ term ecological processes responsible for mussel distributions (Humphries and Winemiller 2009; Randklev, Wolverton, et al. 2010; chapter 3, this volume). Given that conservation efforts tend to be driven by recent, and often limited, historical accounts, the extent or magnitude of the decline of poorly known species such as unionids may not be fully recognized by

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conservation biologists. As a result, the status of a given mussel species may be far worse than is recognized, regardless of whether that species is considered to be rare or common (Régnier et al. 2009). Mussel conservation efforts would benefit from information concerning the long-­term history of unionids because of the high stakes involved in conservation, such as local extirpation. Paleozoological data sets could provide insight on (1) the distributions of threatened species prior to large-­scale impacts (e.g., impoundments) and the degree to which their ranges have changed; (2) the ecological characteristics of those species that have experienced the greatest declines; and (3) locations of prehistoric hot spots for what are today threatened species, and whether or not these locations have been recently sampled. In this chapter, we discuss zooarchaeological data that pertain to several unionid species recently listed for protection, thereby providing information that should inform ongoing conservation efforts.

Background The upper Trinity River drainage is located in north central Texas and is characterized by a humid subtropical climate that is also continental and therefore subject to wide fluctuations in temperature and precipitation (Neck 1990). The major river systems in this drainage (figure 4.1) are the Clear, West, Elm, and East Forks of the Trinity River (Huser 2000). All of these watercourses are now impounded for flood control and for commercial and residential purposes (Randklev, Lundeen, et al. 2010). In general, most of the upper Trinity River drainage is heavily urbanized, which has resulted in groundwater depletion (Garrett 1972). As a result, instream flow is typically low but can rapidly fluctuate as a consequence of surface runoff following heavy local rainfall or impoundment release. Combined with these effects is the discharge of environmental contaminants from both point sources (i.e., wastewater treatment plants) and nonpoint sources (i.e., runoff, septic tanks, and illegal dumping), which has impacted not only the biota within the upper Trinity, but also how the river is managed and used (Coogan et al. 2007; Coogan and LaPoint 2008; Ward et al. 2001, 2002). The Trinity River mussel fauna is typical of those from the West Gulf Province, which includes rivers that drain to the south and west of the Mississippi drainage (Howells et al. 1996; Neck 1982b, 1990). However, very little is known about the distribution or abundance of mussel species in the upper Trinity River drainage (Neck 1990). The few historical records that exist are from the Elm Fork, from the Trinity River at the confluence of its forks near Dallas (Flook and Ubelaker 1972; Neck 1990;



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Figure 4.1. The Trinity River: black dots indicate locations of archaeological sites on the West (41TR114 and 41TR198) and Clear Forks (41TR205), Denton Creek (41DL8), and Rowlett Creek (41DL203). Black triangle indicates modern records for Fusconaia cf. flava (Randklev and Lundeen unpublished data). The black square denotes a single valve of Pleurobema riddellii collected from an archaeological site (41WS38) in the upper West Fork drainage.

Read 1954; Read and Oliver 1953; Singley 1893; Strecker 1931), and from the Clear and West Forks near Fort Worth (Mauldin 1972); these records include data on reservoirs associated with the drainages. Modern accounts have focused on both reservoirs and rivers (Howells 2006). Historical records indicate that two species now considered to be threatened (table 4.1) occurred in this drainage: Potamilus amphichaenus and Pleu-

72

Texas pigtoe Triangle pigtoe Texas fatmucket Sandbank pocketbook Southern hickorynut Louisiana pigtoe Texas hornshell Texas heelsplitter Salina mucket Golden orb Smooth pimpleback False spike Texas pimpleback Mexican fawnsfoot Texas fawnsfoot

Common Name LR/NT LR/NT LR/NT LR/NT LR/NT LR/NT CR EN — — LR/NT CR — NR —

IUCN I CI CI I I CI CI CI CI CI I PE I CI I

NS — U U — — U C U U U U U U U U

USFWS SC SC SC SC SC SC T T T SC T T T EN EN

AFS

T T T T T T T T T T T T T T T

TPWD

The conservation status of each species is designated by the following conservation, state, and federal agencies: International Union for Conservation of Nature (IUCN); NatureServe (NS); US Fish and Wildlife Service (USFWS); American Fisheries Society (AFS; given by Williams et al. 1993); and Texas Parks and Wildlife (TPWD). Abbreviations: C, candidate for listing; CI, critically imperiled; CR, critically endangered; EN, endangered; I, imperiled; LR/NT, lower risk/near threatened; NR, not ranked; PE, possibly extinct; SC, special concern; T, threatened; and U, under review. For definitions of status listings see IUCN (2001), NS (2009), USFWS (2009), Williams et al. (1993), and TPWD (2003). *Mussel species reported in the upper Trinity River drainage.

Fusconaia askewi Fusconaia lananensis Lampsilis bracteata Lampsilis satura Obovaria jacksoniana Pleurobema riddellii* Popenaias popeii Potamilus amphichaenus* Potamilus metnecktayi Quadrula aurea Quadrula houstonensis Quadrula mitchelli Quadrula petrina Truncilla cognate Truncilla macrodon

Species

Table 4.1.  Summary of Status Listings of 15 Mussels Recently Placed on the Threatened List in Texas



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robema riddellii (Howells et al. 1996). Lampsilis satura (Neck 1990; Read 1954), Quadrula houstonensis (Read 1954; Strecker 1931), and Truncilla macrodon (Strecker 1931) have also been reported from this area, but more recent studies have dismissed these accounts as misidentifications (Howells 2000, 2002; Neck 1990; Randklev, Wolverton, et al. 2010). Of the 15 threatened species, only P. amphichaenus has been collected in recent years in the upper Trinity (Howells et al. 1996; Howells 2000; Neck and Howells 1994). Given the limited number of mussel surveys conducted in the upper Trinity River drainage, the distribution and, more importantly, the status of each of the 15 species recently listed as threatened is poorly known. Therefore, the resolution of modern or historic accounts as benchmarks for assessing species distributions and measuring species declines is limited. Fortunately, zooarchaeological data are available that allow an examination of unionid biogeography prior to historical and modern impacts in this drainage. Thus, our goal is to examine the paleozoological evidence to determine if threatened mussels were found in the upper Trinity so that their range decline can be more comprehensively measured. This will allow us to determine if all 15 listed species are indeed threatened or are simply maintaining preindustrial impact distributions.

Materials and Methods To document the biogeography of threatened mussels prior to modern human impacts, we analyzed faunal remains from five archaeological sites dating between 2500 and 600 years before the present (Randklev and Wolverton 2009; Wolverton et al. 2010). Zooarchaeological collections were selected based on their availability and on the presence of unionid remains. Archaeological sites were located near the Clear Fork (official State of Texas archaeological site number 41TR205) and West Fork (41TR114 and 41TR198) of the Trinity River, as well as on Denton Creek (41DL8, Elm Fork) and Rowlett Creek (41DL203, East Fork) in north Texas (figure 4.1); with the exception of the last two, all rivers are now impounded. For each zooarchaeological shell fauna, species identifications were made using freshwater mussel guides (Howells et al. 1996; Parmalee and Bogan 1998) and through comparison to reference specimens in the Joseph Britton Freshwater Mussel Collection housed at the University of North Texas. Identified unionids were counted using two quantitative units: the number of specimens (NSP, both taxomically identified and unidentified umbos) and the number of nonrepetitive elements (NRE, number of identified umbos; Giovas 2009; Mason et al. 1998).

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The absolute abundances of unionids that existed in the upper Trinity River drainage during the late Holocene will never be known, because archaeological assemblages are often modified to some degree by cultural harvesting preferences, differential preservation, and differences in recovery techniques (Peacock 2000). Further, the absence of a particular species from an archaeological site is not necessarily evidence that it was not present at that site (Lyman 2008a). Shell properties such as shape and density affect how well the shell is preserved and therefore whether it can be identified (Kosnik et al. 2009; Wolverton et al. 2010). In highly fragmented assemblages, taxa with spherical and/or dense shells occur more often and are therefore proportionally more abundant. In these cases, species representation may be the result of postdepositional preservation factors rather than a reflection of the late Holocene aquatic environment. To evaluate whether or not differential preservation influenced taxonomic abundance in shell assemblages, we calculated the proportion of taxonomically identifiable umbos from each archaeological site (see Peacock and Chapman 2001; Wolverton et al. 2010, for further details). Based on the presumption that fragmentation influences identifiability, we calculated the ratio of NRE to NSP; the higher the value of this ratio, the larger the number of identifiable umbos and the less fragmented and better preserved the assemblage (Lyman 1994; Peacock and Chapman 2001; Wolverton 2002). Taxa may be underrepresented or absent in an assemblage not only because of poor preservation, but because of lack of recovery. The probability of recovering a given taxon is determined in part by its abundance in the sampled community. Therefore, taxa that tend to be rare on the landscape are typically absent from shell assemblages with small sample sizes, all else being equal (Lyman 2008a). To assess possible recovery bias, we graphed the total number of identified specimens (left and right valves combined) against the number of threatened taxa (NTAXAthreatened) for all five archaeological sites (see Lyman 2008a; 2008b:149–152, for further details). If threatened taxa are rare in, or absent from, small assemblages but present or abundant in large assemblages, then their absence from or rarity in drainages with small assemblages may be an artifact of archaeological sampling rather than a measure of their occurrence in a drainage.

Results Nineteen unionid species were identified in the five zooarchaeological assemblages (table 4.2). Of the taxa considered to be threatened in Texas



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(table 4.1), shells of Lampsilis cf. satura were recovered only from 41TR198 on the West Fork of the Trinity River near Forth Worth, which is outside of its modern range (figure 4.2A). Pleurobema riddellii was collected at archaeological sites on the Clear and West Forks of the Trinity River and on Denton and Rowlett Creeks, suggesting a ubiquitous distribution over the last 2500 years (figure 4.2B). Zooarchaeological specimens of P. riddellii in the Clear and West Forks of the Trinity River (41TR198, 41TR205, and 41WS38) are outside its current distribution (figures 4.1 and 4.2B). Potamilus amphichaenus is absent from all shell assemblages in the upper Trinity River drainage, which is puzzling given its historical and modern occurrence there. However, sample size effects and postdepositional destruction of shells may explain the absence of this species (see below). Fusconaia cf. flava occurs at all five archaeological sites, which are within the modern range of this species (figure 4.2C). This species is not listed for protection because of uncertainties regarding its taxonomic status. However, it has been suggested that if ongoing genetic studies confirm its taxonomic validity, it should be listed as threatened (Howells 2009), and for this reason we have included it in our discussion. Shape and density mediate fragmentation and therefore whether a unionid shell (or fragment thereof) can be identified taxonomically. Species with shells that are rectangular in outline and low in density are less likely to be preserved compared to species with shells that are spherical and relatively dense (Wolverton et al. 2010). Shells of P. amphichaenus are thin as well as elongated and are therefore prone to fragmentation. As a result, it is unlikely that remains of this species would survive; its presence in the upper Trinity River drainage during the late Holocene cannot be ruled out. Shells of Lampsilis satura are thin but more spherical in shape, which increases the likelihood that they will be preserved. However, this species only occurred at one site (41TR198), which also had the largest number of identifiable valves. This suggests that its presence is a function of sample size (see below). Fusconaia cf. flava and P. riddellii are dense and spherical in shape and thus their remains are more likely to be preserved. Both species are present in a number of shell assemblages in the upper Trinity River drainage. However, P. riddellii was absent from 41TR114 (West Fork of the Trinity River), which had the lowest number of identifiable valves. Thus, its absence there is probably the result of sampling error. The probability of discovery of a taxon, assuming it occurred in the region in the past, should increase with larger sample size and/or better preservation (Lyman 2008a; Wolff 1975). Figure 4.3A emphasizes this point for shell assemblages in the upper Trinity drainage; as sample size (log NRE) increases, so does the NTAXAthreatened in an assemblage. For

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Table 4.2. Taxonomic List, Relative Abundance, and Number of Unionids (NRE) Recovered from Archaeological Sites Located in the Upper Trinity River Drainage 41DL8

41TR114

41TR198

41TR205

41DL203

No.

No.

%

No.

%

Species

No.

%

No.

%

%

Amblema plicata Arcidens confragosus Fusconaia cf. flava Lampsilis sp. Lampsilis hydiana Lampsilis cf. satura Lampsilis teres Leptodea sp. Ligumia sp. Ligumia cf. subrostrata Obliquaria reflexa Plectomerus sp. Plectomerus dombeyanus Pleurobema sp. Pleurobema riddellii Potamilus sp. Potamilus purpuratus Quadrula sp. Quadrula apiculata Quadrula mortoni Quadrula nobilis Quadrula verrucosa Toxolasma sp. Toxolasma texasiensis Truncilla sp. Truncilla cf. donaciformis Truncilla truncata Uniomerus tetralasmus Total (NRE) Unidentifiable umbos Total assemblage (NSP) % NRE to NSP

31 — 9 2 8 — — — — 1

29.8 — 8.7 1.9 7.7 — — — — 1.0

10 — 25 3 1 — 1 — — —

14.7 — 36.8 4.4 1.5 — 1.5 — — —

253 8.7 2 0.1 1394 47.8 6 0.2 61 2.1 9 0.3 34 1.2 5 0.2 1 0.0 3 0.1

38 — 4 58 4 — 4 — 31 19

17.9 — 1.9 27.4 1.9 — 1.9 — 14.6 9.0

80 — 4 8 6 — 22 — — —

41.9 — 2.1 4.2 3.1 — 11.5 — — —

2 1 11

1.9 1.0 10.6

— 5 9

— 7.4 13.2

27 — 417

0.9 — 14.3

— 7 19

— 3.3 9.0

— — 8

— — 4.2

— 2 2 2

— — 1.9 — 1.9 — 1.9 2

— — — 2.9

3 259 8 5

0.1 8.9 0.3 0.2

2 1 — 1

0.9 0.5 — 0.5

— 3 — —

— 1.6 — —

6 1 14 1 — — 3

5.8 4 1.0 — 13.5 — 1.0 — — 8 — — 2.9 —

5.9 — — — 11.8 — —

— 5 224 2 117 — 2

— 0.2 7.7 0.1 4.0 — 0.1

1 — 2 — 9 6 1

0.5 — 0.9 — 4.2 2.8 0.5

— — — — 2 1 2

— — — — 1.0 0.5 1.0

2 —

1.9 —

— —

— —

— 2

— 0.1

— —

— —

— —

— —

3 3

2.9 2.9

— —

— —

76 —

2.6 —

2 3

0.9 1.4

— 55

— 28.8

104 112

68 209

2915 4380

212 272

191 1145

216

277

7295

484

1336

48.1

24.5

40.0

43.8

14.3

Sites distributed as follows: Denton Creek, 41DL8; West Fork, 41TR114 and 41TR198; Clear Fork, 41TR205; Rowlett Creek, 41DL203. NRE, nonrepetitive elements; NSP, number of specimens.



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Figure 4.2.  General historical and modern distributions for (A) Lampsilis satura; (B) Pleurobema riddellii; and (C) Fusconaia flava. The solid black line for all three maps is taken from Howells et al. (1996) and indicates the western boundary of historically known or potential ranges. Dashed lines indicate the boundary of ranges for Lampsilis cardium (A) and Fusconaia askewi (C). Historical records are from published accounts dating between 1892 and 1991; modern records are from published and unpublished accounts dating between 1992 and the present; prehistoric records date between 2500 and 600 years before the present.

example, Lampsilis cf. satura only occurs at site 41TR198, which produced the largest number of identified valves (table 4.2). This suggests that if each assemblage had sample sizes similar to that of 41TR198, it is likely that they would have produced more shells of the threatened species. Therefore, the presence of Lampsilis cf. satura at sites with small sample sizes cannot be ruled out. Similarly, the absence of P. amphichae-

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Figure 4.2.  Continued.

nus from 41TR198 and its presence in historical and modern accounts suggests that larger zooarchaeological samples may reveal that it inhabited the upper Trinity River drainage during the late Holocene. The intensity of fragmentation of shells in a particular assemblage may also explain the analytical absence of a species. To determine whether or not this is the case with the threatened species, we graphed the relationship between percentage NRE:NSP and the richness of threatened taxa for each archaeological site. Figure 4.3B suggests that, in general, a higher number of threatened species are identified when shells are less fragmented, but this relationship is statistically insignificant and weak. However, fragmentation exacerbates the influence of sample size on measures of NTAXA when assemblages are small. As a result, the absence of



Biogeography and Conservation Status of Freshwater Mussels

79

Figure 4.2.  Continued.

threatened species from shell assemblages with small sample sizes and high fragmentation rates is likely not evidence of their absence from the upper Trinity River drainage during the late Holocene. In summary, the valves of threatened species that were present and abundant in the upper Trinity River drainage are spherical and/or dense, which indicates differential preservation according to interspecific variability in shell robustness. For threatened species that are thin shelled, presence in the upper Trinity River appears to be a function of sample size. Small sample size and differential preservation may have reduced or eradicated the occurrence of threatened species in shell assemblages for this drainage. Thus, the late Holocene presence of P. amphichaenus and Lampsilis cf. satura within the upper Trinity River cannot be ruled out.

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Figure 4.3.  (A) Relationship between total unionid NRE (sample size) and NTAXAthreatened of Fusconaia sp., Lampsilis satura, and Pleurobema riddellii for five archaeological sites in the upper Trinity River basin. The simple best-fit line is shown for reference (r2 = 0.66, p  0.05). The calibrated ages are in correct stratigraphic sequence (oldest at bottom, youngest at top) and, when modeled as a sequence of events (after Ramsey 1995, 2001), suggest a very brief period of deposition (table 5.1). It seems that the anomalously dense Pismo clam midden deposits at CA-­SCRI-­480 accumulated over a mere 50 years or so. The temporal resolution represented by the five presumably temporally distinct excavation levels reveals the rapidity with which a local population of Pismo clams might respond demographically to a perturbation.



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99

Table 5.1. Radiocarbon Dates for CA-SCRI-480 Level 10–20 cm 30–40 cm 40–50 cm

Accession No. Beta-231846 Beta-231847 Beta-231848

14

C Yr

1750 ± 50 1780 ± 60 1700 ± 40

2 Sigma Calendar Yr

2 Sigma Modeled Yr

Midpoint of Modeled Yr

1180–920 1230–930 1110–890

1090–910 1120–940 1150–960

1000 1030 1055

Note: Modeling after Ramsey (1995, 2001).

Pismo clams and California mussel (Mytilus californianus) shells dominate the assemblage. Based on shell weight, the four deepest (oldest) levels contain a greater proportion of Pismo clams than any other taxon, and more than half of the total weight of shell per level (figure 5.1). This pattern, including California mussel shell weight comprising less than 40% of shell weight in the five deepest levels, reverses in the uppermost (youngest) level, in which the relative abundance of Pismo clam drops precipitously and California mussel makes up 69.9% of the assemblage. Further, a number of small rocky intertidal shellfish species, absent in the lowest deposits, contribute more in this upper level (table 5.2). Of the 13 shellfish species identified, 9 appear in the 40–50-­cm level, 10 in the 30–40-­cm level, 12 in the 20–30-­cm level, and 12 in the 10–20-­cm level,

Figure 5.1.  Proportion of Pismo clam shell weight per level in the CA-SCRI-480 shellfish assemblage.

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Table 5.2. Taxonomic Abundances by Weight (gm) of Shell per Excavation Level Pismo Clam 0–10 cm 10–20 cm 20–30 cm 30–40 cm 40–50 cm

California Black Mussel Abalone

97 775 1209 3141 1363

394 527 501 607 427

27 4 7 48 6

Acorn Barnacle

Sea Urchin

Black Turban

Box Mussel

9 10 23 43 26

16 7 23 38 19

8 10 23 11 8

2 2 5 6 —

and all 13 appear in the 0–10-­cm level (table 5.2). The correlation between taxonomic richness per level and total shell weight per level is negative and marginally significant (ρ = −0.872, p = 0.08), suggesting richness is not a function of sample size and, more importantly, that as the Pismo clam population shrank, more taxa were added by humans, which may have been for diet expansion or the fact that some smaller species attach to the shells of larger ones in the rocky intertidal zone. An MNI of 73 Pismo clams were in the 40–50-­cm level, 81 in the 30–40-­cm level, 32 in the 20–30-­cm level, 13 in the 10–20-­cm level, and 2 in the 0–10-­cm level. Using stratigraphic level as a measure of time, there is a statistically significant decrease in the MNI of Pismo clams that humans harvested and consumed over time (ρ = 0.90, p = 0.037). Finally, the smallest Pismo clam shell (a mere 20 mm) was recovered from the lowermost level; the second largest (90 mm) was recovered from the uppermost level (table 5.3). The mean size of Pismo clam valves in a given level was compared with the mean size of Pismo clam valves in the preceding and in the subsequent levels. Unpooled t-­tests reveal that the differences in means between each pair of adjacent excavation levels are all significant (table 5.4). There is a statistically significant increase in the average size of Pismo clams through time (ρ = −1.0). Table 5.3. Descriptive Satistics of Pismo Clam Shell Length (mm) by Level

Level 0–10 cm 10–20 cm 20–30 cm 30–40 cm 40–50 cm

N of Shells Measured

Mean

Standard Deviation

Minimum

Maximum

3 13 17 111 82

86.02 67.43 48.52 36.52 30.48

4.55 16.78 14.52 4.39 5.92

81.18 46.07 32.01 26.39 20.44

90.21 95.63 75.30 46.75 42.30



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Ancient Actions Predict Modern Consequences

Table 5.2. Continued

Crab

Gooseneck Barnacle

Limpet

Whelk

Ollivela

Chiton

Total Shell Weight

5 1 2 6 3

1 2 3 5 4

1 2 — — —

2 — 1 — —

1 1 2 — —

1 2 2 3 4

564 1343 1802 3912 1865

The pattern of smaller clam shells in the lower levels and larger clam shells in the higher levels requires consideration of postdepositional redistribution. Wood and Johnson (1978) and Erlandson (1984) indicate that downward sorting of fine particles in archaeological deposits may occur as a result of many disturbance processes, not the least of which is faunalturbation. However, no burrowing animals have ever inhabited Santa Cruz Island and thus archaeological deposits on the island maintain excellent stratigraphic integrity. Further, though the Pismo clam shells in the deposits exhibit significant patterning in shell size, such is not the case for other species. Several species of very small mollusks are represented throughout the deposit, contrary to models of size sorting. Indeed, the number of small-­shelled taxa represented increases through time. The statistically significant increase in the size of Pismo clams through time does not appear to be related to postdepositional processes. Over a short period of time, the number and size of Pismo clams exploited by the prehistoric inhabitants of the site changed dramatically. Initially, people were harvesting and consuming large numbers of small clams. Through time the number of individual clams collected decreased and the average size of individual clams concurrently increased. There was a transition from collecting great quantities of relatively small (young) Pismo clams to collecting smaller quantities of larger (old) Pismo clams. What catalysts drove this rapid transition in the demographic characteristics of Santa Cruz Island’s Pismo clams? Table 5.4.  Comparison of Average Pismo Clam Shell Length between Pairs of Adjacent Excavation Levels

t p

Level 1– Level 2

Level 2– Level 3

Level 3– Level 4

Level 4– Level 5

3.479 0.004

3.242 0.003

3.384 0.004

7.783 .000

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Climatic or Prehistoric Anthropogenic Influences? Two possible catalysts drove the transition in demographic (individual size, abundance) characteristics of Santa Cruz Island’s Pismo clams. One is natural; one is anthropogenic. The former involves climatic change, particularly the temperature of seawater. Fluctuation in this variable has been shown to correspond with, and likely drive, change in abundance and size of some shellfish in the Channel Island area (e.g., Arnold and Tissot 1993; Braje et al. 2007; Kennett and Kennett 2000). Studies documenting this correspondence have been at a very coarse temporal resolution and perhaps as a result reveal no shift in water temperature during the time when the CA-­SCRI-­480 Pismo clams were deposited. Further, change in water temperature could cause either an increase in clam abundance and concomitant decrease in size, or a decrease in abundance and concomitant increase in size. On one hand, then, environmental effects on the Pismo clam population can be neither rejected nor supported by available evidence. Anthropogenic causes have been shown to influence the abundance and size of at least some Channel Island shellfish taxa (e.g., Braje et al. 2007; Kennett and Kennett 2000). The pattern in clam (decreased) abundance and (increased) size observed archaeologically is the one theoretically predicted to result from human harvest (e.g., a midden deposit). Although not fully conclusive, the weight of the evidence suggests an anthropogenic cause of the clam size and abundance trends rather than an environmental one. Additional testing with more materials from the site would be ideal but, in the absence of that, I proceed with the tentative conclusion of human predation as the cause. The period of Pismo clam exploitation evidenced at CA-­SCRI-­480 likely spanned about 50 years. Fortunately, this roughly coincides with the potential life span of a Pismo clam. I say “fortunately” because were the time span shorter, the influence of human predation would not be so obvious because only a portion of the life span of the cohort would be represented. Further, if the deposit represented a longer time span, say, 1000 years or more, the temporal resolution provided by the five excavation levels would be much coarser. They would each potentially include multiple colonization and population depletion events, thereby muting the demographic signals (age, abundance) in the prey population that signify predation. It is probable that the midden deposits at CA-­SCRI-­480 reflect human exploitation of a single Pismo clam population shortly after colonization. Such colonization events were likely rare due to the geographic isolation of Santa Cruz Island, the lack of healthy Pismo clam populations on nearby beaches, and intensive oceanographic currents. If there had been multiple colonization events during the time when the CA-­SCRI-­480



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midden was being accumulated, the trends of decreased clam abundance and increased size of individual clams would likely not be evident and certainly would not be so stark. The only explanation for those trends, particularly over the short time represented, is that the archaeological Pismo clams represent a single colonization event. The arrival of hundreds of thousands of Pismo clams in a single colonization provided the prehistoric inhabitants of Santa Cruz Island with an abundant and novel food resource, but one that quickly assimilated demographic characteristics related to predation by humans. Soon after Pismo clams colonized Christy Beach, prehistoric humans began to harvest young Pismo clams in great quantities. For the first several years, a large proportion of the individuals collected were likely at or below reproductive age. Unknowingly, people systematically reduced the possibility that the local Pismo clam population at Christy Beach would be able to effectively sustain itself through local reproduction or contribute to the colonization of adjacent beaches. The accelerated decline of an aging population is evident in the dramatic decrease in the number of Pismo clams and concurrent increase in shell size (and individual age). Pismo clams have a tendency to migrate from the high intertidal zone into the lower intertidal and sometimes into the subtidal zone during their life course. Thus, collection of Pismo clams would have become increasingly labor intensive as the clams decreased in both number and accessibility. Surviving older and difficult-­to-­access Pismo clams contributed progressively less to the shellfish assemblage in successively younger levels. Data presented here illustrate a brief but dynamic period of prehistoric Pismo clam exploitation that exacerbated difficult reproductive circumstances and accelerated the decline of the Christy Beach population. However, the story is not entirely bleak. Other Pismo clam larval recruitment events have occurred at various times in the past at Christy Beach as well as at other beaches around the Channel Islands. This is demonstrated in a second column sample extracted from CA-­SCRI-­480, which also contained a relatively high concentration of Pismo clam shells (but from much shallower deposits), yet this deposit is associated with a much later radiocarbon date than the column sample analyzed here. This suggests the likelihood that even the most devastated Pismo clam population can rebound or that colonization of isolated loci can occur multiple times.

Management Implications Investigation of prehistoric human predation helps us understand human impacts on vulnerable Pismo clam populations and the long-­term consequences of such impacts. The lessons learned here are far from localized.

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This case study clearly demonstrates the potential consequences of sustained human predation on a small biogeographically isolated population, similar to studies of other taxa (e.g., chapters in Dean 2010). This natural experiment provides conservation biologists with a glimpse of the entire life course of a single small population, from first colonization to the verge of extirpation. Very few modern ecological studies approach this fine scale of analysis. Short-­term, fine-­resolution diachronic studies, as exemplified here, are essential for deciphering long-­term population responses to anthropogenic (and environmental) pressures. Data discussed in this chapter illustrate the importance of regular larval recruitment from the metapopulation for the maintenance of healthy local Pismo clam populations. Geographic isolation, due to coastal morphology and oceanographic processes, reduced regular larval recruitment and weakened the population’s link to the regional metapopulation. Aging local Pismo clam populations that receive only poor or episodic recruitment are exceptionally vulnerable to even low levels of sustained predation. In turn, local extinctions can profoundly affect demographic persistence, genetic variation, and evolution of the metapopulation. Metapopulations depend on isolated breeding populations to contribute genetic diversity (figure 5.2). Genetic variability is a key factor in a metapopulation’s ability to adapt to environmental change and will become increasingly important as climate change continues (Willi et al. 2006; Woodruff 1990). Geographic separation of local populations is a primary source of genetic variation within metapopulations (Smedbol et al. 2002). However, isolated local populations must persist long enough to produce local variation. Maintenance of metapopulations for longer-­ term evolution depends upon input of new genetic variability (Reed and Bryant 2000). Thus, the preservation of even the smallest breeding populations is essential to the overall health and survivability of the metapopulation. Local population turnover has been shown to decrease genetic variability in metapopulations, both because alleles are often lost as local populations are extirpated, and because founding of new populations by just a few individuals creates genetic bottlenecks (Smedbol et al. 2002). Although the archaeological evidence suggests that Christy Beach was repopulated by Pismo clams, the decline of the original population due to anthropogenic pressures decreased the overall genetic variability of the metapopulation. As remaining coastal habitats become smaller and more fragmented, it is increasingly important to understand the ecological and evolutionary dynamics of small populations and metapopulations in order to effectively manage and protect them (Lande 1988). The establishment of marine protected areas must compete with commercial and public land use.



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Figure 5.2.  A model of marine shellfish metapopulation dynamics (adapted from Kritzer and Sale 2003:5). Upper: arrows signify gene flow within and between local populations comprising a metapopulation. Lower: probability of gene dispersal between local populations decreases with increasing distance between local populations.

This requires conservation biologists to provide precise prescriptions for how much land or beach is sufficient to support minimum viable populations and how close refugia must be to maintain minimum viable metapopulations (Shaffer 1981). Guidance for such prescriptions can be gleaned from the zooarchaeological record as shown here. Zooarchaeological research described here suggests conservation policies that nurture healthy aggregations of sexually mature Pismo clams and protect young Pismo clams will enhance the resiliency of this species. Just as poor recruitment exacerbated the prehistoric human impact on Pismo clams, regular larval recruitment is a key to Pismo clams surviving future challenges. Conservation of the Pismo clam metapopulation along the California mainland coast and offshore islands must focus on preserving successful larval dispersal between isolated local populations. Without this, minimum viable populations cannot be sustained and cannot contribute to the genetic diversity of the metapopulation. Protecting breeding stocks of sexually mature Pismo clams must also be a priority. Beaches must be able to maintain sufficiently large aggregations of Pismo clams to ensure successful fertilization and larval production. This requires strict clamming protocols to prevent harvesting of sexually immature clams and overharvesting of adults. Further, local populations must be sufficiently close to one another, with consideration for geographic and oceanographic variability, to ensure successful recruitment between populations.

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Arnold, J. E., and B. N. Tissot. 1993. Measurement of Significant Marine Paleotemperature Variation Using Black Abalone Shells from Prehistoric Middens. Quaternary Research 39:390–394. Braje, T. J., D. J. Kennett, J. M. Erlandson, and B. J. Culleton. 2007. Human Impacts on Nearshore Shellfish Taxa: A 7000 Year Record from Santa Rosa Island, California. American Antiquity 72:735–756. Breschini, G. S., and T. Haversat. 1991. Early Holocene Occupation of the Central California Coast. In Hunter–Gatherers of Early Holocene Coastal California, edited by J. M. Erlandson and R. H. Colten, pp. 125–132. Institute of Archaeology, University of California, Los Angeles. Briggs, B. M., K. A. Spielmann, H. Schaafsma, K. W. Kintigh, M. Kruse, K. Morehouse, and K. Schollmeyer. 2006. Why Ecology Needs Archaeologists and Archaeology Needs Ecologists. Frontiers in Ecology and the Environment 4:180–188. Broitman, B., C. A. Blanchette, and S. D. Gaines. 2005. Recruitment of Intertidal Invertebrates and Oceanographic Variability at Santa Cruz Island, California. Limnology and Oceanography 50:1473–1479. Brook, B. W., L. W. Traill, and C. J. A. Bradshaw. 2006. Minimum Viable Population Sizes and Global Extinction Risk Are Unrelated. Ecology Letters 9:375– 382. Caughley, G. 1977. Analysis of Vertebrate Populations. John Wiley and Sons, London. Coe, W. R. 1947. Nutrition, Growth and Sexuality of the Pismo Clam (Tivela stultorum). Journal of Exploratory Zoology 104:1–24. Coe, W. R., and J. E. Fitch. 1950. Population Studies, Local Growth Rates and Reproduction of the Pismo Clam (Tivela stultorum). Journal of Marine Research 9:188–210. Dean, R. M. (editor). 2010. The Archaeology of Anthropogenic Environments. Center for Archaeological Investigations Occasional Paper No. 37. Southern Illinois University Press, Carbondale. Deevey, E. S., Jr. 1947. Life Tables for Natural Populations of Animals. Quarterly Review of Biology 22:283–314. Dugan, J. E., D. M. Hubbard, D. L. Martin, J. M. Engle, D. M. Richards, G. E. Davis, K. D. Lafferty, and R. F. Ambrose. 2000. Macrofauna Communities of Exposed Sandy Beaches on the Southern California Mainland and Channel Islands. In Proceedings of the Fifth California Islands Symposium MMS 99-­ 138, pp. 339–346. Santa Barbara Museum of Natural History, Santa Barbara, CA. Dugan, J. E., D. M. Hubbard, M. D. McCrary, and M. Pierson. 2003. The Response of Macrofauna Communities and Shorebirds to Macrophyte Wrack



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Subsidies on Exposed Sandy Beaches of Southern California. Estuarine, Coastal and Shelf Science 585:133–148. Erlandson, J. M. 1984. A Case Study in Faunalturbation: Delineating the Effects of the Burrowing Pocket Gopher on the Distribution of Archaeological Materials. American Antiquity 49:785–790. Fitch, J. E. 1952. The Pismo Clam in 1951. California Fish and Game Bulletin 36:541–547. Fitzgerald, R. T. 2000. Cross Creek: An Early Holocene/Millingstone Site. California State Water Project, Coastal Branch Series Paper vol. 12. San Luis Obispo County Archaeological Society, San Luis Obispo, CA. Gascoigne, J., L. Berec, S. Gregory, and F. Courchamp. 2009. Dangerously Few Liaisons: A Review of Mate-­Finding Allee Effects. Population Ecology 51:355– 372. Gascoigne, J., and R. N. Lipcius. 2004a. Allee Effects in Marine Ecosystems. Marine Ecology Progress Series 269:49–59. ———. 2004b. Allee Effects Driven by Predation. Journal of Applied Ecology 41:801–810. Glassow, M. A. 1997. Middle Holocene Cultural Development in the Central Santa Barbara Channel Region. In Archaeology of the California Coast During the Middle Holocene, edited by J. M. Erlandson and M. A. Glassow, pp. 73– 90. Perspectives on California Archaeology 4. Institute of Archaeology, University of California, Los Angeles. Hanski, I., A. Moilanen, and M. Gyllenberg. 1996. Minumum Viable Metapopulation Size. American Naturalist 147:527–541. Hastings, A., and S. Harrison. 1994. Metapopulation Dynamics and Genetics. Annual Review of Ecology and Systematics 25:167–188. Hewitt, W. G. 1946. Marine Ecological Studies on Santa Cruz Island, California. Ecological Monographs 16:185–210. Hobday, A. J., M. J. Tegner, and P. L. Haaker. 2001. Over-­exploitation of a Broadcast Spawning Marine Invertebrate: Decline of the White Abalone. Reviews in Fish Biology and Fisheries 10:493–514. Kennett, D. J. 2005. The Island Chumash: Behavioral Ecology of a Maritime Society. University of California Press, Berkeley. Kennett, D. J., and J. P. Kennett. 2000. Competitive and Cooperative Responses to Climatic Instability in Coastal Southern California. American Antiquity 65:379–395. Kinlan, B. P., M. H. Graham, and J. M. Erlandson. 2005. Late Quaternary Changes in the Size, Shape, and Isolation of the California Islands: Ecological and Anthropological Implications. In Proceedings of the Sixth California Islands Symposium, edited by D. K. Garcelon and C. A. Schwemm, pp. 119– 130. Institute for Wildlife Studies, Arcata, CA. Kritzer, J. P., and P. F. Sale. 2003. Metapopulation Ecology in the Sea: From

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CHAPTER SIX

The Overkill Hypothesis and Conservation Biology Lisa Nagaoka

The overkill hypothesis was originally proposed more than 40 years ago as an explanation for the extinction of North American Pleistocene megafauna (Martin 1967; Mosimann and Martin 1975). The hypothesis states that when people arrived in the Americas, they hunted all the megafauna (mammals of ≥44 kg adult body weight) to extinction. By the 1980s the overkill hypothesis had broadened its explanatory scope to cover megafaunal extinctions worldwide (Martin 1984). The model is simple: when people colonize new lands, megafauna (and often other animals) go extinct. Many researchers have evaluated the archaeological and paleontological evidence for and against overkill (e.g., Burney and Flannery 2006; Faith and Surovell 2009; Fiedel and Haynes 2004; Grayson 1984b, 2001; Grayson and Meltzer 2003, 2004; Haynes 2007; Koch and Barnosky 2006; Wolverton et al. 2009; Wroe and Field 2006; Wroe et al. 2006), and the debate surrounding the causes of Pleistocene megafaunal extinction rages on. Nevertheless, many conservation biologists and ecologists treat overkill as a fact (e.g., Brook and Bowman 2005; Johnson 2002; Kelt and Meyer 2009) and use it as a cautionary tale about the negative impacts of human actions (Donlan et al. 2005, 2006; Eldredge 2001; Lyons et al. 2004; Trombulak et al. 2004). Overkill is also used as a point of reference for ecological restoration. For example, it is the basis for the Pleistocene rewilding agenda in North America, proponents of which argue that the

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proper restoration benchmark period should be the day before Homo sapiens colonized a new area (Donlan 2007; Donlan et al. 2005, 2006; Martin 2005). As such, the overkill model is one of the most commonly used examples of applied paleozoology, but it is a poor example of such in conservation biology and restoration ecology because of its implicit assumptions about animal ecology and its troubling use of paleozoological data (or lack thereof). This study examines the integral part of the overkill argument that assumes that continental extinctions were like island extinctions. The extinction of large numbers of bird species following human colonization has been well documented for many island contexts (Anderson 1989; Olson and James 1982a, 1982b; Steadman 1995, 2006; Worthy and Hold­ away 2003). The island analogy part of the overkill model assumes that island extinctions are an appropriate model for continental megafaunal extinctions, particularly the extinction of North American Pleistocene megafauna. To evaluate the validity of this analogy, I compare patterns of extinctions and evidence for prey naïveté among both island and North American fauna. Since the overkill model plays a significant role in several conservation arguments, the soundness of the analogy has important consequences for conservation and wildlife management efforts and the validity of this particular instance of applied paleozoology. Ecological restoration in the form of successful rewilding of a species based on paleozoological data has taken place in New Zealand with the takahe (Porphyrio hochstetteri), a large flightless rail (Lee and Jamieson 2001). This rewilding and other examples of uses of paleozoological data in environmental management (e.g., Emslie 1987; Graham 1988; Lyman 2006; Wolverton et al. 2007) are overshadowed by the much-­publicized agenda to rewild North America with Old World mammals on the basis of the overkill hypothesis. Treated as fact, North American Pleistocene overkill also takes on a support role in general arguments about the impacts of humans on environments worldwide, such as the much-­publicized discussion of the Anthropocene (Steffen et al. 2007) and the Sixth Extinction (Eldredge 2001). I show here that such arguments are built on a house of cards.

Structure of the Island Analogy The Pleistocene overkill hypothesis states that about three dozen genera of North American megafauna went extinct at the end of the Pleistocene as a direct result of overhunting by humans (see Grayson 2006 and Surovell 2008, for evenhanded overviews from an overkill opponent and an overkill proponent, respectively). As Martin (1967) explains, researchers

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had been interested in the question of megafaunal extinction since the late 1800s, but most believed that climate change was responsible rather than anthropogenic causes (see Grayson 1980, 1984a, 1984b, for an extensive history of explanations of megafaunal extinctions). The development of radiocarbon dating in the 1950s provided rigorous temporal control so that the relationship between human colonization of the New World and the extinctions could be examined in fine detail. What the dating showed, according to Martin (1958, 1967, 1990), was that humans arrived in the Americas about the same time that the megafauna became extinct. Since the two events appear to have co-­occurred, Martin concluded that human predation, not climate change, caused megafaunal extinction in North America. Over the years, the details of the overkill hypothesis have changed to incorporate new data and to adapt to new criticisms (Grayson 1984b; Grayson and Meltzer 2003). And it has expanded beyond an explanation for Pleistocene extinctions in North America to become a depiction of what happened to megafauna worldwide (Martin 1984). Human hunting is no longer the only anthropogenic mechanism for extinction. Today, any anthropogenically driven environmental change (e.g., deforestation) is also invoked as a contributing cause of extinction, although human predation is still thought to play a prominent role in the process. With these expansions, the overkill hypothesis can be restated as follows: When people colonize new lands, their impacts (direct, e.g., hunting, or indirect, e.g., habitat modification) cause some faunal taxa to become extinct. One constant in the conceptual history of the overkill hypothesis has been the importance of island extinctions, which provide a graphic illustration of how humans can have significant impacts on native fauna that may culminate in extinction. In particular, avian extinctions in the Pacific have been used to illustrate the magnitude and speed with which human-­ caused extinctions can occur (Martin 1967; Martin and Steadman 1999), such that a “geologically rapid and biotically monumental extinction event will follow human establishment on a virgin land mass” (Martin and Steadman 1999:47, emphasis added), where the “virgin land mass” includes both islands and continents. In addition, the rapidity of the well-­ documented moa extinction in New Zealand led Martin (1967:105) to identify humans as a “superpredator.” Thus, island extinctions became the blueprint for the extinction process for North American Pleistocene megafauna. One statement about island extinctions validating the overkill hypothesis is particularly explicit: The late Pleistocene arrival of humans in North America, South America, and Australia is comparable to the late Pleistocene and late Holo-



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cene human colonization of islands. In both cases, vertebrate faunas unaccustomed to people were suddenly subjected to earth’s most ingenious predator. Whether large mammals on continents or birds on oceanic islands, most species were unable to cope with becoming the preferred prey of bipedal invaders, or at least were unable to survive other ecological changes brought on by people. (Steadman and Martin 2003:142)

The relationship between island extinctions and Pleistocene continental extinctions is structured as an analogical argument. The overkill hypothesis assumes that because extinctions occur in close temporal proximity after the arrival of people in both situations, then the magnitude of the extinctions and their causal mechanisms are similar. Specifically, the naïveté of the fauna and the impacts of humans are argued to be similar in both contexts; thus human arrival in an area causes extinctions that are rapid and “monumental,” to use Martin and Steadman’s term (figure 6.1). In island contexts, naive fauna that evolved in the absence of predators were vulnerable to the arrival of invasive species. Thus, when humans colonized islands, rapid large-­scale extinctions of native fauna often occurred. It is argued that the extinctions were also large in scale in North America; thus the process of extinction on the continent is reasoned, based on the analogy of temporal coincidence of human arrival and extinctions, to be similar to that on islands. However, unlike island

Figure 6.1.  Structure of the analogic argument comparing island extinctions to North America megafaunal extinctions.

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settings, North America had a large number of predators during the Pleistocene. Thus megaherbivores were not naive like island fauna. To accommodate this difference, megafauna are argued to have been naive to human predators who were also hyperefficient hunters capable of causing rapid large-­scale extinctions. In sum, the island analogy is used to assume that continental megafauna were susceptible to changes brought about by the introduction of new predators or invasive species. The megafauna were naive and the humans were hyperefficient predators. As a result, the degree of extinction, that is the rate of extinction, in continental settings should be significant and similar to that seen on islands. However, analogies rely on the appropriateness of the assumptions. That is, humans must be superpredators and fauna naive to accommodate the analogy and the overkill argument. Thus, if any part of the analogy equating islands and continents is weak, then the rest of the argument based on the analogy is also weak. In what follows, I evaluate the strength of the analogy by looking at the rate of extinctions in island and continental settings. In particular, if island extinctions are an appropriate analogy for what occurred on continents and specifically in North America, then the magnitude of extinction and the pattern in rates of extinction over time should be similar across both settings.

Patterns of Extinction In the overkill hypothesis, extinctions on islands and in North America are argued to be “monumental” in nature (Martin and Steadman 1999: 47). This first prediction requires determining if the magnitude and rate of extinctions in both settings were similar. For the Pacific Islands, David Steadman’s (1995, 2006) comprehensive work demonstrates that about 50% of Pacific Island birds became extinct after human colonization. The magnitude of extinctions on these islands is indeed very large. New Zealand and Hawaii have been well studied and provide a useful database on avian extinctions (Boyer 2008, 2009; Duncan and Blackburn 2004; Duncan et al. 2002; Holdaway et al. 2001; Olson and James 1982a, 1982b; Roff and Roff 2003; Worthy and Holdaway 2003). For this analysis, I use published data on terrestrial bird extinctions by Holdaway et al. (2001) and Holdaway (1999a) for New Zealand and by Boyer (2008) for Hawaii. These publications provide extensive data on extant and extinct species. For New Zealand, the data set is limited to taxa found on the three main islands and does not include species endemic only to the subantarctic islands within New Zealand’s political territory (table 6.1). For North America, mammal extinction data were derived from Turvey (2009) for



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extinctions that occurred during the Holocene (10,000 BP to today). The data for late Pleistocene mammal extinctions were gathered from Grayson (1991). The range in table 6.1 reflects a conservative and liberal estimate of the number of species that became extinct during the proposed timing of the arrival of people 12,000 to 10,000 radiocarbon years BP (~13,800–11,400 calendar years BP; Faith and Surovell 2009). In discussions of North American Pleistocene extinctions, the genus is the preferred taxonomic level because of ongoing debates over taxonomic validity of some species. The conservative estimate of the number of extinct Pleistocene megafauna assumes that there was only one species of megafauna per genus, and includes only those genera that have been securely dated to 12,000 to 10,000 radiocarbon years BP (Grayson 2007). The liberal estimate includes all proposed species for all megafauna genera that are assumed to have become extinct under the overkill model. To these numbers, the single extinction that occurred during the Holocene period was added for a range of 31 to 52 species that went extinct after human colonization of North America. For both archipelagoes, the percentage of birds that went extinct after the arrival of humans is high, with the extinction of 37% of terrestrial New Zealand birds and 70% of terrestrial Hawaiian birds (table 6.1). Compared to the island contexts, the magnitude of extinction is far less for North America, with only 7–13% of North American mammals going extinct after humans arrived. Thus, although there were a large number of extinctions in North America, the impact to the total taxonomic richness was much lower than for islands. If “species were unable to cope” with the arrival of humans (Steadman and Martin 2003:142), then we should also see similar patterns in extinction rates over time in both islands and continents. Evaluation of this second prediction of the analogy requires calculation of extinction rates, which for this study is the number of species extinctions per century. Data were analyzed by three temporal categories: prehistoric (pre-­ European contact), historic (premodern), and modern (table 6.2). For the islands, the prehistoric period consists of a range of ages because the exact dates for human colonization of most Polynesian archipelagoes are the subject of much debate (Graves and Addison 1995; Hunt and Holsen Table 6.1.  Comparison of the Magnitude of Species Extinction in New Zealand, Hawaii, and North America after Human Colonization

Number of extinct species Percentage extinct

New Zealand Birds

Hawaii Birds

N. America Mammals

40 37%

78 70%

31–52 7–13%

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Table 6.2.  Calculation of Species Extinction Rates Prehistoric Period

New Zealand Hawaii N. America

No. of Extinctions

Time Span

No. of Centuries

Extinction Rate Range

26 56 31–52

AD 1000/1200–1800 AD 200/800–1800 13,800 BP–AD 1600

6–8 10–16 134

3.3–4.3 3.5–5.6 0.2–0.4

No. of Centuries

Extinction Rate

Historic Period

New Zealand Hawaii N. America

No. of Extinctions

Time Span

14 22 5

AD 1800–2000 AD 1800–2000 AD 1600–2000

2 2 4

7.0 11.0 1.3

Modern Period

New Zealand Hawaii N. America

Endangered Species

No. of Centuries

Potential Extinction Rate

28 31 92

1 1 1

28 31 92

1991; Kirch and Ellison 1994; Prebble and Wilmshurst 2009; Spriggs and Anderson 1993). For the Hawaiian Islands, prehistoric human colonization is argued to have taken place sometime between AD 200 and 800 (Graves and Addison 1995). For New Zealand, colonization occurred between AD 1000 and 1300 (Holdaway 1999b; Wilmhurst et al. 2008). For the end of the prehistoric and the beginning of the historic period, AD 1800 is used. Significant European settlement did not begin until the late 1700s in New Zealand (Sinclair 1990) and early 1800s for Hawaii (Daws 1968). Finally, the start of the modern period is set at AD 2000. Determining the time span for the temporal divisions for the North American sample was simpler. Using the overkill model as a basis, it is assumed that people arrived in the Americas at 12,000 radiocarbon years BP (Grayson 1991; Martin 1967). The beginning of the historic period is set at AD 1600 since early British, French, and Spanish settlements were established around this time. As with the islands, the modern period starts at AD 2000. Thus, the prehistoric period spans about 13,400 calendar years and the historic period is 400 years in duration. To calculate the extinction rates for each region and period, the num-



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ber of species that went extinct during each time period was divided by the number of centuries the time period spans. For the islands, extinction rates for the prehistoric period span a range of values reflecting the range of colonization dates. Since the modern period is currently ongoing, data for the modern period were derived using the number of endangered species for each region instead of extinct species (IUCN 2010). Thus, data for the modern period represent a potential extinction rate that could occur over the next century. As such it may overestimate the future extinction rate, but it accurately reflects the effects of current human actions. Temporal trends in the extinction rate data vary across regions (figure 6.2). Overall, the patterns of extinction for the two islands are similar. The extinction rate doubles from the prehistoric to historic eras, with a large number of bird species endangered in the modern period. This pattern is similar to that seen in global-­level extinction data where species extinction correlates with human population density (Baillie et al. 2004; Cardillo et al. 2004; Ceballos and Erlich 2002). More people result in greater environmental impacts, including species extinction. Thus for the islands, it appears that the pattern of increasing rates of extinctions may be related to greater human population density on the islands. The extinction rate for North American mammals differs markedly from that of islands (figure 6.2). The extinction rate is much less than for the two island groups during prehistoric and historic times. And North

Figure 6.2.  Comparison of species extinction rates across time. The error bars for the prehistoric period for New Zealand and Hawaii represent the range of extinction rates depending on the initial colonization dates. Note also, for US mammals, the number of endangered species extends beyond the scale of this graph.

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America currently has a much larger number of mammal species that are endangered. However, only 25% of the mammal species in North America are endangered compared with nearly all of the native birds in Hawaii and about one-­third of New Zealand birds. Land mass size and thus large metapopulations likely mitigate extinction risk. More importantly, within the context of this discussion, the likelihood of species loss in North America is an order of magnitude greater during modern times than during previous times. This suggests that anthropogenically caused island extinctions are not an appropriate analog for continental extinctions. As presented in figure 6.2, the island extinctions involved an order of magnitude less time than the continental extinctions (table 6.2). Because the late Pleistocene extinctions in North America are believed to have occurred during a short time span rather than being spread over the entire Holocene, the data are separated and late Pleistocene extinctions are all assumed to have occurred in the restricted period between 12,000 and 10,000 BP (13,800–11,400 calendar years BP). The remainder of the prehistoric period spans from 10,000 BP to European colonization and is represented by the Holocene (figure 6.3). As discussed above, a range of extinction rates is shown for the Pleistocene because the number of species that are thought to have gone extinct during this period varies from 30 to 51 (table 6.3). In contrast, only one species went extinct during the Holocene. Even when the late Pleistocene data are considered separately,

Figure 6.3.  Comparison of extinction rate across time with the late Pleistocene megafaunal extinctions (n = 30–51) separated from Holocene extinctions (n = 1) for the US mammal sample. The error bars for the Pleistocene data represent the maximum and minimum extinction rate based on the range of estimates for the number of species that went extinct at the end of the Pleistocene.



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Table 6.3.  Extinction Rates for US Mammals during the Late Pleistocene and Holocene

Late Pleistocene Holocene

No. of Extinct Species

No. of Centuries

Extinction Rate

30–51 1

24 110

1.3–2.1 0.01

the extinction rate is still two to three times less than that for the prehistoric extinctions in Hawaii or New Zealand. In addition, unlike the islands where the extinction rate increased over time, during the Holocene the extinction rate in North America declined (figure 6.3). If the island analogy is appropriate, then we should expect a greater extinction rate in North America since human populations are known to have increased there during the prehistoric period. There are two pos­ sible explanations for the difference in rates between the islands and North America. One is that the nature of extinction varies contextually as do human–environment interactions. This means the extinction pattern (rate and magnitude) for North America should not be expected to be similar to that of islands. However, this undermines the valid use of the island analogy to support the overkill model. Another possible explanation is that there was some other cause for the extinctions at the end of the Pleistocene such that human population density was not a major contributing factor. This would challenge the overkill model particularly since the blitzkrieg corollary of the overkill model assumes that human populations increased significantly during the late Pleistocene. A large human population was required to hunt all species across the whole continent to extinction. However, if this occurred, the issue then concerns why this large prehistoric human population did not continue to have an impact on fauna through the Holocene.

Nature of Extinct Fauna on Islands and Continents When overkill proponents use the island analogy, they assume that since there are large-­scale extinctions in both island and North American contexts, the mechanism or cause for extinction must also be similar. Ecologists have long noted that island fauna are particularly vulnerable to extinction because they tend to be characterized by a high degree of naïveté, endemism, and K-­selected traits (Gaston and Blackburn 1995; McDowall 1969; McKinney 1997; O’Grady et al. 2004; Pimm et al. 1988, 1995). Naive fauna evolve in the context of few or no predators and develop physical and behavioral responses as a result of low predation pressure.

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Physical adaptations associated with naive fauna include flightlessness in birds, while behavioral adaptations can include ground nesting and a limited flight response. These traits increase the vulnerability of island fauna to new predators. In addition to naïveté, many island species have low reproductive rates and small finite populations; thus their populations require more time to recover from an increase in mortality rates. Given their ecological naïveté, K-­selected strategies, and limited population sizes, these species are particularly at risk for extinction related to the variety of environmental perturbations brought about by human colonization. Introduced predators, habitat alteration and loss, and human hunting can swiftly and significantly reduce populations of island species (Berglund et al. 2009; Blackburn et al. 2004; Karels et al. 2008; Sax and Gaines 2008). Relationships between these characteristics and extinction can be seen in the example of the extinct terrestrial birds of New Zealand and Hawaii. Almost all of the terrestrial birds that went extinct in New Zealand and Hawaii were endemic (table 6.4). The two island samples differ in that the extinct New Zealand birds were more likely to be flightless, ground nesting, and large bodied than Hawaiian birds. Because New Zealand has a much longer evolutionary history (80 million years) than the Hawaiian Islands (5 million years), the difference likely reflects distinctive evolutionary biology related to this time depth. While Hawaii and New Zealand have similar overall extinction rates, the causes of extinction appear to differ (table 6.5). In New Zealand, many of the bird species were recovered from archaeological contexts (McGovern-­Wilson 1986; Worthy and Holdaway 2003), while most Hawaiian bird species have been recovered from paleontological contexts (Olson and James 1982b). Thus, human predation appears to be more significant for extinctions among New Zealand birds (Duncan et al. 2002). Indeed the poster taxon for megafaunal overkill in island contexts Table 6.4.  Percentage of Extinct New Zealand and Hawaiian Birds That Were Endemic and Had Traits of Naive Fauna

Endemic Flightless Ground-nesting Large-bodieda

New Zealand Birds (%) N = 40

Hawaiian Birds (%) N = 78

100 80 73 82

96 33 43 92

Note: New Zealand data from Holdaway (1999a) and Holdaway et al. (2001); Hawaii data from Boyer (2006). aSpecies larger than 2 kg.



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is the moa (Order: Dinornithiformes). The hundreds of archaeological sites with moa remains provide significant evidence that moas were hunted and eaten by people (Anderson 1989). However, even though human predation is known to have played a significant role in moa extinction, researchers recognize that introduced animals and habitat loss were likely also important (Anderson and McGlone 1992; Cassels 1984; McGlone 1983; McGlone and Wilmhurst 1999; Worthy 1999). In addition to moas, 12 other extinct avian taxa are also found in archaeological sites in varying amounts (Worthy and Holdaway 2003). In contrast, for Hawaii, introduced predators, habitat loss, and possibly avian disease have been more significant contributors to bird extinction. The difference may be related to resource availability and subsistence patterns. While Polynesian communities such as those in Hawaii were agriculturalists, New Zealand Maori in the southern portion of the country, where moa populations were densest, were hunter-­gatherers. Thus, the harvest of large, naive, K-­selected birds by humans was a significant factor in New Zealand extinctions (Anderson 1989), while the indirect impacts of agricultural land use likely drove extinctions in Hawaii (Athens 1997, 2002). For islands, then, ecological naïveté, among other factors, created a situation where island fauna were particularly vulnerable to human-­ induced environmental change, which led to a high rate of extinction. Were North American megafauna vulnerable like island fauna? Overkill proponents argue that megafauna were naive specifically to a new type of predator—human (bipedal) hunters. This position presumes that humans were hyperefficient hunters such that they decimated populations at a rate faster than megafauna could learn to cope or reproduce, creating effects similar to what we see on islands (Wroe et al. 2006). Rebuttals to this argument have focused on whether naive prey can learn to deal with new predators (Grayson and Meltzer 2003; Wroe et al. 2006). For example, Berger et al.’s (2001) classic study on moose responses to visual and auditory predator cues demonstrates that naive moose who had not Table 6.5.  Comparison of the Causes of Extinction among New Zealand and Hawaiian Birds New Zealand Extinct bird species found in archaeological contextsa No. of introduced predators Degree of deforestationb

Hawaii

60%

8%

7 spp. ~75%

5 spp. ~80%

Data from Worthy and Holdaway (2003); Olson and James (1982b). Data from Athens (1997); McGlone (1983).

a

b

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experienced predation in decades could develop antipredator responses within one generation, and even after one encounter (see also Bshary 2001; Croes et al. 2007; Griffin et al. 2000). In regard to responses to humans, Bates et al. (2007) demonstrate that elephants can distinguish between human groups that pose a threat to them and those that do not based on scent and sight cues. Maasai men attack elephants with spears as part of a rite of passage, and when the clothing of the Maasai are seen or smelled, the elephants flee. Studies like these illustrate that it is possible for fauna to learn (or relearn) predator avoidance behaviors and avoidance of humans in particular, and they learn these behaviors quickly. However, since the fauna of North America were not truly naive because they existed among a large predator guild, it is useful to also examine the literature regarding the impacts of invasive (quadrupedal) predators on native fauna. The magnitude of impact of nonnative invasive predators depends on the nature of antipredator responses of the native prey (Sih et al. 2010). Prey may recognize predators based on specific or general cues and may use single or multiple cues. They can also have specialized or generalized responses to predators. To have a significant effect, the novel predator must be outside the realm of experience of the prey. Thus, fauna that recognize specific cues and have specialized responses are particularly vulnerable to new predators. Naive prey animals are vulnerable because, if they have antipredator adaptations, they are often to specific predators. For example, the kakapo (Strigops habroptilus) is a large flightless parrot endemic to New Zealand. Prehistorically, kakapos were preyed upon by several species of raptors that were sight hunters. Given these predators, the kakapo’s antipredator defenses include feather markings that act as camouflage and becoming motionless when threatened. As a result, the kakapo—with its unfortunately distinctive odor—has been ill equipped to deal with new introduced predators that hunt by scent, such as dogs (Hagelin 2004; Lloyd and Powlesland 1994). Because it has evolved specific adaptations to a particular type of predator, the kakapo is highly vulnerable to novel predators that rely on a strategy distinct from those to which the kakapo is adapted. The critical question therefore is: Were human hunters outside the realm of experience of Pleistocene megafauna? If these taxa had specialized rather than generalized responses, like the kakapo, they may have been particularly vulnerable to human predation. However, Pleistocene megaherbivores existed alongside several species of canids, felids, and ursids, which would have relied on a variety of predation strategies ranging from ambush to social-­cursorial hunting. As a result, they likely had multiple cues for recognizing predators based on sight, sound, and smell, as well as multiple responses for avoiding capture, such as social herd-



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ing behaviors, morphological defensive structures, and large body size (Caro 2005). Bischof and Zedrosser (2009) modeled the impact of new knowledge on prey mortality to understand the impact of learning in prey and of education in aiding survival of naive fauna. They found that educated prey have low mortality rates and slow extirpation rates. In addition, new knowledge related to predator encounters benefited populations of long-­ lived species more than those of short-­lived species. With short-­lived species, more naive individuals are born and the knowledge transmission period is shorter, so the proportion of educated individuals that survive is smaller. With long-­lived species, new knowledge can be transferred over a longer period of time to more individuals and thus have a greater population-­level impact. Long-­lived herding species such as mammoths would have particularly benefited from encounters with humans as educational experiences. Take, for example, the study by Bates et al. (2007) discussed above regarding elephants’ flight in response to the sight or scent of Maasai garments. They found that the elephants’ flight responses were not correlated with direct experience with spear attacks. Instead, it appears that information about the potential threat was transmitted among individuals. Importantly, learning and reduced prey naïveté did not differ between groups of elephants that had less than two experiences in the previous 20 years and groups with more experience. These findings suggest that even if encounters with human predators were rare, happened decades ago, and happened only to certain individuals, elephants in general still perceived the Maasai as a threat and fled. It also appears that contrary to assumptions in the overkill model, greater harvest pressure does not necessarily lead to higher mortality. Bischoff and Zedrosser (2009) demonstrated that with increased harvest pressure, more individuals become educated in light of greater exposure to the new predators, and mortality and extirpation rates declined. This is one of the reasons that extirpation of invasive or pest species is often difficult. Thus, with high harvest pressure on long-­ lived species, there should be an increasing proportion of educated individuals in the population and predator avoidance would quickly become a common strategy. The implications for the overkill hypothesis are significant. Given that most of the megaherbivores were herd animals and several were long-­ lived, whole groups would need to be wiped out at one time to minimize the learning and subsequent establishment of antipredator responses. In effect, people would truly have to be hyperefficient, mega-­superpredators to have hunted all the megafauna to extinction. Proponents of the overkill model have long argued that early human hunters in North America were very efficient, but this is an assumption with as yet minimal empirical support (Grayson and Meltzer 2003). Since it is a crucial component

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of the argument, the null hypothesis should be that the fauna adapted to the novel human hunters quickly. Any claims about hyperefficiency or superpredatory skills should be tested, not assumed, though how they might be tested archaeologically is not at all clear.

Overkill and the Island Analogy Relatively recent faunal extinctions on islands such as New Zealand, Hawaii, and Madagascar are regularly used to illustrate that human predation and anthropogenic landscape change can have a significant impact on native taxa. Several researchers have previously described some of the problems of using island extinctions as an analogy for continental extinctions (Grayson 2001; Grayson and Meltzer 2003; Wroe and Field 2006; Wroe et al. 2006). Burney and Flannery (2006:62), however, have argued that “[analogies] with remote island extinctions do not lie at the heart of human-­agency hypotheses. Of far more importance to the central argument is the coincidence in time of a particular type of extinction event and colonization by people, regardless of land area.” If the island analogy is not important to the model, then what we are left with is simply the timing of the two events—colonization and extinction. As such, this would be a classic example of a false cause (post hoc ergo propter hoc) argument, where outcome or effect B is caused by A simply because it follows A in time. Contrary to what Burney and Flannery argue, the island analogy is important to overkill precisely because it moves the argument beyond one of mere coincidence and correlation. Without the island analogy, the overkill hypothesis lacks a mechanism to explain the process. Interestingly, this is the same critique overkill proponents lay on the climate change explanation—no mechanism is specified to explain how the end of the Pleistocene led to the extinctions (Barnosky et al. 2004; Fiedel and Haynes 2004; Koch and Barnosky 2006). Another conceptual problem with the island analogy is that it treats all extinctions as monolithic and homogeneous: all processes and their attendant effects are the same across time and space. However, as this study shows, extinction patterns (processes and effects) vary across contexts. Extinction rates vary between islands and continents, and the role of human hunting in extinction varies even across islands. Overkill proponents expect the process and cause to be similar just because the outcome is extinction. Even the counterarguments against climate show similar lines of thought. For example, Burney and Flannery (2006) argue that climate change is not a viable explanation because not all extinctions (e.g., Australia, New Zealand, and Madagascar) occurred at the Pleistocene-­Holocene transition. Thus, because not all extinctions occurred at a time when cli-



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mate undisputedly changed, none of the extinctions could be caused by climate change (see also Martin 1990). In effect, overkill sets up a lawlike proposition: people arrive in an area (e.g., virgin lands) and cause mass extinctions. This is not only the simple fallacy of assuming a causal relationship exists between two merely (temporally) correlated events; it is antithetical to an evolutionary perspective and one of the reasons why Grayson (2007) has argued that it would be more useful to conduct a species-­by-­species examination for the causes of extinction.

Overkill and Rewilding While the cause of Pleistocene extinctions is the subject of major debate within archaeology and paleontology, within ecological and conservation biology circles, the Pleistocene overkill model is often portrayed as a fact. It has become an important piece of evidence that is used to set benchmarks for ecological restoration, the most provocative example of which is Pleistocene rewilding (Donlan 2007; Donlan et al. 2005, 2006; Martin 2005). Rewilding in a broad sense is an attempt to restore ecosystems to their former glory and peak biodiversity and can include reintroducing keystone taxa, such as top carnivores or herbivores that have been locally extirpated by humans in particular (Foreman 2004; Fraser 2009; Nicholls 2006; Soulé and Noss 1998). Restoration benchmarks have typically been set at the premodern, preindustrial, or pre-­European era depending on arguments about what is wild or natural or pristine (Hunter 1996). In the case of Pleistocene rewilding, overkill is used to argue that the proper benchmark for North America is the prehuman Pleistocene environment. As such, the modern North American landscape should be repopulated with descendants and relatives of Pleistocene fauna, thereby restoring the continent’s taxonomically depauperate fauna. The logic is that humans have had a significant impact on biodiversity in North America since they arrived, starting with megafaunal overkill. Since people are argued to have been the sole cause of the Pleistocene extinctions, it is also argued that it is our moral and ethical imperative to try to restore what we destroyed. While a number of researchers have pointed out reasons why Pleistocene rewilding is not ecologically viable (Caro 2007; Caro and Sherman 2009; Dinerstein and Irvin 2005; Oliveira-­Santos and Fernandez 2010; Rubenstein et al. 2006; Schlaepfer 2005; Shay 2005; Smith 2005), it is more fundamentally problematic simply because of its reliance on the as-­yet not strongly confirmed overkill hypothesis (Wolverton 2010). Because the validity of the island analogy is questionable, along with other alleged support of the overkill hypothesis, Pleistocene rewilding as a viable conservation strategy is dubious. If ecologists accept overkill, they must also be

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willing to accept that evolutionary and ecological processes related to extinction on continents are similar to those on islands. In effect, though provocative, Pleistocene rewilding taints both rewild­ ing as a conservation strategy and the use of paleozoological data in wildlife management. For an example of a successful combination of both, I turn to New Zealand, where zooarchaeological and paleontological data are often used as just another type of data to understand long-­term population trends and distributions of native fauna (e.g., Beauchamp and Worthy 1988; Boren 2010; Lalas and Bradshaw 2001). An excellent example concerns the conservation of the takahe (Porphyrio hochstetteri). Takahe are large flightless rails that were believed to be extinct until their rediscovery in the subalpine tussockland of the Murchison Mountains of Fiordland in 1948 (Smith 1952). Given their modern distribution, it was believed that takahe were tussock grassland specialists, feeding only on the tender bases of tussock grass. Since the distribution of tussock grassland had decreased during the historic period due to conversion to pastureland, conservation efforts initially focused on increasing the Fiordland population by enriching the tussockland, as well as eradicating predators and competitors, minimizing human disturbance, and hand-­ rearing chicks (Lee and Jamieson 2001). However, in the 1980s, paleozoological data provided a different view of takahe habitat preferences. Prehistorically, takahe were once widespread across New Zealand, and their distribution suggests that they preferred edge habitat and would have done well in forest-­shrubland-­grassland mosaics, refuting the notion that they are tussockland specialists (Beauchamp and Worthy 1988; McGovern-­ Wilson 1986; Trewick and Worthy 2001). Paleozoological data indicate the takahe population in Fiordland is a relict and the Murchison Mountains are merely the last inhospitable refuge from the assault of introduced competitors and predators. This new information pushed takahe conservation in a new direction. Because takahe were no longer considered tussockland specialists, conservation efforts could broaden to include other strategies beyond maintaining the Fiordland population. One option was to translocate takahe to other areas of New Zealand—to “rewild”—to create a geographically extensive metapopulation. At the time, significant effort had been going into ecological restoration of offshore islands by eradicating predators and replanting native vegetation to create “rewilded” refuges for endangered native species (Atkinson 1988). The first takahe were translocated onto offshore islands in the 1980s and over the last 30 years the takahe population has doubled from about 120 to 250 individuals, with about 40% of the population living on four offshore islands. As with the Pleistocene overkill and rewilding contention, paleozoological data were an important part of the argument supporting a rewild-



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ing strategy for takahe conservation. However, unlike the situation with North American Pleistocene megafauna, in the case of the takahe, there is a clear link between the decline in takahe populations and human activities. Of the 85 paleozoological sites that contain takahe remains, 30 are archaeological sites, most of which are midden sites (see Beauchamp and Worthy 1988 for a list of sites). Thus, the link between human predation and takahe decline is much greater than what can be convincingly demonstrated for North American megafauna. The evidence for overkill would be much stronger if these kinds of associations were more common (Grayson and Meltzer 2003; Wolverton et al. 2009). Nevertheless, even in the case of takahe (and the better-­known case of moas), introduced predators and habitat loss were also important in its decline, and they are certainly the major causes of decline in the period shortly after their rediscovery (Bunin and Jamieson 1995). Thus, compared to Pleistocene rewilding, takahe translocation has merit because the population decline was indisputably anthropogenically caused and because there is a viable extant source population for rewilding as well as a clear target environment created by offshore island ecological restoration.

Management Implications The problems of using the Pleistocene overkill hypothesis in conservation strategies are not limited to the argument for rewilding North America. Instead, the problem is much broader because using the overkill model as a warrant for wildlife management policies injects a human-­nature philosophical perspective into conservation practices that may be of limited utility (Rozzi 1999). Pleistocene overkill assumes that as a species, humans have a negative impact on nature. As such, humans and culture are in direct opposition to nature and have been so since the beginnings of Homo sapiens as a species. An excellent example of the problem with this conceptual approach can be seen by comparing discussions about the timing of the Sixth Extinction and the Anthropocene. The Sixth Extinction refers to the latest period of mass extinctions (Barnosky et al. 2011), which is characterized by extinctions related to anthropogenic actions (Leakey and Lewin 1996). The postulated beginnings of this period can be as recent as the modern industrial period or as far back as the origin of humans (Eldredge 2001). The early benchmark is based mainly on Pleistocene overkill and takes the position that humans caused environmental change as soon as they arose as a species and began to colonize new lands. The Anthropocene—a newly proposed geological-­anthropological epoch conceived to occur after the Holocene (Steffen et al. 2007)—was

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originally used to mark the modern period of human impacts and global warming effects, in particular those of greenhouse gas emissions (Crutzen 2002; Crutzen and Stoemer 2000). Ruddiman (2003) has argued, however, that the start date should be pushed back to the beginnings of agriculture (currently estimated to be about 10,000 years ago) because there was a significant increase in carbon dioxide (CO2) and methane (CH4) emissions associated with farming and herding prior to the Industrial Revolution. While the Sixth Extinction and the Anthropocene are similar in that both attempt to specify periods of human-­induced environmental change, they differ in how they determine start dates for the periods. For the Anthropocene, the debate is an empirical one. The question is, when do human activities significantly increase greenhouse gas emissions? Ruddimann (2003) sees a significant increase when agriculture becomes a prominent subsistence strategy. Crutzen and Steffen (2003; Steffen et al. 2007) argue that the increase in greenhouse emissions related to the Industrial Revolution in the late 1800s is more significant. The disagreement is about which scale is appropriate, but both parties present empirically based arguments. In contrast, discussions about the beginning date of the Sixth Extinction are not empirical but rather philosophical. Humans are beyond or separate from nature because our cultural practices, whether they are fire, agriculture, or the production of toxic industrial waste, have significant impacts on the ability of other species to survive. Since delineation of the Sixth Extinction is tied to philosophy, then the agenda determines the definition (see Callicott and Nelson 1998 for a review of a similar debate related to the use of the wilderness concept in environmentalism). This approach may be useful as a point of discussion but has limited practical utility in environmental management. Given their differences, it should be no surprise that the Anthropocene is being considered by geologists as a new epoch (Zalasiewicz et al. 2010), while the Sixth Extinction remains simply a rallying cry among environmentalists. Concepts such as the Sixth Extinction may be more useful if, like the Anthropocene debate, we examine the issue of human contributions to extinction from an empirical perspective. For example, Callicott (2007) argues that setting spatially uniform restoration benchmarks based on perceptions about what is pristine or natural is problematic (Frazier 2010). Pre-­Columbian environments are often advocated as benchmarks in the United States because of conceptions about European (major and significant) versus Native American (minor and insignificant) impacts. However, both European and Native American settlement and impacts varied across time and space. As a solution, Callicott advocates that we



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take a more empirical approach to setting benchmarks by asking the question, when did the effects of human impact outpace previously unimpacted rates of evolutionary change? In other words, in terms of extinction, we can examine when human-­influenced extinction rates surpass those of background extinction as a means of determining when human impacts became significant. Thus, unlike the Anthropocene, for which the question of “how much is too much?” is being debated, the background extinction rate provides the scale beyond which extinction is “unnatural.” Using this comparative method provides an empirical means for establishing local restoration benchmarks or even global-­scale benchmarks such as the beginning of the Sixth Extinction. Compared to using Pleistocene overkill to validate some preconceived notion about human impacts, using paleozoological data to identify variation and change in extinction rates can make a significant contribution to conservation efforts (Hadly and Barnosky 2009). As an example, we can compare the pattern of extinction rates for Hawaii, New Zealand, and North America (table 6.2) to estimated natural background extinction rates to determine possible restoration benchmarks. Background extinction rates have been estimated to be about 1 species per 100 years (Baillie et al. 2004). Extinction rates for Hawaii and New Zealand are clearly beyond what is thought to occur naturally. However, for North American mammals, only modern extinctions can reliably be argued to occur at a rate greater than the background rate. On the one hand, we could argue that for Hawaii and New Zealand, humans had a significant influence on species extinctions as soon as they arrived in those archipelagoes. For North America, on the other hand, it is only during modern times that the potential extinction rate clearly increases beyond the background rate. More importantly, the decline in these endangered species can be linked to human activities. The situation for late Pleistocene megafaunal extinctions is less clear. The extinction rate varies from about the background extinction rate to twice that rate. The higher extinction rate is based on the assumption that ­all megafauna became extinct during this period. The more conservative estimate includes only megafauna that have secure terminal dates between 12,000 and 10,000 BP (Grayson 1991, 2007). Thus, it is important to clarify that the lower extinction rate estimate is based on empirical evidence and the higher estimate is based on an assumption. In addition, overkill also assumes that all extinctions were due to anthropogenic causes because of their association with the arrival of humans in the Americas. However, few of the megafaunal extinctions have been clearly linked to anthropogenic causes comparable to those seen in island settings (Grayson 2001; Grayson and Meltzer 2003). Thus, any conserva-

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tion or management proposal, such as Pleistocene rewilding, that incorporates Pleistocene overkill as part of its argument is based to some extent on this assumption rather than on empirical evidence.

Conclusion The concept of Pleistocene overkill has taken on a life of its own outside of paleontology and archaeology. In conservation biology and restoration ecology, it has become an important anecdote about how long humans have had a significant impact on nature. As Wolverton et al. (2009:35– 36; see also Grayson 2007) have noted, overkill is an inherently archaeological hypothesis and requires testing with archaeological data such as documenting direct associations between humans and all extinct megafaunal taxa (fewer than half currently are well associated with humans), detailed demographic analysis of prey taxa, and so on (e.g., Grayson and Meltzer 2003; Hill et al. 2008). Researchers outside of the discipline are not expected to be able to rigorously evaluate the evidence within another discipline’s debate. But as this study has illustrated, the Pleistocene overkill hypothesis can be evaluated based on the ecological assumptions it makes. By using the overkill hypothesis in conservation arguments, ecologists and conservation biologists should be aware that they are accepting a hypothesis that assumes islands are appropriate analogies for continents and that human impacts are uniform across space and time (and particular taxa). Grayson and Meltzer (2003:590) suggest that researchers who utilize the overkill hypothesis are doing so not based on its empirical strengths but because it has “political capital” and provides an excellent “homily of ecological ruin.” Insofar as this prompts serious thought about how modern society alters earth’s ecology and how we should change our behaviors, it is good. But it certainly does not make the Pleistocene overkill hypothesis correct. References Anderson, A. J. 1989. Prodigious Birds. Cambridge University Press, Cambridge. Anderson, A. J., and M. S. McGlone. 1992. Living on the Edge—Prehistoric Land and People in New Zealand. In The Naive Land, edited by J. Dodson, pp. 199–241. Longman Cheshire, Melbourne. Athens, J. S. 1997. Hawaiian Native Lowland Vegetation in Prehistory. In Historical Ecology in the Pacific Islands, edited by P. V. Kirch and T. L. Hunt, pp. 248–270. Yale University Press, New Haven, CT. ———. 2002. Avifaunal Extinction, Vegetation Change, and Polynesian Impacts in Prehistoric Hawai’i. Archaeology in Oceania 37:57–78.



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CHAPTER SEVEN

Paleozoological Stable Isotope Data for Modern Management of Historically Extirpated Missouri Black Bears (Ursus americanus) Corinne N. Rosania

Prior to the twentieth century, black bears (Ursus americanus) roamed the North American landscape in abundance. Their range reached as far south as northern Mexico and as far north as Alaska (Pelton 2000). But as human population soared to unprecedented numbers during the Industrial Revolution, black bear populations began to dwindle. Humans expanded across the landscape and intensified resource extraction, contributing directly and indirectly to black bear attenuation (McKee 2005; Orians and Soulé 2001; Pelton 2000). Humans intensively hunted black bears for fur and reduced viable habitats in which black bears forage, hibernate, and mate (Pelton 2000). Many populations shrank or disappeared. When faced with ecological change, black bears are impressively plastic. As omnivores they can subsist on a variety of plant and animal tissues. Diet, body size, and hibernation (i.e., duration and location) vary between populations living in different habitats and climates. While this flexibility has enabled black bears to persist in certain regions, they have not been so successful in the state of Missouri. Both paleozoological (zooarchaeological and paleontological) and his-

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torical data indicate black bears were widespread in Missouri prior to the twentieth century (Graham and Lundelius 1994; McKinley 1962). By the end of the nineteenth century, more than 98% of Missouri’s hardwood forests had been cut (Korte and Frederickson 1977), thereby depleting prime black bear habitat. By about 1930 black bears were thought to be extirpated in Missouri (Schwartz and Schwartz 1981), though some suggested a small relict population might remain in the Missouri Ozarks. Whether or not that relict population did indeed exist may be difficult to determine because there is now a population in southeastern Missouri that could be descendants of the relict population of black bears, that immigrated from a population transplanted to Arkansas between 1958 and 1968, or of a combination of the two (Smith and Clark 1994). Analysis of ancient DNA from some of the Lawson Cave bears discussed in this chapter has failed to clarify this matter (Hudson 2009). Whatever the case, management of twenty-­first-­century Missouri black bears is a subject of some interest (Etling 2000; Hudson 2009; Rosania 2010; Wolverton 2008). In particular, what can be done to ensure their survival? The easy answer is to determine appropriate habitats for black bears and then to manage the landscape so as to create (if necessary) and to maintain those habitats. Unfortunately, the extirpation of black bears from the modern landscape prior to study of their local ecology means that we know very little about native Missouri black bears. To learn more about their behavior and ecological requirements, in relation to habitat range, mate accessibility, and diet, modern studies have focused on extant populations, including the recently transplanted populations in Arkansas (Smith and Clark 1994) and the small population in southeastern Missouri (Schwartz and Schwartz 1981). Despite the knowledge gleaned from these studies, we still know very little about black bear behavior in the Midwestern United States (Garshelis et al. 2008). This means that any wildlife management or conservation actions will at best be based on information from other populations in distant regions. Fortunately, information gleaned from paleozoological remains can expand our knowledge of local black bear ecology by elucidating the behavior and ecological requirements of pre-­and periextirpation black bears in Missouri (e.g., Wolverton 2008). Paleozoology can indicate where black bears existed on the landscape and also suggest size of populations. Additionally, paleozoology can indicate what was available on the landscape for black bear consumption and, particularly, what materials were included in their diet. Dietary data will be important for identifying suitable habitats for management of resident black bear populations. We know that black bears are flexible consumers, but research to facilitate future conservation programs should concern questions about what native black bears should eat (or what they did eat prior to historic-­period and modern anthropogenic effects), and not what they often do eat (i.e., human refuse in human modified landscapes) (Beck-



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mann and Berger 2003). Due to extirpation of Missouri black bears, we have no modern reference data or benchmarks with which to assess native Missouri black bear diet. Thus any sort of conservation, management, or restoration activity must assume that native Missouri black bears had the same ecology as extant extralocal (non-­Missouri) black bears. Fortunately, analysis of paleozoological remains can answer some questions about native Missouri black bear diet and habitat use. How might such data be gleaned from paleozoological remains? For over 30 years, geologists, anthropologists, paleozoologists, and zoologists have used stable isotope signatures recorded in organism tissues to study prehistoric and modern human and animal populations. Stable isotopes reflect diet, climate, migration, and weaning age. In this study I examine stable carbon and nitrogen isotopes in a paleontological assemblage of 200-­to 600-­year-­old black bear remains from central Missouri in order to determine native Missouri black bear diet. Specifically, I sought answers to three questions. First, are there observable differences in diet among paleontological omnivores, herbivores, and carnivores from central Missouri? Second, did native black bear diet change through time in Missouri? And third, do native late Holocene paleontological bears differ from modern Missouri black bears in terms of diet?

Background Few paleozoologists with conservation interests have focused on stable isotope data from mammalian species. One is Hughes (2004), who used stable carbon and nitrogen isotope analysis to identify changes in northwestern Wyoming bighorn sheep (Ovis canadensis) abundance, distribution, and migratory behavior during the Holocene. She found that a reAmerican cent shift in migratory behavior that occurred after Euro-­ contact was likely due to the reduction in bighorn migratory range caused by modern human development. Etnier (2004) and Newsome et al. (2007) used stable carbon and nitrogen isotope analysis to identify changes in patterns of northern fur seal (Callorhinus ursinus) behaviors that resulted from commercial fishing and the fur trade. They concluded that modern migratory and breeding behaviors of northern fur seals are artifacts of anthropogenic impacts initiated in the nineteenth century. These studies highlight the value of stable isotope analysis of paleozoological remains for purposes of conservation ecology.

Black Bear Ecology Modern North American black bears persist in forested regions of the northern and eastern portions of the United States, as well as in small

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relict populations throughout the country (Jones and Birney 1988; Lari­ vière 2001; Pelton 2000; Schwartz and Schwartz 1981; US Department of the Interior and US Fish and Wildlife Service 1995; Whitaker and Hamilton 1998). Has native black bear behavior changed over the past several hundred years as a result of Euro-­American settlement? Have habitat use and diet changed? If so, how have these behaviors changed? Extant black bears exhibit exploded polygyny, a mating system wherein one male mates with several females. This mating system influences their residential patterns and thus their habitat use. Females occupy regions based on resource availability (i.e., they exploit large territories in less resource-­rich areas), and males roam much larger territories, which encompass the ranges of multiple females (Landers et al. 1979; Larivière 2001; Whitaker and Hamilton 1998). As a K-­selected species, extant black bears live at or near the carrying capacity of the environment; they have few offspring that mature slowly, thus allowing significant maternal investment during the first two years of life (Doan-­Crider and Hellgren 1996). Therefore, primary productivity strongly influences black bear behavior, including residential and reproductive strategies, because it influences what is available for consumption (Mosnier et al. 2008). Modern black bears are opportunistic feeders, even foraging in human refuse (Beckmann and Berger 2003). Though black bears do consume meat, usually carrion, they have been described as the most herbivorous species of the genus Ursus; plant material usually constitutes the majority of their diet (Boileau et al. 1994; Irwin and Hammond 1985; Mosnier et al. 2008; Schullery 1986). Commonly consumed plants include berries, nuts, grasses, and vegetables (Mosnier et al. 2008). Analysis of stable isotopes in an organism’s tissues does not allow one to determine the actual constituents of a meal, but instead indicates the relative proportions of meat and vegetation. It also allows distinction between two groups of plants (C3 and C4). Therefore, analysis of stable isotopes can be used to determine if late Holocene black bears in Missouri were as vegetarian as their modern counterparts. And if so, it can also reveal which of the two general types of plants they consumed. This information also provides a test of the assumption that native Missouri black bears had the same ecology as extant extralocal (non-­Missouri) black bears for which behavioral ecology has been studied in detail.

Stable Isotope Analysis Simply put, the old proverb “you are what you eat” applies quite literally to stable isotope research. As resources are consumed, stable isotope concentrations in food are absorbed and incorporated into tissues of the or-



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ganism; therefore, the stable isotope value of the organism reflects that of the food resource (plus an isotope-­tissue-­specific fractionation factor). The fractionation factor for bone collagen is 5‰ to 6‰, indicating that bone collagen is enriched in 13C, and thus has δ13C values 5‰ to 6‰ higher than those found in dietary constituents (Bumsted 1983; van der Merwe 1982). Stable carbon isotope values indicate the relative contribution of C3 and C4 plants to an organism’s diet (DeNiro and Epstein 1978). These two categories of plants (C3 and C4) are distinguished by differences in the pathways utilized during photosynthesis. In comparison to C3 plants (herbaceous flowering plants), C4 plants (grasses and some forbs) incorporate a carbon dioxide transport mechanism (Hatch-­Slack pathway) that improves their metabolic efficiency in hot, dry, and/or low carbon dioxide environments. As a result of this modification to the conventional C3 pathway (Calvin-­Benson cycle), C4 plants take in a relatively greater proportion of heavy carbon isotopes during photosynthesis and therefore exhibit higher δ13C values, ranging between −14‰ and −10‰, than C3 plants, which have δ13C values between −30‰ and −22‰ (for further discussion of photosynthetic pathways see Bender 1971; Cerling et al. 1993, 1997; Farquhar 1983; Farquhar et al. 1989; Molles 2002; Winter et al. 1976). These nonoverlapping stable carbon isotope values permit researchers to determine the relative contribution of C3 and C4 plants to an organism’s diet by examining δ13C values in the organism’s tissues. Stable nitrogen isotopes indicate the trophic level at which an organism fed (DeNiro and Epstein 1981; Minagawa and Wada 1984). Plants absorb nutrients (and isotopes) through the air and soil and incorporate them into their tissues. When a producer is consumed, the nitrogen content enters the consumer’s body and is subsequently fractionated by +3.4 ± 1.1‰ because of selective use of the lighter isotope (14N) in metabolic reactions (Minagawa and Wada 1984). Therefore, each consecutive trophic level in a given environment is enriched in 15N (relative to 14N) by approximately 3‰. Consequently, stable nitrogen isotope values in the tissues of carnivorous animals are markedly higher than those of herbivores and producers.

Methods and Materials I performed stable carbon and nitrogen isotope analysis of bone collagen to determine native black bear diet in central Missouri. Stable carbon and nitrogen isotope values in collagen indicate an average diet proportional to the organism’s rate of bone tissue replacement (i.e., bone remodeling rate). The annual turnover rate for cortical bone is approxi-

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mately 2.5% (Libby et al. 1964; Pate 1994). Three kinds of analyses were done to evaluate black bear diet: 1. Stable carbon and nitrogen isotope values for collagen from 10 native Missouri black bears dating to 200–600 years ago were compared to collagen isotope values for other taxa from the same paleontological deposit. 2. Stable carbon and nitrogen isotope values and relative geological ages of the individual paleontological black bears were compared to identify temporal patterns (if any). 3. Stable carbon and nitrogen isotope values for the paleontological bears were compared to isotope values for modern (relict/transplanted) Missouri black bears. To ensure data quality, three bone collagen samples were extracted from each of 10 Ursus americanus individuals recovered from Lawson Cave in Boone County, Missouri (figure 7.1). Lawson Cave is a natural trap cave located in Three Creeks Conservation Area approximately 8 kilometers (5 miles) south of Columbia, Missouri (Wolverton 1996, 2001). The assemblage is curated in the Department of Anthropology Zooarchaeology Laboratory at the University of Missouri–Columbia. It consists of remains of 16 taxa, including U. americanus (North American black bear), Marmota monax (woodchuck), Didelphis virginiana (opossum), Canis spp. (dog or coyote), Sus scrofa (domestic pig), and Sylvilagus floridanus (cottontail rabbit) (Wolverton 1996). Samples were obtained from 10 right ursid femora, the most abundant skeletal element in the deposit and indicating the minimum number of individual black bears in the Lawson Cave assemblage (Wolverton 2008). A majority of these individuals were young adult males who likely fell victim to the cave while foraging for food (Wolverton 2001, 2008). A series of six radiocarbon dates on ursid remains indicates their average age is about 200 years BP, though there is some variability (table 7.1; Rosania 2010; Wolverton 2001). Fluorine concentrations determined by neutron activation analysis suggest the relative order of deposition of all ursid individuals recovered from Lawson Cave (Rosania 2010). Fluorine is absorbed by bone through groundwater; thus, concentrations of fluorine in bones from the same assemblage can be used to determine their relative order of deposition (Lal 1975; McConnell 1962; Van de Water 1953). The measured fluorine values for all 10 individual Lawson Cave ursids do not contradict the previous age rank assigned to several individual bears by radiocarbon dates (dates range between 170 and 630 radiocarbon years before present) (Rosania 2010; Wolverton 2001). Therefore, fluorine concentrations are used here as a proxy measure of time.1



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Figure 7.1.  Geographic regions of Missouri. Lawson Cave is located in Boone County. Modern U. americanus specimens were surface collected from Cedar County and Reynolds County. Spatial data available at Missouri Spatial Data Information Service (2002) and US Environmental Protection Agency (2010).

Bone specimens of herbivores, omnivores, and carnivores from Lawson Cave were analyzed in order to rank the Lawson Cave U. americanus on a continuum of omnivory. Three separate samples were taken from each of 14 skeletal specimens of nonursid taxa to monitor replicability. Twelve of these specimens were from Lawson Cave, including one D. Table 7.1. Radiocarbon Dates (14C Yr BP) on Bear Bones from Lawson Cave Reference Number AA84746 AA84747 AA84748 AA38931 AA38932 CAMS 27141

Age 229 ± 41 206 ± 41 630 ± 42 233 ± 39 207 ± 34 170 ± 60

Source Rosania 2010 Rosania 2010 Rosania 2010 Wolverton 2001 Wolverton 2001 Wolverton 2001

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Table 7.2. List of Specimens from Which Collagen Was Extracted for Stable Isotope Analysis Trophic Level Omnivores

Species Ursus americanus Ursus americanus Ursus americanus

Herbivores Carnivores

Didelphis virginiana Sus scrofa Marmota monax Sylvilagus floridanus Canis spp. Lynx rufus

Common Name North American black bear North American black bear North American black bear Opossum Pig Woodchuck Cottontail rabbit Dog/coyote Bobcat

N

Location

1

Cedar Co., MO

1

Reynolds Co., MO

10

Lawson Cave, MO

1 4 2 4 1 2

Lawson Cave, MO Lawson Cave, MO Lawson Cave, MO Lawson Cave, MO Lawson Cave, MO Oklahoma

virginiana, four S. scrofa, four S. floridanus, two M. monax, and one Canis sp. (table 7.2). Four additional specimens, two Lynx rufus (bobcat, from Oklahoma) and two modern U. americanus (from Cedar County and Reynolds County, Missouri) were selected to establish the carnivorous end of the continuum and to provide a comparative baseline for evaluating the Lawson Cave ursids, respectively. Sample sizes were limited by three factors: (1) the paleontological nature of the Lawson Cave assemblage limited the number of individual organisms and the number of taxa; (2) black bear extirpation prior to modern analyses limited the number of modern Missouri black bear skeletons available for consideration; and (3) population size and distribution of large carnivores on the landscape are constrained by ecological requirements, making them difficult to sample (i.e., large carnivores require large quantities of space and calories). Laboratory protocols, analytical techniques, and resolution measures are described elsewhere (Rosania 2010).

Results Recall that I sought answers to three questions. First, do stable isotopes indicate trophic-­level differences in diet among Lawson Cave taxa? Second, did native black bear diet change through time, as reflected in the Lawson Cave ursids? And third, did Lawson Cave black bear diet differ from that of modern Missouri black bears?



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Trophic-­Level Differences Carnivores have the highest stable nitrogen isotope values, followed by omnivores and then herbivores (figure 7.2). Modern Ursus americanus have stable nitrogen isotope values greater than those of Lawson Cave herbivores, but slightly less than those of Lawson Cave omnivores. Stable nitrogen isotope values from Lawson Cave U. americanus are similar to modern U. americanus values, but are slightly more 15N-­heavy (figure 7.2).

Figure 7.2.  Box and whisker plot of stable nitrogen isotope values for herbivores, omnivores, carnivores, and modern and Lawson Cave ursids. Central marks are medians; box edges are the 25th and 75th percentiles; whiskers extend to samples within one standard deviation of the central mark; and open circles mark extreme outliers (greater than two interquartile ranges from the central mark).

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Herbivores have the lowest stable carbon isotope values, followed by omnivores and carnivores, which have similar average values (figure 7.3). Both modern and Lawson Cave U. americanus have stable carbon isotope values similar to those measured for Lawson Cave omnivores and carnivores (figure 7.3). All specimens analyzed in this study (including modern and Lawson Cave U. americanus) fall within the range expected for C3 consumption (−25‰ and −17‰), except the canid and two pigs, which had carbon isotope values between −17‰ and −9‰. No specimens had carbon isotope values below −9‰. Therefore, all animals analyzed here consumed primarily C3 materials (i.e., C3 plants and consumers), except

Figure 7.3.  Box and whisker plot of stable carbon isotope values for herbivores, omnivores, carnivores, and modern and Lawson Cave ursids. Central marks are medians; box edges are the 25th and 75th percentiles; whiskers extend to samples within one standard deviation of the central mark; and open circles mark extreme outliers (greater than two interquartile ranges from the central mark).



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the canid and two pigs; these individuals likely consumed mixed C3-­C4 diets. Some C4 materials could have contributed to each individual’s diet, but remember that bone collagen records a long-­term average, and thus records the isotope signature of the primary dietary constituents. A Kruskal-­Wallis nonparametric comparison of Lawson Cave herbivore, omnivore, and carnivore stable isotope values identified differences in stable carbon (H = 10.77, df = 2, p = 0.005) and stable nitrogen (H = 13.59, df = 2, p = 0.001) isotope values among the three trophic groups. Two-­sample Mann-­Whitney nonparametric comparisons indicate where differences exist among the three groups. Stable carbon isotope values for herbivores and omnivores are significantly different (U = 9, n1 = 6, n2 = 17, p = 0.002). Stable nitrogen isotope values for herbivores and omnivores are also different (U = 6, n1 = 6, n2 = 17, p = 0.001). Significant differences in stable nitrogen isotope values are also apparent between carnivores and herbivores (U = 0, n1 = 3, n2 = 6, p = 0.010). Thus, herbivores and omnivores consumed different proportions of C3 and C4 vegetation, and as expected, herbivores, omnivores, and carnivores exhibited significantly different stable nitrogen isotope values. Mann-­ Whitney nonparametric comparisons of herbivore and ursid stable isotope values for the Lawson Cave sample were used to investigate whether the two groups were distinguishable. Both stable carbon (U = 6, n1 = 6, n2 = 10, p = 0.005) and nitrogen (U = 4, n1 = 6, n2 = 10, p = 0.002) isotope values for U. americanus from Lawson Cave were different than those of Lawson Cave herbivores. Lawson Cave U. americanus and contemporary herbivores had different diets; the ursids consumed a different proportion of C3 and C4 plants and more meat than the herbivores (figure 7.4). Therefore, the answer to my first question is yes; there are observable differences in the diet of Lawson Cave carnivores, herbivores, and omnivores. Lawson Cave ursids are distinguishable from contemporaneous herbivores. Thus the assumption that native Missouri black bears are no different from extralocal modern black bears in terms of position in the trophic pyramid is confirmed, at least insofar as the coarse resolution provided by stable isotopes is concerned.

Temporal Variation in Ursid Diet Rank-­ordered stable carbon and nitrogen isotope values for Lawson Cave U. americanus were compared with fluorine concentrations from the same skeletal specimens to identify temporal patterns. Spearman’s ρ for both stable carbon (ρ = −0.27) and nitrogen (ρ = −0.43) isotope measurements indicated no significant correlation between the isotopes and fluorine concentrations. Thus, the answer to my second question is no; there is no clear change in U. americanus diet (C3 and C4 plant and meat

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Figure 7.4.  Stable carbon and nitrogen isotope values for carnivores, herbivores, omnivores, and modern (MOU) and Lawson Cave (LCB) black bears. Each point is an average for each individual based on three replicates. Error bars for each sample are based on standard deviations and display the variation among the three replicates.

consumption) through time, but rather stasis. This result indicates that at the coarse scale provided by analysis of stable isotopes, native Missouri black bear diet did not fluctuate during the 400 or so years (based on available radiocarbon dates) represented by the Lawson Cave remains. Because there is no difference between the isotopes of the paleontological and modern bears and because there is no evidence for dietary change in the time span represented, this result suggests, but certainly does not confirm, that modern Missouri black bears may have the same diet as native Missouri black bears that lived on the landscape 200 to 600 years ago. This conclusion, like the previous one, must be tempered by the fact that stable isotopes provide coarse resolution of diet.

Similarity between Lawson Cave and Modern Ursids Stable carbon and nitrogen isotope values for 10 Lawson Cave U. americanus were compared to those of 2 modern Missouri U. americanus.



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Small sample sizes warrant conservative interpretation, but nonparametric Mann-­Whitney U-­tests failed to identify differences between stable carbon isotope values (U = 8, n1 = 2, n2 = 10, p = 0.334), and between stable nitrogen isotope values (U = 4, n1 = 2, n2 = 10, p = 0.099) for modern and Lawson Cave U. americanus. Thus, the answer to my third question is no; there is no stable isotope evidence to suggest that late Holocene U. americanus from Lawson Cave consumed a different diet than modern U. americanus. As with the previous two conclusions, the stable isotopes do not reveal subtle differences in diet. Such may in fact exist between modern ursids and the Lawson Cave ursids. For now, however, stable isotopes have provided previously unknown insights to local ursid diet.

Discussion Late Holocene Ursus americanus from Lawson Cave have stable carbon isotope values similar to those of late Holocene D. virginiana, modern U. americanus, and modern L. rufus; thus, each of these species consumed primarily C3 materials (i.e., plants and/or consumers). Studies of modern black bear diet indicate substantial contributions from vegetation, but insects, carrion, and sometimes moose (Alces alces) and white-­tailed deer (Odocoileus virginianus) comprise a portion of the diet (Austin et al. 1994; Boileau et al. 1994; Irwin and Hammond 1985; Landers et al. 1979; Mosnier et al. 2008; Schullery 1986). Stable nitrogen isotopes reveal whether the stable carbon isotope signature measured in Lawson Cave U. americanus was the result of C3 plant consumption only, or a combination of C3 plant consumption and C3 consumer consumption. Lawson Cave U. americanus have stable nitrogen isotope values similar to those of Lawson Cave S. scrufa and modern U. americanus; an average stable nitrogen isotope value of 4.74‰ indicates meat consumption at the secondary consumer trophic level (i.e., omnivore). Lawson Cave U. americanus likely consumed meat on a regular basis, but their meat consumption was far less than that of Lawson Cave Canis spp. (6.97‰), which would be classified as a tertiary consumer (true carnivore). Spearman’s rank-­order correlation indicates no significant changes in ursid stable carbon and nitrogen isotope values, but rather stasis through time. Additionally, a Mann-­Whitney U-­test comparing Lawson Cave and modern U. americanus diet indicated no significant differences in stable carbon or stable nitrogen isotope values. These results do not suggest any dietary changes for U. americanus in Missouri over the past several hundred years.

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Management Implications Ursus americanus was extirpated from Missouri by the early twentieth century. Therefore, modern attempts to manage this species in the region have no modern local reference data for native black bears (Wolverton 2008). Until now, biologists have had to assume that extralocal (non-­ Missouri) black bears provide accurate ecological and behavioral benchmarks for native Missouri black bears. The Lawson Cave ursids provide answers to questions about U. americanus diet in Missouri obtainable in no other way and provide a test of the above assumption. Results from this study indicate that the isotopic composition of late Holocene U. americanus diet is not significantly different from that of modern U. americanus diet, nor is the trophic level occupied by the two populations different. Therefore, ecologists may confidently use general dietary data on modern extralocal black bears as benchmark reference conditions for native Missouri black bears. Ethnographic research and enactment of wildlife ordinances concerning ursids indicate that many Missourians recognize that local black bear numbers have drastically changed over time, and many people are interested in protecting and rehabilitating the species (Cowan 1970; Etling 2000; Rosania 2010). Fifty years ago, the Arkansas Game and Fish Commission developed the same concern over dwindling local black bear populations (Smith and Clark 1994). Therefore, in 1958 wildlife officials began transplanting U. americanus from Minnesota and Manitoba to Arkansas (Etling 2000; Smith and Clark 1994). These relocations were successful, and the population of U. americanus grew and expanded into southeastern Missouri. Consequently, U. americanus numbers in Missouri have increased over the last several decades. But while wildlife conservationists and wildlife enthusiasts have enjoyed this unintended success, it has not been without consequence. Curious bears (no doubt in search of food and homes) now come into contact with people in residential areas. Some residents, fearing these animals, shoot them. For these reasons, wildlife management programs must find suitable regions for relocation of black bears based on the needs of the animals and local human residents. Wildlife conservationists should look for regions interspersed with trees and reasonably separated from areas with high human population densities. Suitable regions likely exist in the Ozark Highlands (figure 7.1), which offer some of the most biologically diverse habitats in Missouri. Regions can be evaluated for suitability based on habitat and food availability, but successful wildlife management programs must also include public outreach and education programs designed specifically for the species being relocated. Teaching the public to coexist with “new”



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species is an important aspect of conservation today. The combination of past, present, and future knowledge of people, the landscape, and the organisms living on the landscape will encourage successful wildlife conservation programs. Those programs, which are designed to maintain viable populations of organisms, can benefit from long-­term studies of ecosystem trophic structure and of organism habitat use that applied paleozoology can provide. Interestingly, case studies similar to the Lawson Cave research described here can be developed for many species in any state. Paleozoologically derived data, which can expand our knowledge to create a more comprehensive understanding of ecological systems, is an essential component of future wildlife conservation and management decisions. Acknowledgments Funding for this study was provided by the W. Raymond Wood Opportunities for Excellence in Archaeology Fund, Department of Anthropology, University of Missouri–Columbia. Fluorine analysis was performed at the University of Missouri Research Reactor. Additionally, stable isotope analysis was performed at the Department of Geological Sciences Stable Isotope Laboratory at the University of Missouri–Columbia. I thank David Ferring, University of North Texas, Geography for drafting the map presented in this chapter. I am grateful to Steve Wolverton, Charles Randklev, R. Lee Lyman, and Andrew Barker for providing constructive comments on earlier versions of this chapter. Note 1. Direct AMS radiocarbon dating of each femur undertaken after this analysis (Lyman et al. 2012) does not alter results presented here.

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CHAPTER EIGHT

Rockfish in the Long View Applied Zooarchaeology and Conservation of Pacific Red Snapper (Genus Sebastes) in Southern California Todd J. Braje, Torben C. Rick, and Jon M. Erlandson

Rockfishes (genus Sebastes) are a diverse group comprising nearly 100 species of finfish that occupy almost every coastal habitat along the northeast Pacific (Love et al. 2002). They have been exploited by coastal peoples for thousands of years, and are sold and consumed today under the moniker Pacific red snapper. Commercial rockfish fisheries began along the central California coast in the mid-­nineteenth century, rapidly accelerated in the 1960s, and grew to be one of the most economically successful fisheries in western North America (Love et al. 2002:72–77). Overfishing in the last several decades has resulted in an alarming decline of stocks (but see Moser et al. 2000 for alternative explanations), especially in Southern California, where widespread closures of the sport and recreational fishery have been instituted to help reduce the decline of rockfish biomass. In many cases these closures have been retracted due to pressure from commercial and sport fisheries (Lenarz 1986). Over the last decade, it has become obvious to many biologists, ecolo-

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gists, and resource managers that management of numerous modern industrial fisheries has been ineffective at establishing sustainable harvests of the world’s most important commercial fisheries. Scores of formerly productive and seemingly inexhaustible fisheries around the world, from Atlantic cod (Gadus morhua) to Peruvian anchovy (Engraulis spp.) and Pacific abalone (Haliotis spp.) have collapsed, and despite legislation to replenish stocks, many species have shown little sign of recovery (Jackson et al. 2001; Pauly et al. 1998; Pew Oceans Commission 2003). Rising human populations, accelerating harvest of ocean resources, increasingly efficient fishing technologies, pollution, climate change, and a “tragedy of the commons” effect (Hardin 1968; see also Feeny et al. 1990) have all contributed to this collapse, but these do not explain why careful management and monitoring have not been more successful in reversing the trend. In 1995, biologist Daniel Pauly argued that fisheries management policies have typically been based on shallow historic records, rarely extending back more than 50 years, prior to periods of commercial fishing that had already depleted or impacted the resource base, a problem he called the “shifting baselines syndrome” (but see a different definition of this concept by Papworth et al. 2009). Pauly (1995) and others (e.g., Jackson et al. 2001; Pauly et al. 1998; Pinnegar and Engelhard 2007) have argued that effective fisheries management must be built on data sets that extend into the deeper past, prior to commercial overexploitation and, ideally, on data that take into account the potential impacts of ancient human hunters and fishers. This requirement uniquely positions archaeologists, historical ecologists, and zooarchaeologists to apply archaeological and other temporally deep historical data to make significant contributions to research on the long-­term health and sustainability of fisheries and marine ecosystems (e.g., Reitz 2004). In this chapter, we explore the deep (pre)history of rockfish fishing on California’s Channel Islands, documenting that the genus was harvested with increasing intensity over the last 10,000 years (figure 8.1). We employ zooarchaeological and historical data to investigate spatial and temporal variability in the productivity of rockfish fisheries. Despite significant paleoenvironmental changes and human predation that intensified through time, the prehistoric stock persisted for millennia. Our findings extend the ecological and exploitation history of rockfish deep into the past, which should inform the construction of more effective conservation protocols and mediate the shifting baselines syndrome. Our discussion demonstrates that zooarchaeological, historic, and modern records of rockfish stocks are essential for guiding restoration of the fishery to a healthy level that can, once again, become a viable and sustainable commercial and recreational fishery.

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Figure 8.1. Location of California’s Northern Channel Islands and the Santa Barbara Channel.

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Rockfish Ecology Rockfish are a diverse group of finfish, consisting of about 102 species worldwide, 96 of which are found in North Pacific and Gulf of California coastal waters (Alesandrini and Bernardi 1999; Love et al. 2002:5; Johns and Avise 1998). Rockfish can be recognized by the sharp spines on their dorsal fins. They generally occupy midtrophic levels. Members of many species cluster in shallow waters when they are young, feeding mostly on plankton, then move into deeper waters as they age, where they prey on herring, small rockfish, and crustaceans (Love et al. 2002; Murie 1995). The greatest density and diversity of rockfish are found along the Southern California Bight and waters surrounding the Northern Channel Islands, where 56 species live today (Love et al. 2002:19). This diversity is likely a result of the region’s complex mix of oceanographic currents, countercurrents, and eddies, the interaction of which results in different water temperature gradients that form the southern and northern limits for a number of species. Channel Island rockfishes occupy a variety of ecological niches and can be found in water depths ranging from the intertidal to offshore habitats over 200 m deep. The greatest diversity of species along the Southern California Bight is found along shallow (30– 100 m) and deep shelf (100–200 m) habitats (Love et al. 2002:25–26). Unlike most other bony fishes that fertilize external eggs, rockfishes internally fertilize eggs and bear live young (Boehlert and Yoklavich 1984). Similar to most other fish, rockfish are capable of reproducing multiple times during their lifetimes, and some of the longest-­lived rockfish may reproduce 100 or more times (Love et al. 2002:31). There is no evidence that reproductive capacity declines through the rockfish lifetime. Instead, older, larger rockfish produce healthier larvae and far more offspring than younger rockfish. This reproductive strategy seems to have evolved as an adaptive response to variable oceanographic conditions of the northeast Pacific, and the reproductive capacity of older fish may be essential for maintaining healthy populations (Love et al. 2002:31; Boehlert and Yoklavich 1984). The timing of reproductive maturity in rockfish varies by species, with some short-­lived species reaching maturity in one to two years and others requiring more than two decades of maturation during life spans that can exceed 100 years.

Historical Records Historic rockfish exploitation probably began when the first Europeans arrived along the New World Pacific Coast. Large-­scale commercial harvests began in the mid-­nineteenth century, however, when several large



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rocky outcrop species were sold in San Francisco marketplaces (Love et al. 2002:71). The market for rockfish, caught mostly by hook and line, slowly grew throughout the nineteenth century to attain moderate commercial importance. At the close of the century, 1.3 million pounds of rockfish were harvested and sold annually from Washington to California, the vast majority as fresh fish and the remainder salted (Love et al. 2002:72). The character of the rockfish fishery remained virtually unchanged until the early 1940s. With low catch prices paid to fishers and the bulk of the West Coast industry centered on halibut (Paralichthys californicus), salmon (Oncorhynchus spp.), and other more desirable species, rockfish remained a relatively unattractive, impractical, and modestly targeted group of species. With the invention in 1943 of the balloon trawl, a net enabling more effective harvest of rocky reef habitats, rockfish became a more central part of the Pacific Coast fishing industry, but low market prices and limited numbers of fishing vessels continued to keep fishing pressure in check (Lenarz 1986:37; Love et al. 2002:72). A boom in the fishery began in the early 1960s when Russian, and later Japanese, fishers targeted massive populations of rockfish over rocky outcrops using large factory trawlers, tow nets, and roller gear (Love et al. 2002:73–74). These technologies allowed greater and more efficient access to rockfish populations over large rock piles that once provided a deterrent to intensive net fishing. American fisheries were able to compete with foreign counterparts through US government loans and subsidies, which allowed them to retool their engines, add vessels to their fleets, and purchase global positioning systems (GPS). In Southern California, Asian immigrants used gill nets to intensively harvest near-­shore, rocky reef species to supply local communities. Overall, as more traditionally popular sport and commercial species such as giant sea bass (Stereolepis gigas) and yellowtail (Seriola lalandi) were overharvested, the public changed their perception of rockfish, and fishers began to outfit their vessels to exploit the untapped commercial potential of rockfish. This led to dramatic increases in exploitation along the Pacific Coast in the mid-­1970s, with the Washington, Oregon, and California industries peaking in the early 1980s and remaining relatively strong into the late 1990s. In California alone, the annual value of the rockfish fishery may have exceeded one billion dollars at that time (Lenarz 1986:39). In the 1970s, a small commercial fishery began for live rockfish captured in shallow waters (