Yellowstone's Destabilized Ecosystem: Elk Effects, Science, and Policy Conflict

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Yellowstone’s Destabilized Ecosystem

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F R E D E R I C H . WA G N E R With contributions by Wayne L. Hamilton and Richard B. Keigley

YELLOWSTONE’S DEST ABILIZED ECOSYSTEM Elk Effects, Science, and Policy Conflict

1 2006

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Oxford University Press, Inc., publishes works that further Oxford University’s objective of excellence in research, scholarship, and education. Oxford New York Auckland Cape Town Dar es Salaam Hong Kong Karachi Kuala Lumpur Madrid Melbourne Mexico City Nairobi New Delhi Shanghai Taipei Toronto With offices in Argentina Austria Brazil Chile Czech Republic France Greece Guatemala Hungary Italy Japan Poland Portugal Singapore South Korea Switzerland Thailand Turkey Ukraine Vietnam

Copyright © 2006 by Oxford University Press, Inc. Published by Oxford University Press, Inc. 198 Madison Avenue, New York, New York 10016 www.oup.com Oxford is a registered trademark of Oxford University Press All rights reserved. No part of this publication may be reproduced, stored in a retrieval system, or transmitted, in any form or by any means, electronic, mechanical, photocopying, recording, or otherwise, without the prior permission of Oxford University Press. Library of Congress Cataloging-in-Publication Data Wagner, Frederic H. Yellowstone’s destablized ecosystem: elk effects, science, and policy conflict / by Frederic H. Wagner. p. cm. Includes bibliographical references and index. ISBN-13 978-0-19-514821-3 ISBN 0-19-514821-5 1. Ecology—Yellowstone National Park—History. 2. Yellowstone National Park—History. I. Title. QH105 . W8W34 2006 577' . 09787'52—dc22 2005017301

9 8 7 6 5 4 3 2 1 Printed in the United States of America on acid-free paper

To the three most important people in my life, in the order in which I met them: Marilyn, Greg, and Jeff.


Preface

Human reason is beautiful and invincible. No bars, no barbed wire, no pulping of books, no sentence of banishment can prevail against it. —Czeslaw Milosz

THE NEED FOR A YELLOWSTONE SYNTHESIS This book has two major messages. The first is the synthesis of a conceptual model simulating the structure and function of an ecosystem along the northern border of Yellowstone National Park, known as the northern range. It is the winter range of the largest of several wintering herds of Rocky Mountain elk (Cervus elaphus) that summer together on the higher elevation summer range in the central and southern portions of the park and migrate to lower elevations to winter. Up until 1989, the northern range was considered to cover approximately 100,000 ha, 83,000 of this inside the park boundary (Houston 1982). But starting with that date, on average a third of the northern herd has moved each winter to terrain north of the park, and the northern range is now measured at 152,663 ha (Lemke et al. 1998). Archaeological research, photographic and historical records, annual reports of park superintendents since establishment in 1872, and a plethora of research on the effects of elk on the northern range begun in the 1920s provide a massive accumulation of information on the northern range. This accumulation provides the raw material for my synthesis, conducted in chapters 2–15, of the northern range model and effects of the northern herd on the ecosystem represented by that model.

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As readers begin going through these chapters, they will soon become aware that I draw quite different inferences from the evidence than those that have been posited by most of the published researchers since 1971 and reiterated by park personnel since that date. This divergence of views raises the entire question of the role of science in public-policy setting. That, in turn, is part of the broader subject of policy-setting procedures in natural-resources agencies and how those procedures create an environment affecting the objectivity of their science programs. In chapters 16 and 17, I address these two questions in relation to Yellowstone and more generally U.S. national parks. Chapter 16 explores why the park-supported paradigm of the northern range has differed so markedly from the one I propose here. Chapter 17 proposes policy-setting mechanisms for U.S. national parks that would be condusive to the valid science they need for effective policy settings. These discussions in chapters 16 and 17 are the second message of the book. My motives in writing this book have been fivefold: (1) the sheer pleasure of doing science; (2) exploration of the science–policy interface; (3) a concern for the welfare of science itself at a time when science is under attack from a number of quarters; (4) a need to resolve the disputed northern range issue on behalf of numerous agency and academic scientists, both U.S. and Canadian, who have looked with near incredulity on the northern range issue but had no research under way with which to pose factual critiques; and (5) a concern for the welfare of the U.S. National Park System, an asset of incalculable value to the American people.

WHAT IS SYNTHESIS? The central purpose of science is to describe the real world by drawing inferences from observational evidence and from the deduced implications of premises based on real-world evidence. Ecology describes ecological systems at an array of organizational or complexity levels ranging from landscapes and ecosystems to individual organisms, depending on the questions of interest (compare O’Neill et al. 1986; Allen and Hoekstra 1992). Science commonly begins description of a phenomenon of interest at a given organizational level by posing a hypothesis based on existing evidence and observations. It then tests the degree to which the hypothesis is congruent with the real-world phenomenon (its validity or “truth”) by deducing implications of the hypothesis and testing these with observation and experimentation. Typically this procedure investigates mechanisms producing or “causing” the phenomenon of interest at organizational levels below that of the phenomenon itself. This is the process of analysis, “the separation of a whole into its parts for study,” according to the common dictionary definition. As Hilborn and Mangel (1997) comment, Platt (1964) likens this to climbing a research tree where each fork is an experimental result that leads to new hypothetical and experimental branches and, for our purposes here, lower organizational levels.

Preface

As Popper (1979) reasons, it is never possible to prove a hypothesis, that is, establish that it accords perfectly with the natural phenomenon it is posed to describe. It is only possible to establish a degree of probability that it is a valid description. That probability grows (1) with the consistency of subsequent observations and experimental results in elucidating the phenomenon’s causal mechanisms, and (2) with the absence of any more probable hypothesis. Synthesis, “the combining of separate elements or substances to form a coherent whole,” is the reverse of analysis. Synthesis is thus an inductive process, whereas analysis is more deductive. Synthesis, too, strives to describe some aspect of nature and does so by bringing together observations, experimental results, and hypotheses posed at lower organizational levels. Holling (1995:13) discusses these two approaches, calling them the two streams of science. In commenting on the second stream, which I am calling synthesis, he states: “It is fundamentally interdisciplinary and combines historical, comparative, and experimental approaches at scales appropriate to the issues. . . a science of the integration of parts. . . . It is also the stream that is most relevant for the needs of policy and politics.” As I will comment later, science judges its inferences on levels of probability. Analytic science lends itself to quantifying those levels. Well-replicated experiments and repeated measurements of quantifiable variables facilitate explicit probability statements. Because of the varied sources of evidence, some not quantified or quantifiable, and the problem of assessing the relative importance of the different components contributing to some synthetic result, synthetic inferences commonly do not lend themselves to explicit probability statements. Rather, probability judgments are subjective and comparative: That is, judgments are made as to which of two or more inferences is most consonant with the available evidence. Thus the validity of a synthetic inference, that is, the probability with which it correctly describes the phenomenon of interest, depends in a philosophical sense on 1. the validity with which the individual inferences describing the separate, constituent parts accord with the reality of those parts; 2. the internal consistency between the inferences describing the constituent parts, including the known biology of the species/system in question; 3. the absence of any more probable hypotheses; and 4. either conformity with existing theory or a probable rationale, well supported with evidence, for lack of conformity. In a procedural sense, validity depends on 1. consideration of all evidence, inferences, and hypotheses; and 2. explanation of why one inference or hypothesis is accorded higher probability than another.

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Such comparative probability judgments are best made from a thorough knowledge of the phenomena in question, the relevant literature and theory, and more generally the ecological literature and theory. These procedures are tantamount to Chamberlain’s (1965) method of multiple working hypotheses. The method avoids his ruling theory problem of considering only one hypothesis and building an illusion of validity by amassing evidence only in its support while failing to consider others that may be equally or more valid. As already stated, many of the inferences drawn in this synthesis are contrary to those drawn by other authors. In all cases I have followed the previously described criteria by considering those authors’ views, weighed their evidence, and explained why I consider my evidence to be more valid or my conclusions to be more probable. This is a procedure that has not always been followed by other authors. I hope this synthetic effort produces a truer understanding of the role of the Yellowstone northern elk herd in the northern range ecosystem.

ACKNOWLEDGMENTS The stimulus for writing this book has grown over time in response to a sequence of events and interactions with key individuals. In the early 1960s, I accompanied busloads of wildlife-management students on weekend trips to the park each winter. We were hosted graciously and informatively by then Park Superintendent Lemuel (Lon) Garrison, Park Management Biologist Robert Howe, and his assistant, William Barmore. The generally accepted paradigm of elk effects on the northern range at the time was the overpopulation view, to be clarified in subsequent chapters. By the later 1960s, other faculty members at my institution were accompanying the trips. Garrison and Howe had been replaced, and a new northern range paradigm—natural regulation—prevailed. Students and faculty were returning with spirited and often incredulous comments over the new policy that was contrary to the park’s previous half-century position and to prevailing ecological and wildlife-management theory. By the 1970s, a number of professional ecologists with whom I had lengthy conversations emerged as critics of natural regulation. Most incisive were the late Leslie Pengelly, professor of wildlife at the University of Montana, Missoula; A. A. Beetle, botany professor at the University of Wyoming; and range ecologist Robert Ross with the U.S. Soil Conservation Service in Bozeman, Montana. William Barmore, now retired, was originally in accord with the overpopulation paradigm but later supported natural regulation. Bill and I have continued meaningful discussions to the present. These critics and I agreed that natural regulation needed searching appraisal. But none of us had any research under way in the park with which to conduct an outside evaluation of the concept. All research at the time was being carried out by park biologists.

Preface

Two individuals emerged in the 1980s who were particularly aggressive in bringing public scrutiny onto the northern range situation. Alston Chase, former chairman of the Philosophy Department at Macalester College, resigned his academic position and sold his Montana dream ranch to support his family while researching and writing his compelling book Playing God in Yellowstone (Chase 1986). Also in 1986, Charles Kay was a freelance environmental consultant, writer, and photographer who had a master’s degree from the University of Montana and had been employed by the Alaska Department of Fish and Game to police the environmental restrictions on Alaska pipeline construction. Skeptical about natural regulation, Kay was conducting research on northern range aspen at his own expense. At my invitation he moved to Utah State University to conduct a massive, five-year doctoral dissertation study evaluating the natural-regulation paradigm. But Kay’s contribution extends far beyond his northern range study. He had exhaustively reviewed and quantified both the archaeological record for Western North America and the early historical accounts on wildlife resources. The result was an ecological scenario, already advocated by many archaeologists, that pre-Columbian ungulate populations in the Intermountain West had been held at low densities by native carnivores and especially by aboriginal hunting. The point is not only pivotal in the debate over the natural regulation paradigm but more broadly relevant to North American ecology and the implications to national park, wilderness, and protected-area policies. Within the past 15 years, my knowledge of the northern range scene has been advanced further by interactions with former Yellowstone researchers Richard Keigley (now with the Biological Resources Division of the U.S. Geological Survey), geologist Wayne Hamilton (retired), and Douglas Scott (no longer with Yellowstone). I have valued interaction with range science professor Carl Wambolt (Montana State University, Bozeman), who has conducted important northern range research; and with Thomas O. Lemke, biologist with the Montana Department of Fish, Wildlife, and Parks who has generously provided me with updated information on the northern range. I appreciate the efforts of three anonymous reviewers from whom Oxford University Press sought comments before contracting to publish this book, and of Lance Gunderson of Emory University for suggestions on how to improve it based on his reading of a late draft. Stephanie White tirelessly and always cheerfully typed drafts and figures. Andrea Bell’s skills in helping develop the figures have been indispensable. Senior editors Kirk Jensen, Peter Prescott, and Kaity Cheng at Oxford University Press have been a pleasure to work with. All research requires financial support. Kay’s 1986–90 studies were supported by the Rob and Bessie Welder Wildlife Foundation of Sinton, Texas. James Teer was director at the time, served as president of The Wildlife Society during the period, and asked me to chair a society study of wildlife policies in the national parks (Wagner et al. 1995a). Writing this book has been

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made possible by a fellowship to the National Center for Ecological Analysis and Synthesis, Santa Barbara, California. Its director, James Reichman, provided important encouragement. With the most left for last, I am deeply indebted to my wife, Marilyn, for her affection, patience, and selfless support of this effort. Her companionship and help throughout have been essential to its completion.

Preface

Contents

Part I. A Yellowstone Synthesis 1. History of the Northern Range Dispute 3 Part II. Elk Population Changes and the Synthesis Design 2. The Census Period: 1923–2003 15 3. Prehistory to the 1880s 29 4. The First “Experiment”: 1878–1923 40 5. The 132-Year Population Trajectory and Associated Synthesis Design 48 Part III. Elk Effects on Ecosystem Structure and Function 6. Influences on Upland System Structure I: Aspen Woodland 59 7. Influences on Upland System Structure II: The Sagebrush-Steppe Subsystem 91 8. Influences on Upland System Structure III: Conifers and Deciduous Shrubs 124 9. Influences on Upland System Structure IV: The Ungulate Guild 141 10. Influences on Riparian System Structure 172

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Contents

11. Influences on Ecosystem Function I: Erosion, F REDERIC H. WAGNER R ICHARD B. KEIGLEY 215

AND

12. Influences on Ecosystem Function II: Historical Perturbations in Small Lake Basins, WAYNE L. HAMILTON 231 13. Influences on Ecosystem Function III: Bioenergetics 259 14. Influences on Ecosystem Function IV: Nitrogen Biogeochemistry 273 15. Synthesis 280 Part IV. The Role of Science in Policy Process 16. Why the Science Missed the Mark 307 17. The Science–Policy Interface 317 References 335 Index 359

A Yellowstone Synthesis

I Conservation is paved with good intentions which prove to be futile, or even dangerous, because they are devoid of critical understanding. —Aldo Leopold The nation behaves well if it treats the natural resources as assets which it must turn over to the next generation increased, and not impaired, in value. —Theodore Roosevelt


History of the Northern Range Dispute

1 Management efforts fail because of the inadequacy of theory. —C. S. Holling

FOUR MANAGEMENT PHASES Elk, the most broad-spectrum feeders among North American ungulates, subsist by grazing herbaceous vegetation and browsing woody species. Hence, their feeding can affect all parts of the vegetation. During summer, they are found over most of the 9,000 km2 of Yellowstone National Park, then subdivide into several discrete herds to winter outside the park. As commented, the largest of these winters in a low-elevation area in and adjacent to the northern part of the park, the northern range (figure 1.1). An extensive body of literature discusses issues involving this herd in the context of Yellowstone history, with excellent accounts presented in Sellars (1997), Pritchard (1999), and Shafer (2000). The northern herd has changed over time in the course of four distinct management phases, if the term management can be construed to mean human action, whether subsistence use or agency-prescribed protocol. The first phase was the period of subsistence use by native inhabitants, spanning centuries or probably millennia prior to the establishment of Yellowstone National Park in 1872. Early park officials considered that elk populations in the region had been held at low densities prior to European contact by carnivorous predators and aboriginal hunting. They opined that animals that we now consider the northern herd summered in the higher elevations of what became the park; and migrated considerable distances northward down the Yellowstone River valley, out of what later became the park portion of the 3

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A YELLOWSTONE SYNTHESIS

Figure 1.1 Map of the northern range of Yellowstone National Park, redrawn from Cole (1969). A divides the upper and lower ranges. BLA is variously the boundary line, Gardiner, or low-elevation ranges.

northern range, to winter in lower elevations. Elk were not believed to have wintered in large numbers on today’s northern range. These views prevailed among park personnel for nearly a century until challenged in the 1960s and 1970s. The second management phase began with park establishment and continued for approximately a half century. Its protocol was protection, with Native Americans evicted, hunting banned, predators controlled, and ungulates fed artificially in winter. Hunting continued for a period following creation of the park until 1877, when then Superintendent Philetus Norris appealed to market hunters to stop the killing. Hunting abated by the end of the decade, and assignment of a military detachment in 1886 further curtailed killing by humans. The size of the herd began increasing once hunting was brought under control. Park officials began censuses and concluded that the northern herd had


increased to 20,000–35,000 animals by the early 1900s. At the same time, EuroAmerican settlement and hunting outside the park were forcing the herd to winter largely inside its boundaries on the northern range. By 1914, park observers were reporting marked changes in vegetation in the northern range (Smith et al. 1915). These were attributed to heavy grazing and browsing by the enlarged northern herd. Park officials decided that the herd had to be controlled to avoid continued and excessive damage to vegetation and entered into management phase 3, herd control. In the 1920s elk were trapped and removed to reduce the herd. “Direct reduction” (shooting by park rangers) began in the 1930s and continued annually until the herd had been reduced to 3,172 animals counted in the 1968 winter census (Houston 1982). The fourth management phase was announced in 1967 with issuance of three park policy documents: “Management Objectives for Northern Yellowstone Elk” (Anonymous 1967a), “Natural Control of Elk” (Anonymous 1967b), and “Administrative Policy for the Management of Ungulates” (Anonymous 1967c). The policy set forth in these documents is generally referred to as the “naturalregulation policy” (emphasis added). The policy asserted that “natural” processes (e.g., winter weather, intraspecific competition for forage) would limit the elk population without direct human intervention. Predation was included in the 1967 documents as an effective natural control, but was dismissed 4 years later when predation was hypothesized to be a “nonessential adjunct” (Houston 1971). The policy was activated in 1969, the first year without direct reduction.

SCIENCE, POLICY, AND MANAGEMENT Management of natural resources is based on concepts and understanding of nature and ecology. Native Americans likely had sophisticated understanding of elk ecology, which they used to direct their harvesting decisions during what I categorize as phase 1. Ecology, wildlife, and range management had not yet emerged as disciplines in 1872, nor had the development of science-based approaches to management. Hence management phase 2—protection from hunting, predator control, artificial feeding in winter—was based more on conventional wisdom and perhaps analogy with animal husbandry. Phase 3 had an a priori and continuing basis in professional observation and science. As early as 1914, a survey of elk and habitat conducted by the Chief Field Naturalist of the U.S. Bureau of Biological Survey and two forest supervisors of the USDA Forest Service (Smith et al. 1915) reported heavy impacts on willow (Salix spp.) and aspen (Populus tremuloides). The first biologist, M. P. Skinner, was assigned to the park at about the same time and was publishing in the scientific journals by 1921. Science-based management strengthened over the following four decades with the growth of an increasingly intensive and sophisticated research program and a stream of technical publications. Research by a sequence of agency biologists during this phase measured vegetation trends and built exclosures to determine

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vegetation dynamics free from ungulate grazing. Systematic photographs documented conditions and trends in the vegetation, especially the woody component, and complemented photographs taken more randomly in the early years. Phase 4 was adopted without any new scientific evidence, was contrary to a half century of prior research, and contrary to prevailing ecological theory (Sellars 1997; Pritchard 1999). This was pointed out in 1968 by park biologist William Barmore in a memorandum to his superior (1968): we are abruptly “scrapping” current objectives with no stated justification or supporting information. Past objectives were based on supporting information that indicated relationships between elk, their habitat, and associated wildlife were different from what existed in primeval times; that ecological changes since the early 1900s were “unnatural.” . . . I think we are being less than completely “scientific” or completely objective and honest by proposing a drastic program switch without saying why. The natural-regulation policy was adopted for political reasons (Chase 1986), as documented in congressional testimony (U.S. Senate 1967). The park was being besieged by protests from hunters in the surrounding states who objected to the rangers’ shooting and who demanded that the park be opened to public hunting. Then Park Service Director G. B. Hartzog Jr. (1988:252–53) stated: I did agree with [Wyoming] Senator McGee to stop the shooting in 1967; I did not agree that the National Park Service would never shoot again. . . at hearings in Casper and in Washington we carefully laid out the sequence of our wildlife management policy . . . : (a) Natural predation. (b) Trapping and transplanting. (c) Shooting by sport hunters of excess animals outside the parks. (d) Control by shooting within the parks. Senator McGee . . . acknowledged that direct reduction remained a viable management option. This policy was based on recommendations of an outside National Parks Advisory Board chaired by University of California, Berkeley, wildlife professor A. Starker Leopold (Leopold et al. 1963). Four years after the announcement of the natural-regulation policy, a scientific rationale, the natural-regulation hypothesis, was developed to support the policy. The 1967 policy proposed that “winter weather, native predators and the elk population itself interact to naturally control elk numbers within limits set by winter food” (Anonymous 1967b). In 1971 Yellowstone Chief Biologist Glen Cole presented a brief conference paper (Cole 1971) that unequivocally generalized the reality of ungulate natural regulation and proposed on evolutionary grounds that it would occur without significant impact on the vegetation. He cited 11 sources, including 6 intra-agency reports, 3 conference presenta-


tions, 1 Ph.D. dissertation, and 1 peer-reviewed publication, but he neither provided evidence nor referred specifically to the northern herd. In the same year, Houston (1971) set forth the natural-regulation ecological hypothesis for the northern herd. The hypothesis had two key tenets: (1) The elk herd would be limited by intraspecific competition for food and associated winter mortality without influence from predation, and (2) would do so at levels that would not have undue effects on the ecosystem. This hypothesis and predictions were linked to the size of the herd at that point (1971) in history. The effects of nearly half a century of herd control were apparent; 3,172 animals were censused in 1968. Thus Houston’s hypothesis predicted: 1. The northern herd would equilibrate at approximately 6,000–9,000 animals. (However, three years later, following a census of 10,529 animals, the hypothesized equilibrium level was revised to 12,000– 15,000; Houston 1974.) 2. There would be no competitive exclusion of sympatric herbivores. 3. The vegetation would not undergo “retrogressive succession.” Historical photographs would not show the contemporary vegetation to “depart from pristine conditions because of grazing by the elk” (Houston 1982:2). Yet if there were changes, they would not result from grazing by native ungulates. Reduction in fire frequency and climatic changes, if there were any, would be implicated. Moreover, “replacement” (= disappearance) of willow and aspen would not represent retrogressive succession. The new hypothesis was followed with a new research emphasis by park personnel. It produced a number of shorter publications and mimeographed releases, but the seminal documents were a large, two-volume unpublished report on the northern herd (Houston 1974, 1976); a book adapted from the report (Houston 1982); a massive 1980 report by Barmore (1980) eventually published 23 years later (Barmore 2003); and a short, popularized book for general readership (Despain et al. 1986). (I cite Barmore’s work frequently in this book, but I have chosen to use the 1980 report rather than his book. Barmore assures me that no major changes were made for his book, but the report was written at the time of the research, and most of the manuscript for this book was written by the time his book was published.) The research updated some of the vegetation measurements instituted during phase 3, reexamined the historical and photographic record, and studied elk demography and feeding ecology. In general it reversed most of the inferences about the northern range held during phases 2 and 3 and supported the tenets of Houston’s 1971 natural-regulation hypothesis. Soon after its enunciation, considerable skepticism arose in the scientific community about the ecological probability of the natural-regulation hypothesis and about interpretations of the evidence supporting the hypothesis by agency biologists (see Pengelly 1963; Patten 1968; Beetle 1974a, b; Anonymous

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1975; Erickson 1981). But it was Chase’s (1986) strident critique that raised attention levels on the issue. In 1986, Congress appropriated funds for the park to research the question “Is the northern range overgrazed?” The funds were used to support “40 separate research projects” (Singer 1989) both by National Park Service (NPS) and extramural scientists. Shortly after the overgrazing appropriation, additional major funding was made available to study the effects of the 1988 fires and evaluate wolf reintroduction, and some of the research was relevant to the naturalregulation issue. Consequently there was a great deal of park-supported research on the northern range in the latter 1980s and 1990s, and most drew conclusions supporting the park’s hypothesis. During this time, other funding sources supported studies that were critical of the natural-regulation paradigm. The most comprehensive was the work of Kay (1990, 1994a,b,c; Chadde and Kay 1991; Kay and Wagner 1994, 1996), which was carried out concurrently with the overgrazing studies. Wagner et al. (1995a, b), Keigley and Wagner (1998), and Wambolt and Sherwood (1999) also published critical reviews of the paradigm.

TESTING THE NATURAL-REGULATION HYPOTHESIS Several problems complicate a scientifically explicit and objective test of the natural-regulation hypothesis. Undefined terminology and falsification criteria have pervaded the issue since Houston (1971) formulated his concept. The hypothesis was to be considered falsified if the elk herd “erupted.” But the term eruption was never defined. The size of the herd increased by a factor of six from its low point in 1968 (Houston 1982) to the levels of the early 1990s (Coughenour and Singer 1996). But whether that constituted an eruption and falsified the hypothesis has been impossible to say without definition of the term. The hypothesis could also be falsified if the elk induced retrogressive succession. This phase was never defined in terms of the northern range ecosystem. In hindsight, the Clementsian connotation of the term, posed before the newer multiple equilibrium concepts of succession (Westoby 1979/80; Ellis and Swift 1988; Friedel 1991; Laycock 1991) described a more complex phenomenon, has further reduced the term’s utility as a falsification criterion. Even more uncertainty surrounds the term overgrazing. Not only was it not defined ecologically in terms of the northern range at the beginning of the congressionally funded overgrazing research, but the prefix carries a clear value connotation. Aside from the ecological ambiguity, it is not clear when or how grazing becomes overgrazing and by whose values. Coughenour and Singer (1991) explored some of these questions, but their deliberations do not appear to have been specifically applied in a formal review of the northern range situation. Notwithstanding these uncertainties, a large sum of money and 40 separate research projects were applied to the overgrazing question, and the results


are interpreted by park officials and reported to Congress as indicating that the northern range has not been overgrazed (Finley 1997). Yet alternative conclusions have been published in the professional literature. These differential interpretations of factors affecting herd and ecological dynamics are but one facet of a larger and deeper problem that occurs when science and policy interact. As discussed, there are two separate but commingling aspects to the northern range issue. One is the scientific truth of the situation, achieved independently of and without influence from the policy implications of that truth. Establishing the truth of what effects the different policies have had on the size of the northern herd, and consequently the herd’s ecological effects on the northern range, are scientific questions. The second aspect is deciding on the appropriate policy for managing the northern range. Whether a given herd size and its ecological effects are desirable or appropriate is a value and policy question, not a scientific one. Policy setting is a sociopolitical process, not a scientific one (Wagner 1996a,b). Science should illuminate the policy process, but its inferences should not be influenced by it. There has been a pervasive failure to separate these two aspects. The effects of the commingling are twofold. One is that it complicates clear and dispassionate thought, communication, and discourse. Numerous publications combine the two treating them as a single issue. For individuals who favor the natural-regulation policy on the basis of their own personal values, it is difficult to concede the full ecological effects of elk on the northern range ecosystem. To some, stating those effects fully and objectively is tantamount to criticizing the policy. It is difficult to discuss the northern range situation with some people close to the issue without the two elements arising and clouding unambiguous exchange. A more pernicious effect is that policy implications color the objectivity of scientific inference. Park scientists have been in an especially difficult position. The policy was articulated in 1967, the hypothesis in 1971. To question the hypothesis, and by implication the policy, both of which have been the park’s position for more than 30 years, is problematic. Although others shared the view, it was Chase (1986) who most forcefully alleged that the park’s reinterpretation of previous evidence and negation of prior influences were cast for conformity with the new, politically driven natural-regulation policy. Sensitivity to the charge undoubtedly prompted park investigators to the defensive comment in their popularized book (Despain et al. 1986:10) that “Park Service research projects are not aimed at simply supporting some established policy.” But agency research largely continued to support the hypothesis as has much of the park-supported extramural research. A third problem challenging fair yet critical synthesis is the size and diversity of the documentary record. There is an immense body of research information that ranges from unpublished original data, internal reports, theses and dissertations, park reports, and peer-reviewed publications in the scientific literature. The park has an active public information program that issues releases for the written media; publishes popularized books and bulletins (some of a semiscientific nature by virtue of their abundant citation of the scientific literature);

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and distributes tabloids to tourists driving through the park entrances. All assure the reader that its policies are appropriate and achieving sound management. These sources at times refer to unpublished findings without presenting them, and their inferences may be cited subsequently as established fact. Hypotheses—sometimes highly speculative ones without supporting evidence— are proposed and subsequently cited as being well established, in some cases used as the basis for policy support or decisions. This literature is contradictory; some is written by nonscientists with resulting lack of ecological insight. Fair and objective synthesis demands that all of this be consulted and considered. The primary purpose of this synthesis is to review, analyze, and evaluate the existing evidence to address the scientific question of what effect the several management phases have had on the size of the northern herd and in turn its effects on the northern range. The park-supported science is being used to sustain the natural-regulation policy and is being popularized to the public to gain its support. This book examines the science critically to assess whether the claimed ecological effects of the natural-regulation policy are being validly portrayed, to determine whether that policy is being truthfully supported and the public is being accurately informed. As commented in the preface, there is no intent here to advocate a policy for the northern range. But chapter 17 discusses the need for a policy and a number of considerations involved in arriving at such a policy.

IS THE NORTHERN RANGE ENTERING A FIFTH PHASE? The management of the northern range has clearly passed through four phases, and may be entering a fifth. The latest may be moving toward conditions similar to those that prevailed before park establishment, characterized by movement of wintering elk out of the park to foraging areas north of the park and reduction of elk numbers by predation. As discussed previously, the size of the northern range has increased since the late 1980s. This apparent range extension is probably attributable to ecological conditions in 1988. That year was exceptionally dry with limited forage production. The 1988–89 winter was extremely severe. More than 7,000 northern herd elk (Lemke et al. 1998) moved out of the park in search of winter food. This total was approximately 40% of one herd size estimate (Taper and Gogan 2002) for that year. The 1988–89 exodus evidently established a tradition. Between 1988–89 and 1996–97, the number of elk wintering north of the park each year averaged 5,460 (Lemke et al. 1998), on the order of a third of the censused animals in these years (T. O. Lemke, personal communication, June 5, 2002). The number has varied between years, with the fluctuations due to a number of factors, including the severity of winter, location of forage, and perhaps predation pressure. In the winters of 2000–2001 and 2001–2002 the numbers of animals


wintering north of the park were 3,800 and 5,100, respectively (T. O. Lemke, personal communication, June 5, 2002). One stimulus for the new northward movement of wintering animals could be depletion of forage in the park portion of the northern range after decades of heavy use by a large herd and need for the animals to forage more widely in search of food. But one might also speculate that establishment of the wolf population in the park could be a further stimulus. Wolves were introduced into the northern range in 1995. The population probably required 3–4 years to increase sufficiently to affect the herd size. By the winters of 2000–2001 and 2001–2002, elk calves/100 cow ratios had fallen from ~30–40/100 in prior years to 26–28/100 in 2001 and 14/100 in 2002 (T. O. Lemke, personal communication, June 5, 2002). Approximately 45% of the elk killed by wolves are calves (Lemke 2000). It is too early to know whether wolf predation will significantly reduce the size of the northern elk herd. But the wolves have established territories precisely in the Park portions of the northern range. The northern herd winter censuses have declined since the mid-1990s (see chapter 2), and the media have promptly suggested this may be due to wolf predation (Dixon 2002; Stark 2004). But the Western United States has been in the throes of a drought. Merrill and Boyce (1991) and Coughenour and Singer (1996) found strong correlations between variations in summer phytomass and water-year precipitation, respectively, and variations in annual rates of population change in the northern herd. Lemke (personal communication, June 5, 2002) reported that elk herds have declined in other areas of western Montana where there are no wolves. Thus, it is not clear that the northern herd decline has been largely or even partially the result of wolf predation. Both the exodus of a significant fraction of the northern herd out of the park each winter, and whatever the extent to which the wolves may reduce the number of animals wintering in the park, the effect should be to ease foraging pressure on the park winter range. There already is some evidence that this may be occurring. Singer et al. (2001a) observed significant increase in the heights of willows in 2000 above those measured in 1986–92. Richard Keigley (personal communication, July 18, 2002) observed release of cottonwood saplings (Populus angustifolia) in July 2002 in the Soda Butte Creek valley unlike any he had seen during his research in 1992 (Keigley 1998). There is also evidence that the wolves are changing the distribution of elk on the winter range and eliciting vegetation responses in the areas they avoid (Beschta 2003; Ripple and Beschta 2004). This chapter has provided a brief overview of the history of interaction between humans and elk populations in the northern range. At least four (and a possible fifth) management phases were described. Each new phase was characterized by an alternative mental model or hypothesis of the status and trends of the northern herd and its relationship to the northern range. Subsequent chapters examine the complexities of past and current phases.

11


Elk Population Changes and the Synthesis Design

II Right conclusions are more likely to be gathered out of a multitude of tongues than through any kind of authoritative selection. —Judge Learned Hand


The Census Period: 1923–2003

2 The numbers of North American deer are limited by food shortage, which causes a rise in the death rate, especially among the young and senile, and a fall in fecundity. The reduction is density-dependent, though precipitated by heavy snowfalls. —David Lack

ELK POPULATION MEASUREMENTS Inferences from Census Data The first superintendents assigned to the park after its 1872 establishment were civilian; from 1886 to 1916, park administration was assumed by the military (Houston 1982:3). Under both administrative regimes, park officials attempted annual estimates of elk numbers. Houston (1974, 1982) investigated these reports at length and was skeptical of most. In 1916, following passage of the National Parks Organic Act and formation of the National Park Service, Yellowstone National Park was returned to civilian administration. The agency began its own winter censuses in 1923, but Houston (1982:15) concluded that the early ones left something to be desired. Concerted efforts at annual total counts, conducted on the ground, were begun in 1930, conducted annually through the 1930s with an accompanying aerial count in 1935, and carried out four years in the 1940s (Houston 1982:16). The park changed to aerial counts in 1956, varying between a single helicopter or single fixed-wing craft through the 1960s, 1970s, and the single count in the early 1980s (Houston 1982:17; Singer and Garton 1992; Lemke et al. 1996). 15

16

ELK POPULATION CHANGES AND THE SYNTHESIS DESIGN

From 1985–86 up to the present, the counts have been made with four fixedwing planes, with partial counts made by a single helicopter from 1985–91 to assist in evaluating census accuracy. Only 4 counts were made in the 1940s, 2 in the 1950s, and 6 in the 1960s. No counts were made in 1979–80, 1980–81, or the 3 years 1982–85. Moreover, the counts in 1960–61, 1976–77, 1988–89, 1990–91, and 2002–2003 were considered poor counts (Houston 1982:17; Lemke et al. 1998; Lemke, personal communication, February 12, 2003). In this chapter I analyze the census and kill data for insights into the population’s behavior, both to provide understanding of its dynamics and to test the population portion of the natural-regulation hypothesis. There have been several excellent population analyses of the northern herd in recent years, and my procedures are somewhat similar to those of previous authors, although not as detailed as some. I use slightly different numbers for reasons that I describe shortly and update the time series from those used in the excellent Lemke et al. (1998) and Taper and Gogan (2002) studies, which extended through the 1995 census. I also point out several implications of the results that were not pointed out by previous investigators.

Census Accuracy: Variables and Corrections A series of excellent studies has addressed the question of how accurately the different census methods reflect the actual population sizes, explored the variables affecting the censuses, developed correction factors, or estimated population size by other means. As a result, the numbers on this population have probably been bracketed as narrowly as any ungulate population of this size and over an area of this extent. However, the effort has produced at least 10 to 12 different time series of the northern herd, each based on a different analysis of the censuses, or corrections thereof, or for fall or winter, or series calculated from demographic parameters. Hence, some consideration of these is needed to provide background for my own projections. Three sources of variation or bias affect the number of animals counted during the winter censuses. One is simply the fraction of the population actually seen and recorded. This fraction is affected by the census dates, weather conditions, proportions of animals in open terrain, and whether the census is ground or aerial. The second variable is the number of animals removed from the population by park control efforts from the 1920s through 1968 exerted before the winter censuses (Houston 1974:18), and by sport hunting throughout the census period for animals that move outside the park. An October hunting season, varying in number of animals taken, has been in place throughout the years of census and has reduced the population before the winter counts. An annual winter hunt begun in the 1970s has taken varying numbers of animals, in many cases some before the censuses.


The third source of variation is the weather-induced fluctuations in the population itself. In an excellent analysis, Coughenour and Singer (1996) partitioned the weather-related mortality among the seasons and sex-age classes and simulated population trends on the basis of the relationships they perceived. To assess the magnitude of the sightability bias in the ground counts of the 1930s and 1940s, Houston (1982:19) deducted the calves from estimated autumn population censuses and compared the remainders with the late winter censuses of the previous years. The results suggested that the 1934, 1935, and 1937 counts were low by 12%, 9%, and 41%, respectively. The mean of the three is 22%, and he adjusted all other ground counts upward by this amount (his figure 3.2). Two studies have given a firm indication of the aerial counts’ accuracy from 1971–72 through 1990–91. Singer and Garton (1992) summarized studies on visibility bias, which showed fixed-wing counts observing 75% of marked elk in early winter and 50% in the late winter. Comparable figures for helicopters were 83 and 70. The authors constructed a multivariate analysis–based sightability model. In separate studies, Mack and Singer (1992) calculated annual population sizes with models using demographic parameters of the herd. During 10 years of comparison, the censuses averaged 76% of these estimates corrected for lateseason hunting removals. Coughenour and Singer (1996) generalized that the censuses count approximately 75% of the northern range animals. Lemke et al. (1998) further refined the censuses by correcting slight recording and computational errors. They produced “minimum fall” and “minimum winter” population estimates for 1975–76 through 1994–95 by correcting the censuses for hunting removals outside the park. Lemke (personal communication, February 12, 2003) has provided me the continuing numbers through 2002–2003. These estimates must underestimate true population sizes by nearly a fourth because the censuses are not corrected for the 0.75 mean sightability bias.

THE NATURAL-REGULATION PERIOD: 1969–2003 The natural-regulation policy (Anonymous 1967a, b, c) had been announced in 1967 and the management protocol actually began in 1969 with cessation of park herd reduction. Hence I begin interpretation of the herd’s population trajectory with 1969, the first year without culling, and near its low point during the natural-regulation period. Herd trends thereafter have been represented by censuses during a 35-year period without National Park Service (NPS) control efforts, approximating natural-regulation management, but modified by varying sport hunting removals of animals moving outside the park in fall and winter. This is also the period of most effective population censuses. To construct as long a series as possible, I used the 1968–69 to 1974–75 censuses in Coughenour and Singer’s (1996) table 1, the Lemke et al. (1998) data for 1975–76 through 1994–95 in their table 1, and census and hunting

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values from 1996–2003 given to me by T. O. Lemke. These are shown in the second column of table 2.1. Because I did not at this point use any of the sightability correction factors mentioned, this time series is therefore a sequence of indices rather than actual estimates of population number (table 2.1). Houston (1982:20) and Lemke et al. (1998) calculated “minimum fall populations” by adding hunting removals to the winter censuses. The fall hunt occurs before the census, and all animals shot are added to the winter count. The late winter hunts typically begin before the census, and the animals shot before the count are also added to the census to complete estimation of minimum fall number (fifth column of table 2.1). I have based the following analyses on this index of the northern herd population because it is most comparable with calculations of other authors discussed next. Visually, the numbers in table 2.1 point to a continuous increase in the minimum fall population from 1968–69 to 1974–75. There is a suggestion of slowing growth or leveling off between 1975–76 and 1978–79, although the 1976–77 census was faulty. Whether this was actually growth cessation, and if so why—as I will point out, severe winters reduce population size—or faulty censuses is not known. But the 16,473 index in 1981–82 implies that increase continued through the uncensused years of 1979–80 and 1980–81 because the 48% increase between 1978–79 and 1981–82 would be unlikely, biologically, in 3 years. Mack and Singer (1992) pointed out that the low censuses in the latter 1970s gave the impression of equilibration, and Coughenour and Singer (1996) commented that these censuses “probably biased [earlier] mathematical estimates of K downwards.” The lack of three censuses in the early 1980s obscures interpretation of population trend in that decade, but those in the following three years point to continuing increase through 1987–88 to 19,316. Thus a reasonable interpretation of the trend over this 20-year period is continuous growth, with perhaps some slowing in the mid-1970s and occasional years with short-term fluctuation. Increase in the winter censuses from 4,305 in 1968–69 to 18,913 in 1987– 88 is 4.4×. Increase from the 1967–68 census low of 3,172 to the 1987–88 level is 6.0×. To estimate the northern herd’s equilibrium population-index size, I regressed annual instantaneous rates of change (r) in minimum fall populations between year t and year t + 1 on the fall populations of year t. And r is calculated by r = logeNt+1–logeNt, and regressed on Nt. (Note: I use the symbol r in this book as a measure of annual, instaneous rate of change. I use the symbol R to represent the correlation coefficient. To avoid confusion, I will use the lowercase r for the rate and capital R for the coefficient.) To minimize measurement error and maximize precision of estimate, I did not use the aberrant census values for 1976–77, 1988–89, 1990–91, and 2002– 2003. Because each rate estimate requires values for two consecutive years, omitting a single, annual value prevents calculation of two rates. Missing an annual census and hunting kill numbers incurs the same problem. Thus, because of the 4 aberrant years, lack of censuses in the early 1980s and 1995–97,

The Census Period: 1923–2003 Table 2.1 Winter Censuses, Hunting Removals, Minimum Fall Population Estimates, and Hunting Kill as Percentages of Fall Populations, 1968–69 to 2002–2003

Year

Winter Censusa

Corrected Censusb

Hunting Offtakec

Min. Fall Pop.d

Est. Fall Pop.e

Hunting Kill as % of Fall Nf

1968–69 1969–70 1970–71 1971–72 1972–73 1973–74 1974–75 1975–76 1976–77 1977–78 1978–79 1979–80 1980–81 1981–82 1982–83 1983–84 1984–85 1985–86 1986–87 1987–88 1988-89 1989–90 1990–91 1991–92 1992–93 1993–94 1994–95 1995–96 1996–97 1997–98 1998–99 1999–00 2000–01 2001–02 2002–03

4,305 5,593 7,281 8,215 9,981 10,529 12,607 12,014 8,980g 12,680 10,838 No count No count 16,019 No count No count No count 16,286 17,007 18,913 8,739g 14,829 9,456g 12,859 17,585 19,045 16,791 No count No count 11,692 11,742 14,538 13,400 11,969 9,215g

5,740 7,457 9,708 10,953 13,308 14,039 16,809 16,019 — 16,907 14,451 — — 21,359 — — — 21,715 22,676 25,217 — 19,772 — 17,145 23,447 25,393 22,388 — — 15,589 15,656 19,384 17,867 15,959 —

46 50 82 149 265 316 252 1,529 219 1,067 341 661 376 1,359 1,881 2,061 1,571 1,498 1,739 579 2,896 1,299 1,005 4,515 2,055 527 2,538 — — — — 1,102 1,509 1,273 —

4,351 5,643 7,363 8,364 10,246 10,845 12,859 12,354 9,199g 12,941 11,149 — — 16,473 — — — 16,885 17,901 19,316 11,148g 15,805 10,287g 15,587 18,066 19,359 17,290 — — — — 15,640 14,909 13,242 —

5,786 7,507 9,790 11,108 13,573 14,355 17,061 17,548 — 17,974 14,792 — — 22,718 — — — 23,213 24,415 25,796 — 21,071 — 21,660 25,502 25,920 24,926 — — — — 20,486 19,376 17,098 —

0.8 0.7 0.8 1.3 2.0 2.2 1.5 8.7 — 5.9 2.3 — — 6.0 — — — 6.5 7.1 2.2 — 6.2 — 20.8 8.1 2.0 10.2 — — — — 5.4 7.8 7.4 —

aCensuses

for 1968–69 through 1974–75 from Coughenour and Singer (1996), for 1975–76 through 1994–95 from Lemke et al. (1998), for 1996–2003 from T. O. Lemke (personal communication). bWinter census/0.75. cFall hunt plus winter kill before census. dWinter census plus hunting offtake. eCorrected censuses and hunting offtake. fTotal hunting offtake as percentage of estimated fall population. gYears with faulty censuses and therefore inaccurate population numbers.

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and the lack of hunting-kill data for 1997–98 and 1998–99, I could only calculate 15 r-values for this 35-year period. My test for the entire 15 years (figure 2.1) shows a negative relationship significant at p = 0.05 with an R2 = 0.398. The r = 0 intercept of the regression line is 16,800 elk. Taper and Gogan (2002) performed similar calculations, subdividing the time series into a 1964–65 to 1978–79 “early” period and 1985–86 to 1994–95 “late” period, and with what appear to be different equilibrium levels for the two periods. The higher level of the late period coincides in part with the expansion in recent years of the winter range. The latter result emerged both in terms of total population size and their conversion to density, an extremely interesting result, as I will discuss, if the pattern continues into the future beyond their five data points. I similarly divided the time series into early and late periods using the data in table 2.1. But I began my early period in 1968–69, 3 years later than did Taper and Gogan because this was the first year without park removals and is a better start for testing the natural-regulation hypothesis. I ended the early period in 1987–88 because the following year was the first with the large annual movements out of the park into the expanded winter range (Lemke et al. 1998; Lemke, personal communication, June 5, 2002), and is a better

Figure 2.1 Regression of annual rates of change, r, between year t + 1 and year t on minimum fall population of year t for the period 1969–2002. See text and table 2.1 for derivations and data sources.


starting point for the late period. My late period then becomes 1988–89 to 2001–2002. My test for the 10 early rates was significant at p = 0.05 with an R2 = 0.40. Test for the late period, based on only five points, also showed a negative slope but was short of significance. Hence it is not possible to ascertain with my analysis whether the early and late periods differed. Other authors have conducted similar tests for different sets of years. Merrill and Boyce (1991) found a similar inverse relationship for the period 1971–88. They provided no R2 value, but it was presumably smaller than mine because 77% of the variation in r was associated with variations in summer phytomass. Coughenour and Singer (1996) obtained an R2 of 0.45 from a similar test based on winter counts from 1968–69 to 1990–91. In another test, Cheville et al. (1998) calculated an R2 of 0.33 for the period 1969–97. Several points now bear exploration from these tests, and especially figure 2.1. The first obvious one is that the annual rates of change are density dependent. Beside the above r/N tests, several previous authors have documented density-dependent effects in the northern range herd (Houston 1974, 1982; Coughenour and Singer 1996; Singer et al. 1997b) including percentage of yearling cows pregnant, summer recruitment of calves, summer and winter calf mortality, and bull mortality. The causal mechanisms driving these demographic responses have been attributed to intraspecific competition for forage, and by Singer and colleagues to densitydependent summer predation on calves. In their elegant analysis, Taper and Gogan (2002) showed a difference in form between the density-dependent functions for fertility and adult survivorship. Hunting offtake was also roughly density-dependent during the 1968–95 period. The mean percentages of the fall population taken (table 2.1), again omitting the 4 aberrant census years, were 1.3 for 1968–69 to 1974–75, 5.7 for the 4 valid censuses between 1975–76 and 1984–85, and 7.9 for 8 years during 1985–86 to 1994–95. Mean, estimated fall populations for these three periods were 11,311, 18,258, and 24,063, respectively. This effect is perhaps to be expected because the tendency for the animals to move north out of the park has increased as the herd size has grown (Coughenour and Singer 1996; Lemke et al. 1998) and thus become increasingly vulnerable to hunting. But the total density dependence operating on the population only accounts for 23–55% of the variance in r. Thus, perhaps half to two-thirds of the variance is associated with density-independent influences in the environment and with measurement error. Both Merrill and Boyce (1991) and Coughenour and Singer (1996) found a major portion of the variance associated with summer phytomass or water-year precipitation, which undoubtedly is expressed in forage production. Neither Coughenour and Singer nor Singer et al. (1997a) were able to find any correlation between winter mortality and indices of winter severity. The former authors reasoned that failure may be a function of the insensitivity of the indices because there have been years of major loss associated with extreme weather. Indeed Singer et al. (1989) estimated that 3,021–5,757 elk died during

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the severe winter of 1988–89, and Lemke (personal communication, February 12, 2003) surmises that 4,000–5,000 may have died in the winter of 1996–97. In both cases the winter censuses declined in subsequent years. These winters occurred in years of high population levels, which confounds distinction between weather and density-dependent effects. It is this density independence, plus stochastic variation in the densitydependent factors themselves—for example, the hunting removals have been quite variable within the periods averaged above—that produce the scatter about the Figure 2.1 regression line. There must also be some measurement error in the censuses, although I have omitted from the tests the years with obviously aberrant counts. A second set of insights can be deduced from the equilibrium state. I define equilibrium in a fluctuating population in both a state sense—the absence of net or mean trend in numbers over a specified period of time—and a rate sense— the mean rate of change over a period of time r\ = 0. Coughenour and Singer (1996) comment that density-independent fluctuations in forage supply prevent the northern herd from achieving a “static equilibrium” or “steady state.” I believe the term stationarity is more appropriate for the total absence of change to which I infer the authors are referring, as it can be argued that the term steady state can also be applied to the equilibrium achieved by the northern herd. The number of animals at which, on average, a population goes to equilibrium, or equilibrates, or achieves r = 0, is the equilibrium population size. I choose not to use the term carrying capacity for this parameter because of the ambiguities surrounding the diverse uses of the term and the lack of empirical basis for some of them, as discussed by Wagner et al. (1995a:114–27). No population fluctuating in an environment under the influence of density-independent factors will achieve stationarity. That does not negate the reality of dynamic equilibria maintained by density dependence. Each year, the r value is a function of the aggregate forces acting on a population, both density dependent and density independent. A population at its equilibrium size will increase (r will be positive) in a year when the densityindependent factors ease their effects. This can occur in more than 1 year of favorable density independence. Thus, although the equilibrium level for the minimum-fall-population index was 16,800 during 1968–69 through 2001– 2002, the population index rose to 18,066 and 19,359 in 1992–93 and 1993– 94 (table 2.1). But as a population increases, it incurs rising density-dependent pressures (figure 2.1), which eventually force r to the negative despite all but extremely favorable density-independent pressures, and move it back toward its equilibrium point. Moreover, there is no implication here of a population occurring at its equilibrium level for much or most of its time. In a variable environment, it may fluctuate to precisely that number rarely. Using different series of years, either fall or winter numbers, and either correcting for hunting removals or not doing so, other authors have calculated the northern herd, census-based equilibrium level: Merrill and Boyce (1991) at


14,522; Coughenour and Singer (1996) at 13,000; Cheville et al. (1998) at 17,400; and Taper and Gogan (2002) at about 12,500. With a state-structured model using demographic parameters, Coughenour and Singer (1996) calculated the equilibrium population of “sighted” animals at approximately 16,400. As Coughenour and Singer point out, all of these values based on the censuses underestimate the true population size because of the mean 0.75 sightability bias. They conclude that the true population value for their estimate is around 22,000. My 16,800 for 1968–69 through 2001–2002 minimum fall populations, divided by 0.75, becomes 22,400. The estimated population size was consistently above 20,000 between 1980–81 and 1999–2000 (table 2.1). Several authors (Mack and Singer 1992; Coughenour and Singer 1996; Lemke et al. 1998) have reported increased winter migration outside the park since the 1970s with the total area now occupied having increased by as much as 41% (Lemke et al. 1998). These authors have speculated on whether this enlarged winter range area has increased the equilibrium population size. Using different series of years, and the winter censuses, Taper and Gogan (2002) calculated higher equilibrium numbers and densities in their late period than in their early period. They only had five data points and a narrow range of numbers on the x-axis for the late period. As already commented, my own analysis of minimum fall numbers (figure 2.1) with slightly different time periods showed essentially no difference. The data are too few, and too variable, to choose between the two results. But if the Taper and Gogan results hold in future years, the increased density could suggest that the winter range outside the park to which much of the herd now moves, largely ungrazed most of the twentieth century, is superior to the park forage that has been heavily used over the same period. I will present evidence in chapter 9 that the northern herd is under winter nutritional duress down to numbers approaching 5,000–6,000. Taper and Gogan (2002) show fertility declines down to herd sizes approaching 3,000. There is some indication that the herd size has declined since 1994–95 (table 2.1), the last year of the Taper and Gogan analysis. This complicates any inference on the equilibrium level and may explain the difference between theirs and my results for our late periods. Lemke (personal communication, February 12, 2003) states that recent years have been a drought period, and elk populations have declined over much of western Montana. Hence any northern herd decline during this period may be part of a more general phenomenon and not unique to the park. Additionally, the decline may be induced in part by the new wolf population. A final inference to be drawn from figure 2.1 is the declining resilience of the population to additional impacts from environmental factors—whether severe weather conditions, additional hunting removal, or predation—as it approaches its equilibrium size. At any given population level, the distance on the y-axis between the r = 0 line and the r/N regression line is a measure of how much additional impact from environmental factors the population could sustain, on average, without being forced to decline. The r-value differences at

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minimum fall population sizes of 5,000, 10,000, and 15,000 are 0.3, 0.14, and 0.05, respectively. Converted to percentages by % = (1 – antiloger) ×100 these become 26, 15, and 5.

THE 1923–68 PERIOD The comments that follow in this section are based on the data in Houston’s (1982) tables 3.1 and 3.2 and my calculations from those data in my table 2.2. The northern herd was only censused 5, 9, 4, 2, and 5 years during the 1920s, 1930s, 1940s, 1950s, and 1960–61 through 1967–68, respectively. Except for an aerial count in 1935, all of the early counts up to 1956 were ground counts that Houston (1982:19) calculated to underestimate the actual number by a mean of 22%. I have estimated actual fall populations by correcting the winter censuses with Houston’s 22% for the years 1922–23 through 1948–49, Mack and Singer’s (1992) 0.76 for 1955–56 through 1967–68, and adding hunting and park removals to the corrected values (table 2.2). Houston (1974:18) commented that all censuses from 1935–68 were made after the hunting seasons and park removals. The mean, estimated fall populations for the censused years of the 1920s, 1930s, and 1940s were 14,624, 13,918, and 14,239. The mean, annual hunting removals for these same decades were 3.7%, 6.9%, and 20.1% of the estimated fall numbers. On average, the population evidently maintained its numbers in the face of these hunting removals. But the averages do not disclose the detailed trends in the population. The gaps in the data complicate interpretation and make it tentative. But provisionally, the hunting removals were highly variable, falling below 5% in some years, exceeding 15–20% in some. The population appears to have increased in series of years when removals fell below 11–15%, then held its numbers or declined following one or a series of years in which removals exceeded 11–15% (table 2.2). Thus there is some suggestion of population increase in the 1920s until an 11.6% removal in 1928; increase through the early 1930s until 3 out of 4 years with removals ranging from 14.3% to 20.0% from 1935–38. With only 4 years of census in the 1940s, it is difficult to discern a trend. But 3 of those years experienced >15% removals, and there may have been something of a decline or at least absence of trend. By the 1950s the park intensified its removal efforts and strengthened them further in the 1960s. At this point, the herd was clearly reduced, reaching its low point with a censused 3,172 in 1968, the last year of park reductions. In sum, the population evidently maintained its numbers during the 1920s through the 1940s in the face of outside hunting removals which in occasional years exceeded 15–20% and set it back to some degree, but in most years fell below 10% and allowed recovery. The approximate result for the 3-decade period was stasis.


Table 2.2 Estimated Fall Populations and Hunting Kills as Totals and as Percentages of Fall Populations, 1922–23 to 1967–68 Year

Winter Census

Corrected Hunting Censusa Offtake

Park Est. Fall Hunt. Off. as % Removal Populationb of Fall Pop.c

1922–23 1924–25 1926–27 1927–28 1928–29

11,648 12,428 12,488 8,959 9,122

14,933 15,933 16,010 11,485 11,695

33 366 719 1,529 15

49 59 107 187 0

15,015 16,358 16,836 13,201 11,710

0.2 2.2 X|%=3.7 4.3 X| pop.=14,624 11.6 0.1

1929–30 1930–31 1931–32 1932–33 1933–34 1934–35 1935–36 1936–37 1937–38

8,257 7,696 10,624 11,521 10,042 10,112 10,281 8,794 10,976

10,586 9,867 13,621 14,771 12,874 12,964 13,181 11,274 14,072

212 316 290 177 136 2,598 2,287 257 3,587

110 2 37 2 11 667 557 574 236

10,908 10,185 13,948 14,950 13,021 16,229 16,025 12,105 17,895

1.9 3.1 2.1 1.2 1.0 X|%=6.9 16.0 X| pop.=13,918 14.3 2.1 20.0

1942–43 1945–46 1947–48 1948–49

8,235 8,513 7,815 9,496

10,558 10,914 10,019 12,174

6,539 2,094 970 2,837

691 73 39 49

17,788 13,081 11,028 15,060

36.8 16.0 X|%=20.1 8.8 X| pop.=14,239 18.8

1955–56 1958–59

6,963 4,884

9,162 6,426

3,900 372

2,635 1,334

15,697 8,132

24.8 X|%=14.7 4.6 X| pop.=11,910

1960–61 1961–62 1964–65 1966–67 1967–68

8,150 5,725 4,865 3,842 3,172

10,724 7,533 6,401 5,055 4,174

25 125 1,012 1,108 116

1,434 4,619 892 1,540 984

12,183 12,277 8,305 7,703 5,274

0.2 1.0 X|%=6.0 12.2 X| pop.=9,148 14.4 2.2

aWinter

census/0.78 for 1922-23 through 1948–49, and divided by 0.76 for 1955–56 through 1967–68. Winter censuses from Houston (1982:15–17). bCorrected census + hunting offtake + park removal. These are actual population estimates. c(Hunting offtake/est. fall population) x 100.

Not until the 1950s and 1960s, when intensified park reductions approached or exceeded the magnitude of hunting removals, and together they exceeded 20– 25%, was the herd progressively reduced to the 1968 low. Outside hunting alone would not have achieved it, nor would the park effort by itself have done so except with the especially intensified removals. Even then, the reduction was somewhat stepwise by virtue of variations both in hunting kill and magnitudes of park removal. These removal effects on population trend indicate that they were partially if not largely additive. The population evidently declined with removals

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>20–25%. That lesser removals did not cause decline does not indicate lack of additivity, which I define here as absence of effect on r-values. Figure 2.1 shows that at 10,000 animals, the population could withstand an additional 15% reduction in r without being forced to decline, but with reduction in r and magnitudes of population increase nevertheless.

DISCUSSION The calculations of this chapter make it possible to test the population aspect of the natural-regulation hypothesis. (As discussed in chapter 1, there were two major aspects to the hypothesis: the population behavior and the effects on the northern range ecosystem). But first, some further definition to ensure semantic and communication clarity (compare Wagner et al. 1995a:127–34 for sources and elaboration). The term regulation is one of the most ambiguously and inexplicitly used terms in ecology. But a substantial number of population ecologists and literature have converged on the principle that regulation is the density-dependent maintenance of equilibrium by density-dependent factors once that level has been attained. In rate terms, it is the maintenance of r\ = 0. As a noun, it is roughly synonymous with stabilization. The northern herd only reached the above-cited equilibrium population sizes of the latter half of the twentieth century by the 1980s, and only then could be regulated. The r/N relationships, like that in figure 2.1, are persuasive evidence that the population was regulated during the 1980s and 1990s. Determination of the population size (or density) at which a population achieves equilibrium is appropriately termed limitation. It is effected by the combined action of both density-dependent and density-independent influences. In rate terms, limitation is the reduction of r to zero at the equilibrium population size. The northern herd was limited in the 1980s and 1990s, according to my calculations, to a mean, minimum fall equilibrium size of approximately 16,800 but an actual number when corrected for sightability bias in the censuses of 22,400 or more. I agree with Coughenour and Singer (1996) that no further mean population increase was to be expected as long as the complex of factors operating at that time remained the same. The recent expansion of the winter range could allow some increase, but wolf predation could work in the opposite direction. There are three, basic tenets to the population aspect of Houston’s (1971) natural-regulation hypothesis: (1) The population would be limited without human intervention. (The process Houston termed regulation, I infer, is actually limitation.) (2) The equilibrium population size was first postulated to be 6,000–9,000 (Houston 1971), later revised to the range 12,000–15,000 and averaging about 12,000 (Houston 1974), and still later at 15,000 (Houston 1982:66). (3) The population would not erupt.


The first tenet has become true. But this was essentially a foregone conclusion. As Wagner et al. (1995a:132) commented, there are limits to the resources of any area, and no population can continue growing indefinitely. Second, hypothesized equilibrium population size was revised upward as the population grew, and three were proposed. Each has been surpassed as the population grew to the levels shown by the recent r/N tests, even without correcting these upward for the sightability bias. This tenet of the hypothesis must be considered falsified. Finally, it is not clear whether the third tenet of the population aspect of the hypothesis—that the population would not erupt—can be tested definitively because Houston (1971) never defined what he meant by an eruption. To some observers, the 6× increase from 3,172 censused animals in 1968 to 18,913 in 1988, or the 2.6× increase from 7,281 in 1971 to the 18,913 in 1988 would constitute eruptions. Seemingly, Houston had nothing like this in mind in 1971 when the hypothesis was proposed and he was projecting a 6,000–9,000 equilibrium population. Caughley’s (1970) definition of eruption—“an increase in numbers over at least two generations, followed by a marked decline”—is not helpful. This simply describes most fluctuating populations. Leopold (1933:50), who may have coined the term originally, defined it as “severe but irregular fluctuations of no fixed amplitude . . . fluctuations of over 50 percent from average or normal density.” This, too, is not helpful, also describing many fluctuating populations. If the density prevailing before park formation, which I will discuss in the next chapter, is considered average or normal, the increase of recent years is far beyond that level. To some authors, the abundant evidence of density dependence in the herd, and the fact that it has equilibrated, imply the validity of the hypothesis. But the foregoing analysis shows that the nonobvious tenets of the population portion of the hypothesis have been falsified by the herd’s trend since it was proposed. This only bears on the population aspect. The second aspect is the absence of significant effects on the northern-range ecosystem. Coughenour and Singer (1996) recognize this dual nature of the hypothesis but conclude that there has been no major effect on the vegetation, as have a number of other authors. I will address that aspect in part III of this book. Two final points bear mentioning. The size range to which the population rose in the latter 1980s and early 1990s—22,000 to 25,000, now corrected for the sightability bias—is similar to the numbers which park officials were reporting for the population around the turn of the twentieth century before herd control was instituted. Those numbers were challenged by Houston (1974, 1982) and subsequent park authors, as I will discuss in chapter 4. That the herd could rise to these numbers in recent years using a vegetation that had been heavily grazed for nearly a century lends credibility to the early accounts. Secondly, park publications have pointed out that northern range winters moderated in severity during the twentieth century, perhaps explaining the levels

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to which the contemporary population has risen (compare Meagher and Houston 1998). But the herd rose to a range of 20,000–35,000 around the turn of the century, as I will argue in chapter 4, during the earlier period when the indices showed the winters to have been more severe. If moderating winters have eased weather pressures on the population, they may only have offset to some degree the nutritional pressures imposed by decades of grazing alteration in the forage.

Prehistory to the 1880s

3 Nothing in biology makes sense except in the light of evolution. —Theodosius Dobzhansky

THE PRISTINE STATE? One of the pivotal points in the natural-regulation debate has been the abundance of elk in the Yellowstone area at the time the park was established in 1872. Early photographs and reports of vegetation condition and abundance of other animals raise questions about the number of elk that would permit such unmodified conditions to persist. Moreover, the natural-regulation hypothesis posed a northern herd equilibrium level or “carrying capacity” of about 12,000–15,000, with predation playing a “nonessential” role in establishing this number. If such a carrying capacity obtained for the twentieth century, consistency on the part of park scientists called for a population of roughly this same size wintering on the northern range prior to 1872, and in the presence of an uncontrolled predator guild. Yet this number of animals obviously impacted the northern range in the 1900s. Early park officials also reported that the herd migrated out of the northern range in winter. Houston (1982) disagreed with this interpretation of the evidence, inferring that large numbers of elk wintered on the northern range prior to park establishment, also consistent with the hypothesized situation under natural regulation. In this chapter I examine a wide range of evidence bearing on these questions.

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ARCHAEOLOGICAL EVIDENCE In an effort to use archaeological data as one source of evidence on the early abundance of elk in the Yellowstone area, Kay (1990) examined the entire archaeological literature for the region. At the time he wrote, 274 archaeological sites had been cataloged in what is now Yellowstone National Park (YNP); more than 300 in Jackson Hole, Wyoming, south of the park; and over 1,000 in other parts of what is now considered the Greater Yellowstone Ecosystem (GYE). No major sites had been excavated in what is now the park, although Hadly (1990) had explored a small cave for which packrats and predators had been the main taphonomic agents with no sign of human activity. Archaeologists have concluded that the GYE was inhabited by Native Americans beginning some 10,000 years BP. Kay analyzed the ungulate remains unearthed at four sites ranging from 4 to 80 km north and east of YNP according to archaeologists’ parameters MNI (minimum number of individuals [animals]) and NISP (number of individual specimens [portions of animals]). Among 313 ungulate MNI and 3,708 NISP, elk were only 5% and 3%, respectively. Deer and bighorn sheep together made up 75% and bison 15% of MNI, and 59% and 37%, respectively, of the NISP. Kay contrasted these results with the contemporary makeup of the GYE ungulate populations in which elk comprise, by his estimate 79%. Archaeologist G. C. Frison (1971) had earlier explored the Jackson Hole area, finding bison and mule deer remains but no elk. In total, ungulates were a minor part of the native people’s diet, with elk no fraction of that part. Frison (1971) commented on the “strong orientation toward Archaic plant gathering” among early inhabitants of the Wind River and Bighorn basins south and east of Yellowstone. He remarked on the scarcity of elk among the ungulate remains in archaeological sites and the absence of evidence of traps and communal procurement practices. Archaeologist G. A. Wright (1982) spent 10 years excavating “more than two dozen sites” in the Jackson Hole area. He found some bison, mule deer, and bighorn bones, but no elk in an area where today 2,000–4,000 elk summer on the floor of Grand Teton National Park on the west side of Jackson Hole and where 7,000–10,000 winter on the adjacent National Elk Refuge (Boyce 1989). Wright (1984) proposed a subsistence model of the early inhabitants called a high country adaptation. Hunter-gatherer cultures depended heavily on plant resources. Fishing was also important between late spring and early fall. Winter was spent outside the area hunting ungulates, the inhabitants of the park area moving north to the plains in search of bison. Jackson Hole occupants moved southeast in winter to the Green River–Red Desert area or west into Pierre’s Hole. No evidence was found that large numbers of elk occurred in Jackson Hole in either winter or summer. Kay (1990), in addition to quantifying the ungulate remains in archaeological sites around the Yellowstone area, analyzed data from sites throughout the Intermountain West. The results were comparable over the entire region: Ungu-


lates were a minor component of the natives’ diets, and elk were a small percentage of the combined and limited ungulate remains. Kay (1990, 1994a) used the archaeological findings as one source of evidence to support his aboriginal overkill hypothesis proposing that ungulate populations in the Intermountain West of the United States, particularly elk and moose, occurred at low densities in pre-Columbian times. This was achieved by a combination of aboriginal hunting and predation by large carnivores. Kay et al. (2000) marshaled similar evidence for western Canada. A number of authors—most of whom have been closely associated with YNP, and only two of whom are archaeologists—have questioned Kay’s inferences from the archaeological evidence (Barmore 1987; Hadly 1990; Cannon 1992; Schullery and Whittlesey 1992; Boyce 1998; Klein et al. 2002) on three grounds. Some emphasize that the results show elk to have existed in the area for millennia, a point never questioned by Kay. Second, some argue that the archaeological remains cannot be used for estimating animal numbers, something Kay never attempted other than to infer, along with other evidence, that the dearth of ungulates in the natives’ diet indicates relative scarcity. Third, some authors protest on taphonomic grounds that the ungulate species composition in the sites cannot automatically be taken to indicate, as Kay inferred, the composition prevailing in the pre-Columbian populations. This is a composition markedly different from the contemporary one. In fact, Kay addressed the taphonomic issue at some length. More generally, these are one-sided, negative arguments seeking to falsify Kay’s hypothesis. They do not grant the possibility of the positive alternative that the site composition could roughly reflect the prevailing abundance and makeup of the ungulate evidence that tends to support his inferences.

HISTORICAL EVIDENCE Inferences of Scarcity by Early Biologists Because there were no censuses of winter elk numbers on the northern range in the 1870s and before, all attempts at approximating them, including the one proposed shortly, have been based on interpretation of anecdotal reports by early trappers, miners, explorers, hunters, tourists, and military personnel in diaries, journals, oral accounts, and a variety of publications. There is a consistent core of accounts used by all authors, but some draw on a wider range of sources than others. The differing impressions of abundance result from which accounts have been used, and how they have been analyzed. M. P. Skinner (1927), the first park biologist, cited accounts in the Lewis and Clark journals of game abundance on the plains but scarcity once the party reached the mountains. Skinner also cited reports from the 1870 Washburn, Langford, and Doane expedition through the Yellowstone area to the effect that they “found very few animals.” He inferred from the accounts of the 1871 Hayden survey that the party only saw one deer.

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Biologist W. M. Rush (1932) quoted at length from the Lewis and Clark journals on game scarcity in the mountains and Native Americans subsisting on foods other than meat. When they met the Shoshoni Indians at the headwaters of the Jefferson River, “even the Indians had only salmon and berry cakes to trade them.” More recently, Martin and Szuter (1999) cited the Lewis and Clark accounts of abundant game on the plains and scarcity in the mountains and westward along the Columbia River. The authors emphasized that regions of game abundance on the plains coincided with Indian nations at war where hunters were reluctant to venture freely and far from tribal centers.

Later Analyses Inferring Abundance Murie (1940) was apparently the first agency biologist to question the early impressions of scarcity. Citing accounts from some of the same expeditions as Skinner’s (1927), Murie pointed out that some of these reports included accounts of large numbers of animals. He argued that positive evidence must take precedence over negative, and failure to see game could not be taken as certain indication of its absence. He compiled 17 early accounts of wildlife abundance and concluded that it must have been abundant in the Yellowstone area during the 1800s. Houston (1982) cited 20 accounts of wildlife numbers from 1836 through 1881, many of the same sources used by Skinner and Murie. All his sources reported seeing and/or shooting elk, with accounts ranging from individual animals to “scattered flocks, large band, everywhere, abundant, all manner,” and so on. YNP historians Schullery and Whittlesey (1992) compiled the most extensive set of early accounts, reviewing statements from 168 sources for the period 1806–81. Reports of 131 sightings of individual elk, “small groups,” and “herds”; 9 of sounds; 25 of meat, hides, bones, or antlers; 31 of tracks, trails, or other signs; and 84 general statements of presence totaled 280. Of the 56 statements by observers who commented one way or the other on game abundance in the park area, 91% noted that game was, in Schullery and Whittlesey’s words, “very abundant.” These authors concluded that “elk were common throughout the Park [prior to 1882], and were observed at various times in large numbers in virtually every part of the park where large numbers now occur” but “the record is not sufficiently detailed . . . to allow us to say with any confidence that elk numbers on the Northern Range during any given year in that period equaled, exceeded, or were less than, at present.” In reading through this extensive compendium, one develops a sense of great elk abundance from the sheer number of reports. That sense is strengthened by the frequent allusions to abundance of other wildlife: six other species of ungulates, beaver (Castor canadensis), numerous species of carnivores, flocks of waterfowl, and streams brimming with trout. But with further consideration, that sense is tempered with several reservations:


1. The perspective of these early observers cannot be judged from our own. What they considered abundance may or may not match our experience. One cannot get any reasonably quantitative sense from subjective descriptions used more than a century ago by people who had very different outdoor experiences from our own. 2. As Kay (1990) pointed out, some observers exaggerated. Schullery and Whittlesey conceded that in his frequent use of the word thousands, A. Bart Henderson probably meant something more like “lots.” 3. More significantly, Murie, Houston, and Schullery and Whittlesey set out primarily to compile reports of game abundance, evidently to support the contention that wildlife was abundant in the YNP area in the 1800s. In so doing, they almost certainly did not compile comments on game scarcity, inability to shoot game for food, and food shortage. Murie and Houston reported none; Schullery and Whittlesey mentioned only a few. Hence their compendia are based on selected evidence. I agree with the latter authors and Murie that focusing only on anecdotal reports of failure to see wildlife cannot provide an objective indication of its absence or necessarily its scarcity. But it is equally true that reporting only observations of abundance, as Murie and Houston did, biases information in the opposite direction. When reports of scarcity are numerous from large parties of mounted, experienced outdoorsmen who depended on shooting game to provision lengthy excursions, they begin to carry weight and discount reports of game abundance to some degree. Kay (1990) raised this issue in pointing out Houston’s (1982) reference to two passages from prospector Walter DeLacy’s journal of his 1863 trip. Houston cited two passages in which DeLacy observed elk, but did not point out that these were the only reports of elk by a party of 25–40 prospectors on a 27-day excursion. Thus any balanced assessment depends on equally thorough consideration of abundance and scarcity reports and some form of quantification based on temporal and/or spatial scaling.

Quantifying the Historical Record Aware of the difficulties in deriving any reliable quantitative sense from anecdotal reports, Kay (1990) devised a means for quantifying such observations prior to Schullery and Whittlesey’s compendium. Kay analyzed the journals of 20 expeditions through the park area in the 1800s on what he called a “continuoustime basis.” He followed the historian’s discipline of using only first-person accounts recorded in journals and diaries at the time of observation, and tabulated these in terms of occasions on which wildlife was seen, and number of references to lack of game or lack of food, all in terms of the time period of the excursion. His results were: 1. The 20 parties varied in size from 3 to 60, averaged about 20, and together spent a total of 765 party-days in the Yellowstone area. The

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mean number of days per party was 765/20 = 38, and the total number of person-days was about 765 × 20 = 15,300. 2. The 20 parties reported seeing elk on 42 occasions, or approximately two times per party and once per 18 party-days (765/42). Kay did not attempt to estimate the number of elk seen because there was no way to determine the number of animals in a reported herd or group. 3. The parties recorded lack of game or lack of food 45 times, or approximately the same number of times they reported seeing elk. 4. The parties reported seeing ungulates of six species on 121 occasions. Elk observations constituted 35% of the total ungulate observations. Mule deer (Odocoileus hemionus), pronghorn (Antilocapra americana), and bighorn sheep (Ovis canadensis), in that order, were the next most common ungulates seen. Kay concluded from these calculations that elk were “not abundant throughout the Greater Yellowstone Area during the period of early historical records.” Moreover, as with the archaeological evidence, they constituted a much smaller fraction of the collective ungulate population than they do today (79%). In an effort to examine whether there might be significant variations in the species’ visibility, affecting their frequency of observation, Keigley and Wagner (1998) evaluated five factors that would potentially affect sightability. They concluded that these factors tended to offset each other among the species, and that variations in visibility were not likely to have been serious impediments for seasoned hunters searching for the animals. Schullery and Whittlesey (1992) implied that because their compilation was so much more extensive than Kay’s, their conclusions of game abundance were more valid than Kay’s inference of scarcity. The data sources were quite different between the two studies. Schullery and Whittlesey did not restrict their sources to first-person, on-the-spot accounts, as did Kay; they used some newspaper and magazine sources, published accounts after time lapses of months to years, and some secondhand reports. Nevertheless, I will accept these at face value and analyze them according to the Kay procedures. The results are: 1. Schullery and Whittlesey’s 168 sources reported seeing elk on 131 occasions, or only 0.78 times per source. This is only about one-third as often as Kay’s twice per party. I did not attempt to tabulate the total source- or party-days because of the greater variety and indirectness of the authors’ sources. 2. Schullery and Whittlesey’s sources reported seeing ungulates on 330 occasions. Thus the 131 elk occasions constituted 40% of all ungulate observations, not very different from Kay’s 35%. Mule deer, pronghorn, and bison (Bison bison), in that order, were the next most frequently seen ungulates other than elk, again similar to Kay’s results except that bison replaced bighorn sheep at fourth in the order.


3. Although they reported seeing signs of or hearing wolves and mountain lions (Felis concolor), Kay’s 20 parties did not report seeing either species during their 765 party-days. This contrasts with the contemporary situation where wolves are readily seen, even by casual tourist observation, on the northern range since their restoration by the park in 1995. Kay considered the failure of his seasoned outdoorsmen to see carnivores to be another indication of low ungulate numbers because in other parts of the world with large ungulate (prey) populations (e.g., East Africa, the early North American plains, and now the YNP northern range), large predators are numerous and readily seen. Schullery and Whittlesey’s sources did report seeing wolves and mountain lions. But again when quantified on a per source basis, their evidence tends to support Kay’s position. Their 168 sources reported seeing wolves on five occasions, mountain lions on three. Over 95% of their sources failed to report seeing these two species, supporting Kay’s inference of low abundance. More generally, Kay’s methods of quantifying historical reports provide a more objective and accurate picture of early wildlife abundance than subjective appraisal of anecdotal accounts. And those methods point to relatively low ungulate densities.

Was the Northern Range a Major Elk Wintering Area in Prehistory? One aspect of the northern range debate has been the question of whether elk occupying the higher elevation summer ranges of the park area, whatever their numbers, migrated north to lower elevation winter ranges beyond present park boundaries before its establishment in 1872. By some accounts, elk migrated down the Yellowstone River valley to the north beyond the present town of Livingston, Montana (89 km from the park’s north entrance), as far as 100–300 km (Shore 1912, Graves and Nelson 1919; Skinner 1928; Bailey 1930; Cahalane 1941). Park historian A. L. Haines (1977) wrote “Crow Indians who lived along the lower Yellowstone River . . . [called it] ‘E-chee-dick-karch-ah-shay’ . . . which means ‘Elk River,’ a name derived from the fact that it provided a migration route for those animals while passing between their summer range on the Yellowstone Plateau and their wintering ground at lower elevations.” Park Superintendent M. Harris (1887) commented 15 years after park establishment about professional hunters camping outside the park boundaries “to intercept the game when driven out of the mountains by the deep snow, it seeks the lower valleys.” Klein et al. (2002) quote two early (1882, 1926) unpublished sources to the effect that “the buffalo, deer, and elk were accustomed to living on this plateau during the summer. In winter they migrated to lower and warmer regions outside the Park area until settlements of farmers in the country surrounding the Park made it impossible for them to use their long used winter homes.”

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Houston (1982) and Schullery and Whittlesey (1992) had rejected these accounts as indicating migration out of the northern range in winter, maintaining that elk had always wintered in large numbers in that area. But coupled (1) with the fact that early, prepark travelers largely moved through from May to October and there were few, if any, observations of wintering elk prior to 1872; (2) the penchant for elk elsewhere (including in the park) to move out of highelevation summering areas to distant, low-elevation wintering areas; (3) the above-mentioned early accounts; and (4) Keigley and Wagner’s (1998) more detailed analysis of this question, the weight of the evidence points to the validity of the early view.

POPULATION EVIDENCE OF EARLY NUMBERS Keigley and Wagner (1998) combined numbers of elk reportedly killed by market hunters, population responses, and known rates of elk population increase to approximate the number of animals in the northern herd at the time of park establishment. Until 1870, a few elk were shot in the YNP area mainly for subsistence, according to early accounts. From 1871 through 1877, larger numbers were killed by market hunters for their hides and tongues. Park Superintendent P. W. Norris (1875, 1877) and other travelers through the region reported that approximately 7,000 elk were shot during the period 1871–77, with 4,000 shot in the winter of 1874–75. (These numbers differ from Schullery’s [1997a] interpretations. I suggest consulting Keigley and Wagner (1998) for their resolution of the difference.) By the summer of 1875, Strong (1968) remarked that market hunting had reduced the game “until it is a rare thing now to see an elk, deer, or mountain sheep along the regular trail from Ellis [now Bozeman, Montana] to Yellowstone Lake.” Norris (1877) considered “the game in most of the park, especially along the main routes of travel, as too much decimated to justify extra efforts for its protection west of the Yellowstone Lake, River and Grand Cañon.” In 1877, Norris appealed to the market hunters to stop the slaughter. By 1879, hunting had eased and “choice animals” were observed to “have increased” in numbers. In 1883, it was reported “that between the Mammoth Springs and Cooke City [essentially the entire span across the northern range] there are at least 5,000 elk; that the yearly increase, to place a low estimate, is 1,000” (Anonymous 1883). Three years later Superintendent Harris (1886) reported that “there is more game in the Park now of every kind than was ever known before.” The following year, several thousand elk were estimated to winter in the Lamar River valley and its tributaries (Harris 1887). Wolf control began in the 1870s, and Norris (1881) stated that “their easy slaughter with strychnine-poisoned carcasses of animals have nearly led to their extermination.” Using these accounts and estimates of population growth rates, Keigley and Wagner (1998) approximated prepark numbers at 5,000–6,000 animals. These are of course crude estimates that depend on the validity of the early reports


and can only be considered order-of-magnitude approximations. But they do point to low numbers at park establishment in 1872 and are not suggestive of a population two to three times as numerous.

EVOLUTIONARY-ECOLOGICAL EVIDENCE In research entirely independent of the northern range debate, a number of plant ecologists have converged on the view that the northwestern U.S. vegetation west of the Rocky Mountains did not coevolve with heavy herbivorous pressure from large numbers of grazing animals. Mack and Thompson (1982) contrasted the morphology and consequent ability to withstand grazing of grasses east and west of the Rockies. They divided the two vegetations into the Bouteloua gracilis (blue grama) type of the Great Plains and the Agropyron spicatum (bluebunch wheatgrass, now Pseudoroegneria spicata) type of the northwestern United States. The Great Plains grasses, coexisting for millennia with the region’s bison herds, characteristically evolved a rhyzomatous growth form capable of spreading vegetatively and less dependent on flowering and seeding for reproduction. Species of the Agropyron type have erect, bunchgrass growth form that make them vulnerable to grazing, as would their dependency on flowering and seeding for reproduction. The authors considered that these morphological characteristics survived in the absence of heavy grazing pressure because “the steppe communities west of the Rockies . . . lacked large herds of mammals throughout the Holocene.” They also pointed out the low numbers of dung beetle species compared with the Great Plains and the prevalence of cryptogam soil crusts, which are vulnerable to trampling even by small numbers of ungulates. Caldwell et al. (1981) studied the ecophysiological responses of A. spicatum and A. desertorum, a grazing-tolerant bunchgrass introduced into western United States from Eurasia. A. desertorum allocated more plant resources to foliage growth following clipping experiments to simulate grazing. A. spicatum continued allocation of resources to root growth unabated, rendering it less effective in “coping with herbivory” to which it was not well adapted. Milchunas et al. (1988) generalized the effects of evolutionary history on the responses of grassland community structure to grazing by large herbivores. They considered the semiarid regions of northwestern United States and southwestern Canada to have had a short evolutionary history of grazing and as a result the grasses to have been unable to cope with heavy grazing. These basic ecological findings have been widely incorporated into rangemanagement theory. The American range literature (compare Young and Sparks 1985; Heitschmidt and Stuth 1991; Pieper 1994; Knapp 1996; Donahue 1999) repeatedly comments that the Intermountain West was not populated by large numbers of grazing ungulates in pre-Columbian times. Hence the vegetation had not evolved strong adaptations for tolerating hervbivory and was severely impacted by the introduction of European domestics.

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DISCUSSION Several independent lines of evidence point to low ungulate numbers in the Yellowstone region prior to park establishment. Each is circumstantial, and each depends on assumptions of unknown probability. No one line is unequivocal evidence of low pre-Columbian numbers. But all derive from independent scientific sources, and all converge on the same implication of low numbers. There is no significant body of contrary evidence except subjective appraisal of historic, anecdotal accounts, which, when quantified on a time scale, points in the same direction as the other evidence sources. The archaeological evidence indicates that ungulates were an inconsequential component of the natives’ diet in the region. Archaeologists are now widely adopting the ecologists’ optimal-foraging models (see authors in Kay and Simmons 2002). If highly ranked food sources like large ungulates had been abundant, they would have attracted a major amount of foraging time and would have been major dietary components. The archaeological evidence also indicates that elk were a smaller fraction of the ungulate guild in the region in pre-Columbian times than in contemporary populations. There were numerous accounts of ungulates by early travelers in the Yellowstone region. But when these are scaled on a temporal basis, the reports are actually infrequent. Moreover, reports of game or food scarcity have an equivalent frequency. Combined with rare observations of large carnivores, which implies low prey density, the collective historical evidence points to low ungulate abundance. And, like the archaeological evidence, the historical accounts indicate a lower fraction of elk than in today’s combined ungulate guild. In one small population instance, a market-hunting removal of 4,000 animals drove northern range elk numbers to rarity. This is a significantly greater impact than would have affected a population of 12,000–15,000 and suggests a northern herd in the range of 5,000–6,000. Finally, several plant ecologists have pointed out that western North American grasses did not evolve the morphological and physiological adaptations that would enable them to tolerate herbivory, as have North American Great Plains and East African species that have coevolved with large numbers of grazing animals. As a result, the western vegetation was seriously impacted by the introduction of domestic livestock. Why ungulates occurred in low numbers in pre-Columbian times in the YNP area can only be speculated on. Wagner et al. (1995a) review a number of cases in which ungulate populations are held to low densities by large carnivores. Barmore (1987) concedes that “it seems likely that an intact predator fauna influenced the size and dynamics of several ungulate species populations and the interrelationships between them [in the YNP area] to an unknown degree, but more than previously thought.” In addition, Kay (1994a) has postulated that Native American hunting significantly reduced ungulate numbers before European contact. American ecolo-


gists have been slow to embrace the archaeologists’ and anthropologists’ burgeoning evidence that preindustrial cultures significantly alter their landscapes. But the mounting literature (see Krech 1999; Butler 2000; Kay and Simmons 2002; Mann 2002) surely cannot be ignored any longer. Various authors show evidence of resource depression, either through exploitation or competition. The Yellowstone evidence makes it clear that the ecosystems of the region cannot be fully understood without knowledge of the prehistoric influences and protection therefrom. Without that understanding, insightful comparisons cannot be made with outside areas subject to modern impacts. That prehistoric humans were a component of a system at some other time is relevant to understanding it today.

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The First “Experiment”: 1878–1923

4 [E]very single organic being may be said to be striving to the utmost to increase in number, that each lives by a struggle at some point of its life, that heavy destruction inevitably falls either on the young or the old. . . . Lighten any check, mitigate the destruction ever so little and the number of the species will almost simultaneously increase to any amount. —Charles Darwin

HISTORY REVISED For nearly a century, the prevailing view among Yellowstone National Park officials on the northern herd’s population trajectory, following its recovery from about 1883 after the market-hunting period, was rapid increase to high numbers around the turn of the century followed by some decline in the early decades of the 1900s. This was based on early reports by administrators and their efforts at census. From the 1920s to the 1960s, a procession of park biologists concurred in this view. In 1974, 3 years after proposing his natural-regulation hypothesis and 7 years after the park announced the natural-regulation policy, Houston (1974) rejected the earlier view and concluded that the herd had not risen to the high levels that previous investigators had maintained. Along with the debate over elk numbers in prehistory, this disagreement over the numbers to which the herd rose continues to be one of the contentious questions in the northern range issue.

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Houston (1974, 1982) extracted comments on animal numbers, condition, and distribution from park superintendents’ annual reports and from scout and ranger diaries and log books, during the years from park establishment to 1920. These vary in detail and subject matter, and hence interpretation is provisional and judgmental. In general, as Houston remarks, the reports comment on increasing numbers of elk in the late 1800s and early 1900s. By 1879, after he had appealed to market hunters in 1877 to stop shooting elk in the park, Superintendent Norris (1879) commented that “choice animals” had increased. By 1886, Superintendent Harris (1886) reported “that there is more game in the Park now of every kind than was ever known before.” The following year, several thousand elk were estimated to winter in the Lamar River valley and its tributaries (Harris 1887). Predator poisoning began in the 1870s (Norris 1880), and by 1881 Norris (1881) spoke of the “easy slaughter” of wolves and their approaching extermination. Superintendents’ reports continued to recount elk increases through the 1880s, with the first estimates of total park summering populations at 25,000 in the early 1890s (Houston 1982:214). Reported numbers continued to increase with one account (1890) placing the total at 35,000–60,000. Estimates for the entire park continued in the range of 35,000–40,000 through 1910. The first reported estimates of the northern herd apparently appeared in 1912. Shore commented on 15,000 in the “upper Yellowstone Valley area,” and Superintendent Brett reported 30,101 “along the northern border . . . an approximate accurate census” (Houston 1982:219). Superintendents’ northern herd estimates continued in the range of 20,000–35,000 through 1918 or 1919, then dropped back to 11,000–13,500 by 1919–20 bringing the numbers down to the range of the first National Park Service (NPS) censuses in the 1920s (refer to table 2.2). The decline was generally attributed to severe winter losses in the 1917–20 period plus large hunting kills—7,000 in 1918, 8,000 in 1919, 3,206 in 1920—on animals migrating out of the park. Houston (1982) rejected most of these accounts on a number of grounds. Some of the reports did not clarify whether they referred to the summer population of the entire park or to animals wintering on the northern range. There was confusion at times over which of the several northern wintering herds were included in what today is considered the northern herd. Some figures were manufactured. Some estimates were based simply on a percentage increment over the previous year’s numbers. Field procedures varied; some were faulty. Special short-term investigations by the U.S. Forest Service and Bureau of Biological Survey personnel in 1915–19 concluded that the counts were high, perhaps by a factor of two. Houston (1974) also tabulated annual numbers of elk observed by scouts and rangers on horseback and ski patrols to get some impression of whether any trend was indicated. He perceived none and inferred that this also pointed to the absence of any sharp rise in numbers by the early 1900s. He concluded that following the market hunting of the 1870s, the northern herd recovered and stabilized in a range of 12,000–15,000. Probably not

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coincidentally, this was also the range in which his second prediction forecast the herd’s equilibrium point following the low point of the 1960s. He did not think the evidence warranted the contention of heavy winter losses in 1917–20.

HISTORY RESTORED Houston’s (1974) evaluation of the early reports is persuasive and appropriately critical. Clearly there were problems with interpreting and trusting them. But it argues only on the negative side of the question and does not equally explore whether there are reasons that might suggest the approximate validity of the reports. He does not offer a biological rationale for why the herd should have equilibrated at 12,000–15,000—especially in view of the fact that he was predicting 6,000–9,000 only 3 years earlier—nor does he present any argument why 20,000–35,000 was biologically implausible. I find five lines of evidence and reasoning that collectively build a case for the reality of a wintering northern herd around the turn of the century and the early 1900s on the order of 20,000–35,000 or more. Not all are individually strong bases for drawing conclusions. But they are consistent and in the aggregate provide a basis for what I consider to be the most probable inference. The first is the consistency of the reports. The range of 20,000–35,000 was reported for nearly 30 years. That it should fall in this magnitude by pure serendipity rather than, say, 10,000 or 40,000 asks a considerable amount of chance. I have difficulty accepting the view that these reports were all fictional and without basis. As was discussed in chapter 2, contemporary organized censuses, whether ground or aerial counts, tend to underestimate actual numbers by roughly a fourth. If there were any tendencies on the part of the superintendents to exaggerate their counts, those tendencies would have been offset to some degree by what probably were less organized and systematic censuses that likely undercounted to a greater degree than today’s counts. Second, during summer 1914, a task force of two national-forest supervisors and the chief field naturalist of the U.S. Bureau of Biological Survey examined elk range in Wyoming, Montana, and Idaho surrounding YNP (Smith et al. 1915). The task force reviewed YNP population estimates for the northern range and Forest Service estimates for the southern herd (those wintering in Jackson Hole), clearly distinguishing the two. As Houston points out, they were critical of the park counts. But he did not point out that they considered the April 1914 counts to be “the last reliable estimates” of the “northern herd.” They subdivided the numbers into 14 subpopulations of which the Yellowstone and Lamar River Valley population—the heart of what we refer to today as the northern herd— numbered 27,800. At that point the various subpopulations wintering in the northern portion of the park were clearly recognized. In his 1918 annual report, Superintendent Lindsley commented “while no accurate count was made of the herds of elk during the past year, more than 20,000 were seen in the park in the month of January with no special effort having


been made to count them” (Houston 1982:228). By 1919, the northern herd size had apparently begun to decline, but even in that year Graves and Nelson (1919)—a Forest Service investigator and the chief of the Bureau of Biological Survey, respectively—reported a 1917 census by personnel of these two agencies plus employees of the state of Montana at 19,000 animals in “the northern group.” Third, part of Houston’s (1974) negation was based on his inability to see any increase trend in the scouts’ and rangers’ diaries and logs. Indeed the numbers are highly variable and one questions whether any inference, trend or no trend, should be drawn from them. But if no trend is inferred, it does not bear on the question at hand. The earliest chronological sequences that Houston compiled began in 1898. Half began in the early 1900s. By these dates the population had undoubtedly already reached the magnitude in question. As already reported, the large numbers for the park were being reported by the latter 1890s. The herd may well have been approaching or at rough equilibrium by the turn of the century and the first two decades of the 1900s, just as the contemporary herd equilibrated at some level above 20,000 during the 1980s and 1990s (refer to table 2.1). At rough equilibrium, no trend should have been shown by the diaries and log books. Fourth, three consecutive, large hunting removals of 7,000, 8,000, and 3,206 were reported for 1918, 1919, and 1920 (Houston 1982:230, 234, 235). Nonlocal hunters shipped their animals by rail from the railroad terminus in the gateway community of Gardiner, Montana, at the park’s northern entrance. Shipping records, interviews of local hunters, and surveys of crippling losses provided a reasonable record of the hunting kills. These numbers provide a basis for approximating the herd size by a method that does not depend in any way on the censuses or population estimates. Houston (1974:15) questioned the validity of these kill estimates, considering “the most acceptable” values to be on the order of 3,500. I will accept his conservative estimates for the purpose of approximating the population size and assume removal of 3,500 each in 1918, 1919, and 1920. Population size can be approximated by considering what percentage these removals were of the population, and therefore what they imply about total numbers. The data in table 2.2 show that in all but 1 year in the 1920s and early 1930s, when there were censuses, the hunting removals fell well below 10%, and in all but 1 year were below 5%. The mean for the 1920s was 3.7%, for the 1930s was 6.9%. If I assume 5% for the 3,500-animal harvests in each of the 3 years, the implied population in each of the years is 70,000. It would be reasonable to assume a higher rate of kill, perhaps 10%. As discussed in chapter 2, the proportion of the population moving out of the park and becoming vulnerable to hunting is higher at high populations and there may have been a stronger vestige of the earlier emigration tendencies than in recent decades. If the kills in each of the 3 years were 10%, the implied population size is 35,000; if 20%, a level reached only once in the 1922–37 period, the implied population is 17,500.

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Thus, I have now bracketed the population size around the numbers claimed by the early reports. If one accepts Houston’s surmise that the reported kills overstated the true values by a factor of 2, and if one accepts reasonable values for the proportions of the population shot based on the measured percentages of following decades, the implied 1918–20 northern herd size falls well within the range of 20,000– 35,000, reported by park officials for some 30 years. These numbers, based on projections from the kill estimates, do not depend in any way on the disputed park counts. Given the conservative nature of all values used in these projections, it is possible that the range of 20,000 to 35,000 may well underestimate the true population size of that period. Fifth and finally, as discussed in chapter 2, the range of 20,000–25,000 is the level to which the herd returned in the 1980s–90s following cessation of park control. And it did so on a forage base that was heavily impacted during the first century of park history, as I will discuss in subsequent chapters. If the herd could rise to these levels in the 1980s–90s, there is no reason to suspect that it could not have done so at the turn of the century. In fairness, this cannot be used as a critique of Houston’s hypothesis, which was posed before the herd rose to its recent numbers. But it is another basis, again independent of the disputed, early park counts, for the collective lines of evidence and reasoning which point to the validity of the pre-Houston scenario on the 1878–23 herd trajectory.

DISCUSSION The predictions of Houston’s natural-regulation hypothesis that the herd would stabilize at some “moderate” level—first 6,000–9,000 in 1971, then 12,000– 15,000 in 1974—following the low point in 1968 was consistent with the other tenets of his hypothesis: that equilibration would occur without significant impact on the ecosystem. But the earlier park views of a herd increasing to 20,000– 35,000 the first time it was freed of human and other constraints early in park history raised questions about the plausibility of the new predictions. Some reconciliation was needed between the new hypothesis and the early accounts of an elk population burgeoning to extremely high numbers. That was achieved by rejection of the early accounts. Moreover, as discussed in the last chapter, the earlier park views of low elk densities in the region in prehistory were also inconsistent with an hypothesized equilibrium level, or carrying capacity, or K value of approximately 12,000–15,000, characteristic of what is now the northern range . Hence the efforts of park publications since 1971, described in the last chapter, to make a case for significant elk numbers in prehistory. There was no empirical or theoretical basis in ecology for hypothesizing in 1971 that a North American elk herd would equilibrate short of significantly impacting its forage base to the point of incurring nutritional shortages and reduced population process. The contrary view prevailed for North American ungulates at the time. Leopold et al. (1947) had published a classic paper that


played a significant role in setting a frame of reference for large-ungulate management of that era. The authors concluded that a combination of predator control and protection from hunting allowed deer populations to irrupt to the point of significantly impacting the vegetation and cited numerous examples across the country. In the 1950s and 1960s, state departments of natural resources nationwide changed from bucks-only to antlerless deer hunting laws in efforts to prevent permanent destruction of deer ranges, often over the protests of hunters who wanted continuing high deer populations and hunting success. Leopold himself entered the fray during the 1940s, arguing for liberal deer-hunting regulations in his home state of Wisconsin before his death in 1948. By the 1960s, Canadian moose and wolf specialist Douglas Pimlott (1967) was generalizing that intrinsic population constraints on North American ungulates that would prevent them from increasing to the point of damaging their food base might not have evolved because the prevalence of heavy predation obviated the need for such adaptations. Barmore (1980:441) cited a number of authors in the 1970s who concurred in the absence of such intrinsic self-limiting mechanisms. By the 1970s Caughley (1976) was proposing new theory of ungulate population behavior . But his first paper that caught attention of ungulate biologists (Caughley 1970) was a largely empirical study summarizing his observations on introduced species in New Zealand. His first substantive theoretical contribution did not appear until 1976, 5 years after proposal of the natural-regulation hypothesis and 2 years after Houston (1974) had completed his population analysis and revised portrayal of the northern herd population trajectory. Caughley generalized that after an ungulate species and vegetation first came together—reflecting his experience with the introduction of exotic species in New Zealand, and of uncertain application to North American systems in which ungulates and vegetation have coexisted for millennia—they would oscillate over a period of time and eventually come to some joint equilibrium in the absence of human intervention. He called this point the ecological carrying capacity. It included extreme reduction of plant biomass. Although Wagner et al. (1995a:117–25) commented on the lack of a strong empirical basis for Caughley’s theory, his model and the appropriateness of ecological carrying capacity for national parks are often cited in YNP publications. The important point here is that there was no strong theoretical climate or body of evidence in 1971 for the natural-regulation hypothesis that an elk population would somehow limit its numbers at some moderate level without significantly impacting its ecosystem. It was posed at the end of a half century of park research and a procession of biologists who had uniformly recounted high elk numbers and significant alteration of the northern range ecosystem. In 1879, major constraints had been removed from the northern herd. Wolf control had been under way for several years, and Native American and Euro-American hunting inside the park had been stopped. The herd was increasing with a lush forage base that had not previously been subjected to heavy herbivory.

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I know of no evidence suggesting any other constraints on the herd except for the density-dependent intraspecific competition for the forage as the herd reached high levels and interaction with weather factors (Merrill and Boyce 1991; Coughenour and Singer 1996) that were discussed in chapter 2. There is obvious evidence of density dependence in the contemporary population. But that density dependence was not sufficient to stop population growth between 1968 and 1988 short of 20,000–25,000. I see no reason why it should have been any different in the late 1800s and early 1900s when market hunting had stopped, Native Americans had been removed, wolves were controlled, and park protection became a reality. The herd had 36 years between abatement of market hunting in 1879 (Norris 1879) and the Smith et al. (1915) report of 27,800 in the northern range in 1914. Increase from the low point in 1968 to 20,000–25,000 in 1988 occurred in 21 years. Thus the herd had nearly twice as much time for the first increase. Houston (1982) reported that some poaching continued for several years after park formation, and the superintendents commented that some animals were shot during the 1880s to provide meat for the park hotel dining rooms. But as discussed, the increase in the second half of the 1900s took place in the face of substantial hunting kills on portions of the herd leaving the park in fall and winter. Why the northern herd declined by the early 1920s is less certain than the 1880–1914 events. Houston (1982) rejected the early contentions of heavy winter losses, yet in his progress report (1974:76–77) he listed 10 winters between 1880–1962 when the number of dead animals counted ranged from 476 to 1,888. More recent, heavy winter losses have been well documented. Lemke et al. (1998) calculated the 1988–89 winter loss at somewhere between 3,021 and 5,757. Coughenour and Singer (1996) estimated the loss in this same period at 6,718, 30% of the herd. Lemke (personal communication, February 12, 2003) suspects a winter loss of 4,000–5,000 in 1996–97. As inferred from figure 2.1, at any given population size, the difference in r between the r-value on the regression line and the r = 0 line indicates the amount of additional mortality and/or reproductive suppression the herd can absorb without declining (r turning negative), on average. Clearly this resilience declines as the herd increases. Thus a given proportionate winter loss is more likely to reduce the population at high numbers than at low. A small proportionate loss is more likely to induce decline in a large population than in a small one. Moreover, as already discussed, the fraction of the population shot outside the park rose as the population increased from 1968 to the 1990s, apparently because an increasing fraction migrates out of the park in winter and is vulnerable to hunting as the population rises. This may well have happened in the early 1900s, as suggested by the large hunting kills of 1918, 1919, and 1920. Thus a large population of the early 1900s, with its rates of change reduced on average to near zero by density dependence, could well have been reduced by half by a sequence of 2 or 3 years with somewhat above-average winter kills and large hunting kills. As already stated, Coughenour and Singer’s (1996) es-


timated 6,718 winter loss during 1988–89 alone reduced the population by 30% between early winter and April. In sum, the evidence is persuasive that the northern herd has twice been freed of major constraints, particularly human exploitation. It has twice risen to the range of 20,000–35,000 animals. I will present evidence in subsequent chapters of what effects those increases have had on the northern range ecosystem.

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The 132-Year Population Trajectory and Associated Synthesis Design

5 But biology can advance here only when in this fighting the better argued claim, based on better evaluation of evidence, is being selected. —Holmes Rolston III

THE POPULATION TRAJECTORY The purpose now is to reconstruct the northern herd population trajectory for the 132-year period since park establishment and before from the data presented in chapters 2–4. The reconstruction is shown in figure 5.1. I placed the prepark population size on the order of 5,000–6,000 according to the arguments in chapter 3. The effects of market hunting at and immediately following park establishment and the 5,000 count in 1883 are also discussed there. I have accepted a general population increase from 1883 to the 1890s and early 1900s to the range of 20,000–35,000 with the 27,800 in the Yellowstone and Lamar River valleys in 1914 reported by Smith et al. (1915) as “the last reliable estimate” (figure 5.1). I have also shown the Graves and Nelson (1919) count of 19,000 for 1917. As discussed in the last chapter, these may be conservative. The counts and estimated values for 1923–68 are the winter censuses and the estimated fall populations that appeared in table 2.2. The same values for 1969–2002 are the winter censuses from table 2.1, each divided by 0.75 for the sightability correction, and then incremented by number of animals taken in the fall and the numbers taken in the late hunt prior to the censuses.

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Figure 5.1 Reconstructed times series for the Yellowstone northern range wintering elk population from before park establishment in 1872 to 2002. Black dots are actual counts: those reported in the early years (see chapters 3 and 4) and park winter censuses for 1923–2002 (see tables 2.1 and 2.2). Open circles are total population estimates (1) as calculated for the 1800s in chapter 3, and (2) fall population estimates for the 1900s shown in tables 2.1 and 2.2. The dashed line is the surmised trend between counts and estimates until the first NPS census in 1923.

The entire time series begins with a northern wintering herd prior to park establishment in the general range of 5,000–6,000, largely migrating out of what is now the park to lower elevation winter range. Following reduction to lower levels during the brief market-hunting period at and immediately following park establishment, the herd increased to a range of 20,000–35,000 in the 1890s and early 1900s. It declined in the second decade of the 1900s to censused numbers in the 1920s through 1950s, largely between 7,000–12,000, and estimated fall populations largely in the range of 11,000–18,000. The low density at park establishment, reduction during the market-hunting period, subsequent rise to the range of 20,000–35,000, decline in the twentieth century to the censused numbers beginning in 1923 were the pattern claimed by park officials and researchers for roughly a century until Houston’s (1974, 1982) revision. That revision has been maintained by park personnel up to the present (Despain et al. 1986; Bishop et al. 1997). I have returned to the earlier view because I think the evidence, reviewed in the last three chapters, supports that scenario. There is of course no certainty in reconstructing the past. But science never gives certainty. My conclusion here is my judgment—based on the evidence, logic, and the ecological record and theory—on where the highest probability lies.

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FORCES DRIVING THE POPULATION TREND The northern herd has evidently undergone some short-term (i.e., year-to-year) fluctuations. Winter mortality has been implicated in a number of cases: alleged for the 1918–20 decline, as discussed for the winter of 1988–89 (Coughenour and Singer 1996), and 1996–97 (Thomas Lemke, personal communication, February 22, 1999). The 1918–20 decline occurred in three years with large hunting removals, as discussed in chapter 4. The 1988–89 decline occurred after a drought year. All of these declines occurred at times when the population had risen to high levels, was sustaining heavy density-dependent constraint (refer to figure 2.1), and had limited demographic resilience for coping with environmental pressures. Summer mortality has received less recognition but may be more influential than winter loss. In their excellent paper, Coughenour and Singer (1996) reported mean summer calf mortality to be 0.52 and mean winter loss 0.33. Thus summer calf mortality rate averages 1.7× winter mortality. Both are inversely correlated with water-year precipitation of the previous year. Evidently the nutritional condition of the calf, determined by the forage conditions of the summer, influences its vulnerability to winter conditions. As discussed in chapter 1, these authors and Merrill and Boyce (1991) observed correlations between annual rates of population change and annual precipitation or its surrogate, vegetative production. However, these weather-induced fluctuations have only been variations on the general trends in the population’s numerical trajectory over the past 132 years (figure 5.1). The sweep of that trajectory has been significantly a function of the extent of human constraint: 1. Although it cannot be known with certainty, I find the evidence and logic persuasive in Kay’s (1990, 1994a) hypothesis that the low level of ungulate numbers in the Yellowstone region prior to park establishment was maintained in particular by aboriginal hunting and to some degree by mammalian carnivores. A substantial amount of evidence and literature are in concert with this view. 2. Increase to the population high at and after the turn of the century was made possible by release from the pre-Columbian population constraints afforded by park establishment and protection provided by park personnel. 3. Decline to the 1968 low point was induced by increasingly intensive sport hunting outside the park and agency control efforts. 4. The rise from 1968 to recent numbers approaching those of the turn of the twentieth century was made possible by cessation of park control efforts, park protection once again, and relatively light hunting removals (refer to table 2.1) outside the park. In the absence of major human constraints, the population appears to have been limited in the latter 1980s and early 1990s by a combination of densitydependent forces—intraspecific competition for forage (Coughenour and Singer


1996), limited outside hunting, and some predation effect (Singer et al. 1997b)— and randomly fluctuating weather influences to an actual fall population level of approximately 20,000–25,000. Thus the equilibrium level was set largely by nonhuman influences, but increase to that level was once again permitted by park protection. It would doubtless have been higher without the limited outside hunting kills that occurred during the period. Given low populations prior to park establishment, it follows that elk numbers in the late 1800s and through most of the twentieth century have been well beyond the bounds of historic variability. And the northern range ecosystem has experienced levels of herbivorous influence to which it had not previously been exposed, at least perhaps since the last glacial retreat, human arrival, and megafaunal extinction. I will examine the nature of those influences in the chapters of part III.

A NEW PHASE? As discussed in chapter 1, the herd and range may be entering a new phase in the first decade of the twenty-first century by virtue of two changes under way. The first is an increasing exodus of animals out of the park in winter. Thomas Lemke (personal communication, February 22, 1999) reports that major fractions of the herd have moved northward out of the park each winter to the region of Dome Mountain (refer to figure 1.1). This pattern began in the winter of 1988–89, and has averaged a third of the herd over the succeeding winters. In some winters (e.g., 1988–89, 1998–99) the movement reaches or exceeds 60% of the herd. The result must be some easing of herbivorous pressure on the park portion of the range. The wolf introduction in 1995 also may be making inroads on northern herd numbers, as discussed previously. Thus a combination of these two changes could reduce the number of wintering elk inside the park’s northern range to another phase of lower numbers. A well-designed vegetation monitoring effort seems highly desirable.

SYNTHESIS DESIGN In an imaginary world, with unlimited resources and freedom of action, the managers of Yellowstone National Park could have evaluated the effects of a large elk population on its ecosystem with a grand experiment. Replicated experimental treatments with different densities of elk could have been maintained, associated hypotheses posed, and the effects on the system assessed with a menu of measurements specified at the outset. Replicated controls with no animals and parallel measurements could have matched the treatments. Despain et al. (1986) called natural-regulation management an experiment. But Kay (1987) and Wagner et al. (1995a:152) questioned the use of the term on the grounds that there were no replications, no controls, no stated hypothesis

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(although perhaps Houston’s [1971] natural-regulation hypothesis could have been invoked to serve in this role), and no regimen of measurements specified a priori that would be continued until a stated time when the results of the experiment would be declared. Bishop et al. (1997) later agreed that the term was not appropriate in this situation. However, the synthesis that follows bears some characteristics of an experiment that tests the hypothesis of significant ecosystem effects by large numbers of elk on the northern range ecosystem. Three independent measurement protocols allow comparison of system response between situations of heavy elk use and those with light-to-moderate or no use, and permit inferences of cause and effect. The first protocol involves the changing elk numbers shown in figure 5.1. The population trend constitutes two periods of high numbers that emulate two experimental replicates and two periods of low density that serve as two quasicontrol replicates. That these four treatments have alternated provides a de facto switch-back design, and any alternating responses will strengthen an inference of causation over mere one-way correlation. I have arbitrarily designated the periods during which the early estimates and later censuses consistently fell below 6,000 as the low-density or control periods. These are the periods from prehistory through 1884, and 11 years from 1958–59 through 1969–70 (figure 5.2). The high-density or experimental periods then become 1885 through 1958, and 1971 to the present, 73 and 33 years, respectively. However, I am not taking these as absolute limits on the response periods for several reasons. Different components of the ecosystem may respond differently to a given herd size. I will also present evidence in later chapters of what appear to be lagged responses, some appearing in the evidence only some time after the beginning of a treatment; others continue a trend begun in one treatment and extend for some time into a new one. A related question is the length of time required for the effects of a given population treatment to develop and become measurable in the ecosystem, if at all. Yet another question is the state of the system at the time a given number of animals began using it. A dense and diverse vegetation that had not experienced heavy herbivorous pressure since the last glacial retreat might have been able to absorb the pressure of 20,000–35,000 elk for some years or even decades before their effects became obvious in the early 1900s. The earliest report I have found of grazing effects was 1914 (Smith et al. 1915) although I place the beginning of the first high-population treatment in 1885 (figure 5.2). By the time of the second population high after 1970, the system had experienced the heavy use of the first high for 73 years. Thus, although the population rose to roughly comparable levels twice and constituted two replicate treatments, each treatment was applied to a different system. Consequently the two high-population phases were not in fact two well-replicated experimental treatments. Similarly, park personnel state that there was no response of the northern range during the population reduction of the 1960s and 1970s (see Singer et al. 1994). I will present evidence in later chapters that there were responses, but it


Figure 5.2 Synthesis design showing major evidence sources and chronology of control and experimental periods based on changes in elk numbers.

might not have been surprising if there had been none. By 1959, the northern range had been subjected to the first high-population treatment for 73 years. The system could have been so depleted that an elk population reduced below 6,000 could still have been sufficient to prevent recovery. Moreover, the lowdensity period in figure 5.2 only spanned 11 years (1959–70). That might not have been sufficient time for a depleted vegetation, still being fed on by a reduced herd, to show significant recovery in this semiarid environment. The inability to control variables other than herd size is another departure from a formal experiment. Several environmental factors were varying during the 132 years of park history. A number of these have been invoked as causes of ecosystem changes that have occurred, in some cases discounting elk effects simultaneously. Recourse to specific aspects of the evidence is needed to test these hypotheses as well as that of elk effect. The spatial distribution of elk, and consequently their use, is not uniform across the northern range. The animals tend to move west through the winter season, and concentrate toward the west end. As I will discuss in later chapters, evidence of effect varies across this herbivorous gradient. The retrospective nature of this synthesis is another departure. It largely relies on measurements and evidence sources not specified a priori. The nearest approach was a set of exclosures established in 1957 and 1962 at the time of

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intensified herd reduction and periodic inside and outside vegetation measurements made over the following 30 years or so. But numerous other studies have spanned only 1 or 2 years. Thus it is necessary to depend on an extensive and diverse mass of evidence for the 132-year period. There is evidence for all four stages of the population trajectory, but it is of different forms, and no form spans the entire period, except perhaps photography. In many cases, the coincidence of a given form of evidence and a given stage in the population trajectory is fortuitous. Several problems arise with this use of discontinuous and diverse forms of evidence. One is interpreting causation in cases where there is evidence for change in the state of the system only between the time of park establishment and the present. Because there were two intervening stages of high elk populations, assigning time and causation of change is problematic. An analogous problem with data discontinuity is that system responses developing in one population stage may be obliterated by a subsequent stage. If measurements were not made during the first response, there may not be any evidence of it by the time the following stage is well under way. For example, any changes occurring during the reduction period of the 1960s may have been eliminated by the population growth of the 1970s, 1980s and 1990s and therefore undetectable by current measurements. Another potential problem is comparison of system response during two or more population phases, when the responses are measured with different methodologies (figure 5.2), and the forms of evidence are different. Comparison must be made with care, is often qualitative, and is rarely subject to statistical testing. Some authors have criticized the use of photographs, asserting that a photograph only provides evidence for a point in time. I do not consider this a valid criticism of photographs that show perennial vegetation, especially where the ages of the plants can be approximated. The presence of a plant in an area attests to the fact that the environment was sufficiently suitable at the time of its establishment to permit it to begin life, and through its lifetime to permit its survival. Plants of a number of ages in a photograph provide evidence of environments over a period of time. Moreover, repeat photographs of the same sites do provide a basis for judging change (Kay 1990; Meagher and Houston 1998) if interpreted with care and objectivity. In a number of cases, it has been necessary to use anecdotal information. Using such information is always risky unless it can be placed in some context within which its validity can be judged. That context may be a knowledge of the ecological situation about which the comments are made. Or it may be the observations of knowledgeable scientists or other experts whose experiences and familiarity with scientific procedures, and whose understanding of the significance of their comments, can be expected to temper them. In the ideal, anecdotal information should be quantified. Chapter 3 presents examples of how quantification through temporal scaling, as originally devised by Kay (1990), and the use of population parameters give quite different indications of animal numbers than subjective appraisal of the same anecdotal data sets.


Finally, hypotheses of no effects and significant effects are hypotheses of dynamics. Inferences about dynamics are problematic in the case of single measurements or observations. But with two or more measurements or observations in time, change may be evident and dynamics may be inferred. As I will discuss in later chapters, virtually every subsystem of the northern range ecosystem has been in continuous change through park history. A system of exclosures inside the Park provides for a second protocol of quasiexperimental measurements. Singer et al. (2003) inferred from previous references to conditions inside park exclosures by Kay and Wagner (1994) and Wagner et al. (1995a) that “Some hold the view that ungulates should have no effect on plant species composition . . . totally ungrazed sites inside exclosures were used . . . for . . . evidence of too much ungulate herbivory” (emphases added). The authors neither stated nor intended such value judgments. The conditions inside the exclosures were cited in response to frequent park statements discounting grazing effects on the northern range, a scientific and not a value question. Those conditions are used in the same context in this book: benchmarks at one end of a grazing-intensity continuum along which ecosystem effects can be examined objectively. Inside-outside measurements provide the basis for inferences of effect. When these are repeated over time, they provide evidence of dynamics. System measurements inside and outside park boundaries provide a third protocol for inferring elk effects. Obviously these make for imperfect comparisons. The inside and outside systems differ in ways other than elk use and cannot be controlled. There are resident elk outside the park, and some northern range elk winter outside YNP. But there is no question that far greater numbers of elk winter inside the park portion of the northern range: several fold prior to the winter of 1988–89, and 2 to 3 times as many, on average, since that winter. Thus there are 3 bases for evaluating the effects of herbivory by a large elk population on the northern range ecosystem. Except for the exclosure measurements, none of these was based on a research design established to address the questions of elk effects. Most of the evidence is circumstantial. But taken together and combined with the relevant literature and theory, it provides what I infer to be a consistent and coherent paradigm, set forth in the chapters that follow, about the northern herd effects on the system.

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Elk Effects on Ecosystem Structure and Function

III The key bridges [to understanding] the paradigms of how nature operates are . . . provided by views that include a systems perspective, interdisciplinary, nonlinearity, and cross-scale views. —Lance H. Gunderson, C. S. Holling, and Stephen S. Light


Influences on Upland System Structure I: Aspen Woodland

6 One of the penalties of an ecological education is that one lives alone in a world of wounds. Much of the damage inflicted on land is quite invisible to laymen. —Aldo Leopold

ASPEN BIOLOGY AND NORTHERN RANGE ELK EFFECTS Aspen in western North America grows in clones, and what may appear as a “stand” or “grove” of trees is in actuality a cluster or patch of stems, or ramets, growing from a single root mass: thus a single organism (figure 6.1). Kay (1998) commented that a recent random sample of clones in the northern range averaged more than 0.4 ha (1 acre) in area. Some authors claim that there are no authentic records of clones being established in the West through sexual reproduction during recorded history (Mitten and Grant 1996; Kay 1998). However, there are some reported cases (Peterson and Peterson 1992, Romme et al. 1997. Given the rarity of sexual reproduction in the West, the species occurs as an extensive, regional population of clones largely occupying the intermediate elevations of the western U.S. and Canadian mountain ranges but also occurring on the lowlands across the Prairie Provinces and in Alaska. In much of the literature, western clones have been considered to be extremely old, commonly dating back to the Pleistocene (or even the Pliocene or Miocene, see Schier 1975), and becoming established at a time when the climate was more favorable for seedling survival (Mitten and Grant 1996; Kay 1998). But recent genetic research 59

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ELK EFFECTS ON ECOSYSTEM STRUCTURE AND FUNCTION

Figure 6.1 Aspen clone in Eagle Creek area ~5 km outside of the park’s northern entrance. Aspen branches near the ground, diversity of age classes (including young ramets on the fringes of the clone), and absence of bark scarring are characteristic of relatively unbrowsed aspen throughout the West. White spot in background is thermal terraces above the park’s Mammoth headquarters. Photo by Charles E. Kay, June 6, 1983.

is showing clones that were established more recently, presumably from sexual reproduction. The trees are short-lived, typically attaining 100–130 years longevity (Krebill 1972; Schier 1975; Jones and Schier 1985; Mueggler 1989). They senesce and die from a number of causes and are replaced by new cohorts of root suckers that mature and grow to tree stature. Thus over several centuries or millennia, the clones generate a succession of tree generations. Moreover, vigorous clones expand by proliferating root suckers on their outer fringes (figure 6.1), so typically the ramets in a given clone exist over a wide range of age classes. A clone can survive for a time without trees, supported by the photosynthesis of short (e.g., up to 1 m) suckers in a growth form called “shrub aspen” (figure 6.2). This clonal growth form has produced much of the ambiguity in discourse on aspen. When an author speaks of a decline in aspen, unless specified it is not clear whether the reference is to decline in area occupied by trees, in the number of clones, in the total area occupied by clones, or some combination of these. When aspen researchers study the ages of aspen “stands” by coring trees (see Mueggler 1989, 1994), they are in fact studying the ages of the older ramets in the current tree generation and not the ages of the clones.


Figure 6.2 Shrub-aspen clone in Yancey’s Hole in the northern range. Photo by Charles E. Kay, July 18, 1986.

Despain (1990; Renkin and Despain 1996) has hypothesized that shrub aspen is the “normal” growth form of the species, an adaptation for clone survival in the face of presumed centuries and perhaps millennia of heavy herbivory. The ecophysiological question is whether the shrub form, subject to chronic tissue removal, has sufficient photosynthetic surface to support indefinitely the respiration of clones with their extensive subsurface root masses and provide energy for annual growth attempts of the new ramets. There is evidence that shrub-aspen clones are less vigorous than those with trees (see Bailey et al. 1990). Kay and Wagner (1996) measured characteristics of regrowing stems in 1989 following the 1988 fires on the northern range in both shrub-aspen clones and those that had trees before the burn. Mean number of regrowing stems in shrub-aspen and tree clones were 19,170–1 ha and 120,941–1 ha (p < 0.001), respectively. Mean stem lengths at the end of the growing season and before elk browsing were 20.9 cm and 70.5 cm, respectively (p < 0.001). Mean widths of the largest leaf on each sucker were 26 mm (n = 159) and 79 mm (n = 127) for the two types (p < 0.001). These authors also observed that the shrubaspen ramets are short-lived. Stem ages on 22 shrub-aspen clones in 1986 before the fires were largely under 4 years of age, and none was older than 15. Renkin and Despain (1996) observed that clone vigor, as measured by sucker density and growth rate, is a function of root biomass. They suggest further that shrub-aspen is the characteristic growth form on marginal sites: “It is probable

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that shrub aspen is a consequence of the physical environment and reflects the inability of a particular site to support aspen as a canopy component.” But almost certainly, such clones have been subjected to decades of intense elk browsing. A combination of the stress of site marginality and a century of annual tissue removal by browsing may have so weakened the clones that they cannot generate significant tree growth. In a 21-year study of aspen in Rocky Mountain National Park, Olmsted (1997) traced the declining vigor of aspen clones subjected to annual browsing by a large elk population. Browsing rates of more than 30% of twig volume removed were correlated with declining stand density. Furthermore, “Aspen shoots can only be browsed a few times before diffusion of apical dominance makes it unlikely that the growing shoots will mature into a tree.” Renkin and Despain examined early photographs of aspen on sites that were subsequently invaded by conifers and were thought to have been obliterated. On examining the sites in the field, they found aspen still present in the shrub form. Identifying these in the early photographs could only have been possible if the clones were in tree form. Kay and Wagner (1996), as reported, identified clones in tree form in early photographs. At the time of their studies, two-thirds had been converted to shrub form. Thus the typical situation appears to have been clones in tree form in the early park years subsequently converted to and maintained in shrub form. These clones appear to be less vigorous than those with trees, are declining in area, or are disappearing. Elk affect aspen in several ways. They bite off portions of the bark in intermediate-age and young trees (figure 6.3). If the tree is girdled, it is killed. If not, the bite wounds provide entry for fungi that kill the trees (Krebill 1972; Hart 1986). The bark of the older trees may be too tough to bite through, but the gnawing produces black scarring around the trunks that is universally characteristic of older trees in the northern range (figure 6.4). Elk browse leaves and lower branches as high as they can reach, producing browse lines (figure 6.5). They browse out the ramets in winter down to the snowline, preventing growth to tree stature. With no replacement ramets, the older trees reach the end of their longevity, die out (figure 6.6), and leave the clone in shrub-aspen form (figure 6.2).

CHRONOLOGICAL CHANGES IN NORTHERN RANGE ABUNDANCE Decline in Area Occupied by Trees Probably the first quantitative estimate of the amount of aspen in the northern range was an unpublished one of M. P. Skinner’s cited by Barmore (1980:370) at 28,350 ha, made “probably . . . prior to 1922.” This area would be 41% of the 69,000 ha that Barmore considered the northern range, 28% of Houston’s (1982) 100,000 ha total area, or 34% of the latter author’s 82,600 ha in the park. Barmore considered Skinner’s estimate excessive, but Kittams (1948) cited Skin-


Figure 6.3 Aspen bark gnawing by elk in Eagle Creek outside park’s north entrance. August 11, 1986.

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Figure 6.4 Characteristic scarred appearance of northern range aspen. Note depauperate understory, absence of branches near the ground, and absence of young aspen ramets. Compare with figure 6.1. Photo by Charles E. Kay, August 3, 1986.

ner as saying “I would call the aspen abundant on the northern half of the Yellowstone.” Barmore gave more credence to a 1930 map (also cited by Houston, neither author giving the source) which showed the area of dominant and mixed aspen at 4,490 ha (Barmore’s sum of the two types, unexplainably, was 4,455). This would constitute 6.5% of Barmore’s 69,000 ha of northern range, approximately 4% of Houston’s 100,000, and around 5% of his 82,600 in the park. Barmore and Houston mapped aspen on the northern range in the 1960s and 1970s and considered the area to be 1,418 ha, or with some rounding error, approximately 2% of Barmore’s 69,000 ha and Houston’s 83,000. That a decline from 4,490 ha in 1930 to 1,418 in the 1960s or 1970s constitutes a 68% reduction, over that period. Houston (1982:92) generalized the area of aspen on the northern range during his study at 2–3%, that “in the original photos” at 4–6%, and hence about 50% decline. These values have become the standard park position (Despain et al. 1986:103–104, Bishop et al. 1997:51), although Renkin and Despain (1996) cited 6% and 2%, implying a 67% decline. More important, the larger figure is taken as the percentage of the northern range at the time of park establishment. “Aspen probably occupied between four and six percent of the range at the time the park was established in 1872” (Despain et al. 1986:103). An unpublished park document makes this same statement.


Figure 6.5 Aspen grove, photographed in 1932, showing initial stage of browsing highline. From Wright and Thompson (1935).

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Figure 6.6 Decadent aspen stand characteristic of those on the northern range. Trees are 100 to 130 years of age, mean longevity for the species, and are dying. No ramets (root sprouts) have escaped browsing to replace the trees. Photographed July 17, 1997.

If Barmore’s unsourced 1930 map is taken at face value, its implied 68% decline pertains only to the 30–40 years between it and Barmore’s and Houston’s studies. Roughly 40 years have elapsed since Barmore’s work and 29 since Houston’s (1976) vegetation report that was the basis for his book. There has most likely been further decline over this period, conceivably at the same or higher rate per unit time as occurred from 1930 to the 1960s. Moreover, there was almost certainly decline prior to the 1930 map. Elk populations by 1930 had been high for nearly a half century (refer to figure 5.1). As cited, concerns were being expressed as early as 1914 for the killing of “whole groves” by bark peeling (Smith et al. 1915). Skinner (n.d.), who began his studies in the 1920s, commented in an undated report: But before spring, . . . [other] food gave out, and the Wapiti attacked the aspens. . . . Many were pushed over and all bark, small branches and twigs were devoured. Most of the remaining aspens had their trunks so severely scraped of bark that they died. So much so, that the next summer I noted that all aspens on thousands of acres had been killed. Barmore (1980:370) commented, “Prior to about 1930, P. tremuloides in the northern part of the Park was apparently somewhat more widespread.” The time elapsed from the beginning of the first high-population period (1885, refer to figure 5.2) to 1930 was 45 years; from 1930 to the mid-1960s


during Barmore’s work was about 35 years; and from 1965 to 2004 is 40 years. If the rate of aspen decline in the first and third of these periods was roughly similar to that from 1930 to 1965, that is, 67% of aspen present at the start of each period, then the 1884–2004 decline would be: (1 – 0.333) × 100 = 96% I do not suggest that this is necessarily the correct amount. Obviously we do not know that the first and third rates were equal to the second. But it gives strong reason to conclude that the standard 50% park figure and the 67–68% suggested by Renkin and Despain (1996) and implied by Barmore’s data, all of which are based on 1930 to about 1965 numbers, substantially underestimate the 1872–2004 decline in area occupied by trees. By the same token, the amount of aspen on the winter range in 1872 must have been two to three times the amount in 1930, conceivably 8,000–12,000 ha and on the order of 10–15% of the park portion of the northern range. As I view the photographic record reproduced by Skinner (1927), Kittams (1948), Houston (1982), Kay (1990), and Meagher and Houston (1998) for the 1920s and before (see also figure 6.7), I am impressed with the number and extent of dense, robust aspen stands occurring in numerous areas of the northern range, unlike anything that can be found today except in exclosures. I have little doubt

Figure 6.7 An aspen stand behind a detachment of the Minnesota National Guard photographed in 1893 in the Little Blacktail area of the northern range. Trees range in age from small samplings with branches near the ground to mature trees decades in age and without bark scarring. Photo by F. Jay Haynes, courtesy Montana Historical Society, Helena, MT. This is figure 22 in Kay (1990).

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that the 6.5% implied by the unsourced 1930 map is conservative to a considerable degree in estimating the amount of 1872 aspen and that aspen was a more conspicuous feature of the northern range landscape than park spokespersons grant in their publications. Park publications (Singer 1996a; Bishop et al. 1997) reinforce the view of extremely small aspen acreage with aspen pollen evidence of Engstrom et al. (1991). These authors cored sediments of eight small lakes in the northern range and found aspen pollen extremely scarce, prompting Bishop et al. to generalize the Engstrom et al. results as showing aspen to be “a marginal species.” However, it is general knowledge among palynologists that aspen pollen is fragile and ephemeral. Engstrom et al. (1991) themselves describe some of these same difficulties with interpreting aspen pollen. The pollen data cannot tell much about the species’ abundance a century or more ago.

Chronology of Tree Establishment The Prepark Period Romme et al. (1995) hypothesized a scenario for northern range aspen growth from before park establishment to the present. The authors cored mature trees in 15 stands of the northern range, determined their ages, and backdated the trees to their establishment years. Their results and conclusions were: 1. The ages of the 77 trees (their figure 6) which they backdated placed the establishment years largely in the period 1870–90. 2. Because there were four trees as old as 154 to 167-years of age that placed their establishment in the 1820s–30s, they concluded that aspen has the potential to live this long. The near absence of trees of that age in the park must imply constraints on establishment prior to 1870. Those constraints, the authors surmised, were the high prepark elk numbers claimed by Houston (1982), which they accepted uncritically without considering the contrary Kay (1990) paradigm. 3. The authors also cited a study by Warren (1926) who aged 31 trees in the northern range in the early 1920s to develop the relationship between tree size (measured dbh) and age. All but two were established in the 1870–90 period. 4. The 1870–90 period coincided with the market-hunting elk reduction described in Chapter 3 along with other hypothesized factors suggested by western tree-ring chronology and Houston’s (1973) reconstruction of northern range fire frequencies. 5. More generally, they hypothesized that this period: was historically unique: numbers of elk and other browsers were low, climate was relatively wet, extensive fires had recently occurred, and large mammalian predators of elk (e.g., wolf, Canis lupus) were present. This combination of events has not occurred since 1900.


Park officials (see Bishop et al. 1997; Schullery 1997a; Huff and Varley 1999) promptly adopted the Romme et al. scenario as the official park position on the relationship between elk and aspen on the northern range: “aspen . . . owe their present abundance on the northern range to one demonstrably brief period of success in escaping browsing in the late 1800s . . . aspen are only an occasional occupant of the northern range” (Bishop et al. 1997:75). And it has been accepted by investigators outside of the NPS (“an ecologically unique period”; Coughenour and Singer 1996), and by freelance authors writing about the park (Baskin 1999). However, two lines of evidence falsify the pre-1870 part of the scenario. One is more in-depth analysis of the Romme et al. and Warren (1926) data. 1. Aspen in Yellowstone are extremely difficult to age because most of the older trees have heart rot. Kay (1990) attempted to core over 400 trees on the northern range, found it impossible in more than 90% of the trees, and was forced to abandon it. In reconstructing age distribution and establishment dates of trees outside the park boundary, he was obliged to use established regressions between tree diameter and age (Kay 1990:63). Romme et al. (1995) do not mention this in their work. When I asked coauthor Linda Wallace (personal communication, September 27, 1997) about it, she stated that the technician coring the trees commented on the difficulty of getting usable cores. According to Wallace’s recollection, something on the order of three-fourths were unusable. Thus, they were able to age only 77 out of something on the order of 300 trees. Because heart rot becomes more prevalent as the trees age, it is likely that the 77 trees did not constitute a random sample of the aspen age distribution on the northern range, with the older trees, in all likelihood, underrepresented or not represented at all. Consequently trees established before 1870 were almost certainly underrepresented in the Romme et al. sample. 2. The aspen literature attests to the short life span of aspen trees (see Jones and Schier 1985), most trees dying by 100 to 130 years of age. The relative absence of trees more than 120 years of age in the Romme et al. sample cannot be taken as evidence that large numbers of trees had not established prior to 1870. A large fraction of such trees would have died by the time these authors conducted their study. The authors indicate a knowledge of and comment on the short longevity of aspen trees but apparently did not consider it an important issue. The general tenor of their discussion indicates a belief that tree establishment was rare prior to 1870. 3. The presence of four trees established in the 1820s and 1830s cannot be taken as an indication of typical aspen longevity, as the authors suggest. Population survivorship curves characteristically have a few lowprobability, advanced-age individuals on the tails of the curves.

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Thus there is reason to question whether the 77 trees sampled validly reflect the pattern of tree establishment prior to 1870. Additional evidence indicates that there was significant establishment prior to that date. Romme et al. (1995) cite only a small portion of Warren’s (1926) data, the ages of the 31 trees the latter author used to relate diameter growth to age. In establishing this relationship, Warren almost certainly picked trees over a range of sizes. Hence, it cannot be assumed that his 31 trees were a random sample of the trees in his study area. Warren also measured the circumferences of 135 stumps of aspen trees cut by beaver. These measurements can be converted to tree ages by converting circumferences of the stumps to diameters at breast height (dbh) and converting these to ages with Warren’s relationship. I measured circumferences of 200 trees at 17 inches (mean height of Warren’s [1926] beaver cuts) and at breast height at 1,846 m elevation east of Logan, Utah. I converted these to diameters, regressed the 17-inch measurements on dbhs, and then converted Warren’s 135 stump circumferences to dbhs. I then converted the latter to tree ages with Warren’s table 1 and backdated these from his 1925 measurement date with the following tree establishment dates: Establishment Decade < 1850 1850–59 1860–69 1870–79 1880–89 1890–99 1900–09 1910–19

No. Trees 7 15 30 36 20 22 4 1

Thus the establishment dates are distributed over a flat curve from 1850 to 1899, with 39% established before 1870. If establishment of significant numbers of the contemporary trees are to be attributed to the 1870–90 elk reduction (e.g., Romme et al. 1995; Bishop et al. 1997; Schullery 1997a, b), consistency demands that the pre-1870s trees in Warren’s samples must also imply low elk numbers in the earlier 1800s. This is consistent with the chapter 3 inference of low elk abundance and winter migration out of what is now the northern range prior to 1872. It also falsifies the view that aspen trees were established on the northern range only as a brief episodic event in 1870–90. More recently, Ripple and Larsen (2000) aged aspen trees in the same sites studied by Warren and randomly over the northern range. They did comment on the frequency of heart rot and the difficulty of getting usable cores. In one sample of 20 trees aged in stands studied by Warren, 14 (70%) were established before 1871, 4 (20%) of them before 1801. In their random sample of 98 trees across the northern range, approximately 11% were established before 1871. Here again, the realities of tree longevity and inability to core the older trees would impair detection of trees established in the earlier decades. Nevertheless, they


concluded “that successful aspen overstory recruitment occurred on the northern range of YNP from the middle to late 1700s until the 1920s.” A second source of evidence on the chronology of aspen-tree establishment is the photographic record. Kay (1990) compiled early photographs of the northern range showing aspen vegetation. His figure 22, shown here as figure 6.7, depicts an entire hillside covered with aspen at Little Blacktail in 1893. Numerous trees in the background have boles with diameters equivalent to or greater than those of the heads of the men in the foreground. Warren’s diameter-age relationship places trees with diameter 5.82 in (14.8 cm) at 60 years of age. Thus there clearly were trees established in the early 1800s. Moreover, the stand shows a diversity of age classes down to very young saplings, implying the establishment over several decades of the 1800s up to the date of the photograph and the absence of heavy browsing pressure throughout that period. Kay’s figure 23 is an 1899 photograph showing several beaver-felled aspens. There are no humans in the photos, but several of the trees are obviously quite large, perhaps with diameters at least 6–9 in (15.2–22.9 cm) and almost certainly established in the early 1800s. An 1893 photo from the F. J. Haynes Collection (figure 6.3) shows a hillside with dense aspen growth, the trees in the left background obviously several times the height of the men in the picture. They were probably several decades in age, and again established well before the 1870–90 period. Meagher and Houston (1998) show early photos—plates 46.1 by Russell in 1929, 54.1 by U.S. Army Corps of Engineers in 1900, and 59.1 by Nelson in 1898—which have trees that could be decades old and predate 1870. I do not suggest that all aspen trees on the northern range were several decades in age at the end of the century. But the photographic record supplements the Warren evidence by indicating an aspen vegetation in the 1800s with significant areas of mature trees that could not be disclosed by the Romme et al. methodology. I conclude, in ending this section, that the reliable available evidence points to an aspen vegetation with a range of tree ages, and hence diverse establishment dates, in the northern range at and before park establishment. This is consistent with the view developed in chapter 3 of low elk densities in the 1800s prior to park establishment and the evidence of low densities in the 1870– 83 market-hunting era. The evidence does not support the view of a rare, episodic establishment adopted by the park (Bishop et al. 1997:52, 53, 75; Meagher and Houston 1998:245).

Aspen Trend during the First Elk Population High I started the first elk high-population period with the herd rising above censuses of 6,000 in 1885 (refer to figure 5.2). I hypothesized a continuous population increase, on average, to the 1914 count of 27,800, on which Smith et al. (1915) commented approvingly. I then suggest decline in the following decades but still numbering on the order of 19,000 in 1917 according to Graves and Nelson (1919). Early aspen photographs (compare figure 6.7, Kay 1990:figure 23) taken in 1893 some 9 years into the 1885–58 period with a northern herd numbering

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more than approximately 6,000 (figure 6.2) show young trees well within elk browsing heights, branches also within those heights, and trunks unscarred by elk chewing. By summer of 1914, some 30 years into the first population high (see figure 5.1), Smith et al. (1915) observed, “In places the bark is peeled from aspens to such an extent that whole groves of the small trees are killed.” Barmore (1980:371) commented that “deterioration . . . first became apparent or was first mentioned in the late 1920s,” and inferred that this was the beginning of accelerated aspen decline resulting from unspecified change in conditions (p. 375). But Skinner, who commented in the 1920s, was the first biologist assigned to YNP. Except for the Smith et al. (1915) survey, there had been no professional biologists previously assigned to the park who could have observed what was occurring and comment on it. The herd increase had been under way for approximately 40 years. It must have taken a period of years for the effects of the enlarged herd to become evident. Even so, effects were evident to the 1914 Smith et al. expedition. Romme et al. (1995) found that only 3% of the trees they sampled had been established after 1921. Only 5% of Ripple and Larsen’s (2000) trees escaped in this period. The latter authors pointed out that 1914–26 was a period of intensified wolf control and eventual eradication and that the failure of aspen tree establishment followed this period. By implication, wolves had limited elk numbers sufficiently to allow aspen recruitment prior to their eradication. This hypothesis has received widespread notice, including the syndicated media (see Young 2000; Lane 2002). However, the northern herd rose precipitously from 1884 to the 1910s while wolves were still present, then either leveled off or declined until 1968, well after wolves were gone. Thus there is no evidence in these trends that wolf predation was a significant determinant of the elk herd trajectory during this period. It may lend support to Kay’s (1990, 1994a) paradigm that aboriginal hunting was the major constraint on elk numbers before 1872. Moreover, emphasizing the absence of aspen reproduction after the 1920s places too much weight on that cessation. Samples by Romme and colleagues show abundant aspen-tree recruitment in the latter decades of the 1800s, dropping off sharply after 1890. My analysis of Warren’s beaver stumps shows a sharp drop after 1900, as do Ripple and Larsen’s two data sets. These indicate the beginning of major reproductive inhibition of aspen trees by an elk herd now risen to high levels (see figure 5.1) when wolves were still present. That inhibition has continued to the present and indicates that the 1920s wolf eradication was not a major influence on aspen recruitment trends in the twentieth century. By the 1930s, roughly 50 years into the high densities, all authors were reporting heavy impacts: Bailey (1930), Wright and Thompson (1935), Grimm (1939). Grimm established three 6 × 6 m exclosures in 1934 on the northern range on sites with aspen suckers or saplings. Successive photographs on one exclosure (Range Plot 10) in 1939 and 1965 show robust growth of aspen trees inside but none outside (Kay 1990:88).


Kay (1990:104–8) showed sequential photographs of Grimm’s Range Plot 25 in Yancey’s Hole. This exclosure began as a hay corral in the 1920s with the first photograph in 1932 showing a dense stand of aspen saplings perhaps 2 m tall inside. A 1986 photograph (see figure 1c in Keigley and Wagner 1998) of the same exclosure shows a robust, multiage group of trees up to 15 m high inside. There are no aspen trees on the outside, but short (< 1 m) suckers can be found in the meadow in the vicinity of the exclosures. W. H. Kittams (1959) identified 394 young aspen “trees” > 2 m tall in 1947, and recounted them annually through 1952. A total of 300 (76%) died by the end of the study, in Kittam’s view because of heavy browsing. Kittams (1948) also presented a series of six photographic pairs. The first shots were taken in 1887, 1922, 2 in 1923, and 1 each in 1927 and 1937. These were matched with his own shots taken in 1948. All showed sharp reduction or absence of trees in areas ranging from the Lamar River valley on the east to the Mammoth headquarters on the west. In total, there is abundant evidence of aspen decline, at least of trees, during the first period of elk abundance. The evidence is not continuous in space and time, and it is of varying forms. But there is little doubt of its reality.

Trends during the Population Reduction Barmore (1980:359–60) measured aspen sucker height and density inside and outside a 6.1 m2 exclosure at the Mammoth Park headquarters in 1965 and compared his results with earlier measurements. There was no increase in sucker density between 1935 and 1965 on the outside and actual decrease inside, which he attributed to self-thinning. Sucker height increased on the inside, but not on the outside. He commented that “disappearance” of aspen on the site was “far advanced” when the exclosure was established in the “early 1930s.” He made similar measurements inside and outside the Yancey’s Hole Range Plot 25. He noted that 81 of the 86 exposed trees outside the new exclosure were dead by 1941. The density of suckers outside the exclosure declined from 1936– 43, then increased from about 0.02–0.3 ft–2 from 1943–65, but the heights were suppressed by browsing. The latter increase in outside sucker density coincided with the growth and maturation of the trees inside, perhaps another piece of evidence that clones with trees are more vigorous than shrub clones. At two exclosures established near the Lamar River in 1957 and 1962, there were no changes in sucker densities either inside or outside by his measurements eight and three years later. Sucker heights did not increase outside either exclosure, or inside one, but they did increase inside the other (p. 363). Referring to measurements in 20 aspen stands across the northern range, Barmore commented (p. 364): “The consistent but insignificant increase in mean sucker height after 1964 and 1965 probably reflected addition of sample stands on the fringes of the winter range when some increase in sucker height occurred after reduction of the northern Yellowstone elk herd.” Barmore found that sucker browsing remained high during the population reduction. He (p. 364) measured mean “leader use” between 1963–69 at 50%,

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80%, and 44% outside the Mammoth, Yancey, and combined Lamar exclosures, respectively. In 5 of the 20 stands he measured over the northern range, mean overwinter leader use during the same period was approximately 75% of these observations. He commented (pp. 364, 369): Browsing on P. tremuloides suckers begins as soon as elk arrive on the winter range. Leader use was about 40 percent in early December 1964 when other forage was easily available. . . . By early to mid January 1965, mean use of leaders above the snow surface reached the winter-long maximum of 60 percent. Similar browsing early in the winter has been reported by others for the northern Yellowstone winter range. . . . Most suckers that project above the snow are consumed before mid or late winter when they otherwise could provide higher-quality food than cured grasses. Bart O’Gara, former leader of the Cooperative Fisheries and Wildlife Research Unit at the University of Montana, was a doctoral student at (what is now) Montana State University in the 1960s conducting research in YNP on pronghorn. He commented to me (personal communication, February 8, 1996) that he observed some aspen recovery by 1967. When he returned to the park five years later, that growth had disappeared. Kay (1990:78–81) analyzed the age composition of aspen trees in 74 stands near Eagle Creek, an area ~5 km north of the park’s northern entrance to which some elk move in winter when the population is high or winter conditions are severe. Fewer animals move out to that point at low population densities and mild winter conditions. This density-dependent pattern of winter exodus was discussed in chapter 2. Kay noted that a substantial number of trees had grown and survived in the 1960s whereas few had become established prior to or after that decade (figure 6.8). In total, there is limited but nevertheless suggestive evidence of slight aspen response during the population reduction. That response came after the clones must have experienced heavy use for at least a half century by the first large population increase, and their vigor was doubtless reduced. Moreover, the time for recovery was short. I have placed the population reduction dates at 1959– 70 (chapter 5), encompassing 11 years. Barmore’s comments also indicate that aspen is highly palatable to elk and sought advertently. As a result a lower population heavily utilized what must have been a greatly reduced resource, which was then largely prevented from significant recovery. If there was any recovery not observed by the limited research in the 1960s, it must have been eliminated by the burgeoning population of the following 33 years. The northern herd increased in the latter 1800s into a lush and extensive vegetation that had not previously been subject to heavy herbivory. It may have taken a number of years before the herd had sufficiently reduced the vegetation to have a continuing impact. And by the 1960s, the aspen resource may have been so depleted that even a reduced elk herd could suppress the species’ growth.


Figure 6.8 Percent of aspen stands (n = 74) regenerating by five-year intervals in Eagle Creek area outside of park northern entrance, and number of elk wintering north of the park (Kay 1990:figure 5).

Trends during the Contemporary Population High Little more can be said in reference to the current situation. Kay (1990), Romme et al. (1995), and Ripple and Larsen (2000) found no ramets escaping to tree stature in the northern range, and the latter authors’ evidence indicates no significant tree establishment for the past century, whether the stems were removed as small new growth or young trees killed at later stages. The remaining aspen trees on the northern range are at the limits of their longevity. If the browsing pattern of the twentieth century continues, and no ramets are allowed to form trees, aspen can be expected to disappear almost totally within the next decade or two as a deciduous woodland type. Some clones remain in shrub aspen form, but the evidence discussed next indicates clone disappearance and shrinkage. Thus aspen as a species in the northern range and as organisms living in the area for millennia, are also declining. How much longer shrub aspen clones can survive is an open question.

Trends in Clone Abundance As discussed, under continuous browsing a clone loses vigor. To examine whether the clones can survive for significant lengths of time, Kay and Wagner (1996) searched photographic archives for pictures of YNP taken before 1900 that showed stands on sites that could be relocated. They examined the sites in the field to determine whether the clones still existed. They measured the areas of the surviving clones in 1986 and estimated the areas of the same clones as best they could in the photographs. Of the clones represented in the early photographs, one-third had disappeared. The remaining two-thirds had been converted to shrub aspen, and the mean area of these had declined to 20% of the areas interpolated from the photographs. These results are consistent with Mueggler’s (1994) findings on 713 aspen “stands” in nine national forests in Utah, Idaho, and Wyoming: Clones that show diminished vigor may not “have the potential to replace themselves successfully.” The Kay and Wagner (1996) sample indicating disappearance of one-third of the clones and 80% shrinking in the areas of the surviving clones implies a shrinkage of 87% in the total area occupied by the clones in this sample: 33 + (0.8 × 67) = 87

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This assumes that the average areas of the eliminated clones and those of the surviving ones were approximately the same. The result is of the same magnitude as the values already cited, and once again far above the values typically cited in park publications.

CAUSATION Elk Browsing The evidence of elk effect on trends in aspen abundance is of four forms: 1. The chronological correlation with elk numbers. The Warren (1926) and Ripple and Larsen (2000) data, plus photographs of mature trees around the turn of the century, indicate that tree establishment occurred widely over the northern range in the decades prior to park establishment. According to the weight of the evidence, elk occurred in low densities and migrated out of what is now the northern range in winter in these years. Tree recruitment declined sharply after 1890–1900, 5–15 years after the herd had risen to high numbers, and stopped completely by the 1920s with continued high elk numbers. Professional observers began reporting heavy impacts by 1915. A sequence of authors reported decline. With the 1960s herd reduction to low numbers lasting at most 11 years, and with a much reduced forage resource, there is evidence of slight aspen response. Research since the population recovery has consistently reported almost total use of suckers and shrinkage or disappearance of clones. 2. Aspen morphology. Aspen in early photographs (see figure 6.7) shows (a) a diversity of age classes including young saplings that today would be browsed out, (b) trees with branches near the ground (i.e. no highlining), and (c) absence of bark scarring. Aspen trees in the park today (a) are uniformly of advanced age classes (100 to 130 years), (b) have no branches near the ground, and (c) are universally bark scarred. Clones without trees are in shrub aspen form. 3. Inside-outside park comparisons. Stands near to but outside the park and at comparable elevations largely have characteristics of 1893 stands (figure 6.1). 4. Exclosures. The park established six exclosures, standing today or until the 1988 fires, on sites that have/had aspen clones: Range Plots 10 and 25 in 1934 and 1936, and Mammoth, Junction Butte, Lamar West, and Lamar East (1957 and 1962). Three of the last four are on sites with surface-water seepage, the clone root systems experiencing similar inside-outside moisture environments. All six of these exclosures have/had robust, mixedage aspen stands on the insides. There are no trees on the outsides, although some produce root sprouts that are typically less than 1 m tall.


The 1988 fires inadvertently produced small exclosures on the northern range in the form of jackstraw piles of dead conifers that had fallen over after the burn. Ripple and Larsen (2001) measured aspen sprouts inside and outside 28 of these on the northern range and found mean inside height of aspen shoots to be 1.46 m and those outside the piles 0.54 m.

Climate Change Park publications have postulated climate change as one contributor to aspen decline. Houston (1982) invoked the drought of the 1930s. More generally (p. 103) he pointed to a mean annual temperature increase of 0.5–1.0°C from the latter 1890s to 1972 at the Mammoth Hot Springs Park headquarters. He also summarized rainfall for the same period, which suggests a slight increase in springsummer moisture, decrease in winter, and a reduction in mean, annual totals of perhaps 20 mm (about 5%). Park officials also point out that YNP was established at about the end of the Little Ice Age (ca. 700–100 years B.P., Hadly 1996) and the 1872 vegetation must have been a response to its cooler, wetter climate. Balling et al. (1992) calculated that park summer temperatures increased 0.87°C from 1890 to 1990, and seasonal antecedent precipitation declined 61 mm. Hence Houston and Balling et al. do not agree on summer precipitation trends. The most recent analyses (Kittel et al. 2002; Baldwin 2003) show slight precipitation increase in the northern Rockies, primarily in summer, during the twentieth century. There is no evidence that what appear to be real but slight climate changes have been involved in the aspen decline, nor has any ecophysiological mechanism linking the two been proposed. As Renkin and Despain (1996) comment, “Most attempts to define the influence of climate on aspen are philosophical . . . rather than empirical.” It is a hypothesis that has been repeated so often that it appears to take on a tangible level of probability, notwithstanding the absence of evidence. Several considerations argue against a significant climate role in the northern range aspen decline. As already discussed, there is evidence that the decline began before or by the turn of the twentieth century, and concerns were being voiced by 1915, both before or barely into the climate changes of 1900s. At that point the twentieth-century climate change was not a statistical reality. Aspen increased inside exclosures during the 1930s drought (Range Plots 10 and 25). Where subjected to elk browsing, it declined in moist areas in the park, as Houston (1982) and Kay (1990) have pointed out. And as Kay (1990) commented (see figure 6.8), aspen established ramets that matured to tree stature outside the park in years when few elk wintered in the area. In general the species is more vigorous—lacking highlines, showing clonal expansion with suckering on the fringes of stands—outside the park. Thus aspen has fared well in this century inside the park when protected from elk browsing and has declined on optimum sites in the northern range when not protected.

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In total there is no tangible evidence indicating, and substantial amounts arguing against a significant role of climate change in the northern range aspen decline of the past century. The National Research Council (NRC) review of the northern range situation (Klein et al. 2002) concluded that “the twentieth century fluctuations in annual indices of drought stress fall within normal fluctuations over the past several centuries . . . in Yellowstone and throughout the Rocky Mountains, climatic change does not appear to be closely associated with aspen regeneration . . . aspen regeneration occurred during both favorable and unfavorable climatic conditions between 1871 and 1920.”

Fire Suppression Another factor commonly cited in park publications as a major factor in northern range aspen decline is fire suppression. Houston (1982:127) simply concluded that “widespread fire suppression was the primary cause of the [aspen] decline.” An extensive body of literature portrays three modes of interaction between fire and aspen. Although commonly held to be a seral species, Kay (1998) considers this not to be an appropriate characterization because of the clonal growth form, the ages and relative immobility of the clones, and the species’ consequent inability to move into many available disturbed areas. Moreover, aspen scholars consider many stands in the northern Rockies to be climax, probably on sites too xeric to permit conifer growth (Beetle 1974b; Jones and DeByle 1985; Mueggler 1989). Nevertheless, conifers can and do move into and replace aspen growth in some situations. Gruell (1980a, b) presents evidence for the preemption of aspen stands by conifer invasion in the Bridger-Teton National Forest south of Yellowstone. Fire is commonly implicated as a deterrent to such invasion and is one of three means by which fire could maintain aspen in abundance, and fire suppression could facilitate aspen decline. Second, some aspen stands in western North America appear to senesce as trees mature and exert apical dominance, which prevents adequate replacement by suckers (Krebill 1972; Schier 1975; Mueggler 1989, 1994). In many cases, destroying the older trees with fire stimulates dense sucker growth. Third, some authors have attempted to reconcile the abundance of aspen in prehistory, or prior to fire suppression, with hypothesized early elk abundance (see Gruell and Loope 1974). They have postulated an analog of large prey numbers swamping predators by suggesting a need for extensive, large-scale fires that stimulate so much aspen sucker growth that the ungulates cannot make significant inroads. There is considerable evidence for the frequency of fire in the YNP area prior to and around the time of park establishment. Houston (1973) found evidence for a 20–25-year prepark fire frequency on the northern range. Risbrudt (1995) reported mean fire intervals mostly at 20 years or less in the Northern Rocky Mountains between 1500 and 1950.


A number of early photographs of the northern range, some with burned trees in the backgrounds, show dense clumps of young ramets that must have been stimulated by fire: see Houston (1982:plate 34), Kay (1990:figure 19) for two 1885 photos; Kittams (1948:figure 21) for an 1887 picture. Later photos around the turn of the century show older, but still young, even-aged aspen thickets that may have been released by fire: see Meagher and Houston (1998:plates 49.1, 54.1) and figures 6.3 and 6.4 in this chapter. These were not universal in the northern range, however. Other photographs show stands with mixed age classes (Kay 1990:figure 23; figure 6.5 here). Aspen researchers have tended to stress the role of lightning in the frequency of early fires (Gruell and Loope 1974) as have YNP publications (Despain et al. 1986). But Kay (1990, 1998) points out that aspen vegetation is notoriously fire resistant and difficult to burn except in spring and fall. At the same time, lightning frequency is highest in summer when aspen has what researchers call its “asbestos” characteristics (DeByle et al. 1987). Kay presents evidence of frequent lightning strikes on the northern range during the twentieth century when aspen burns have been few (Loope 1971; Despain et al. 1986) and small except for the 1988 fires. Rather than lightning as a common prehistoric ignition source, Kay stresses the role of aboriginal fire use. The basic question then is whether and to what extent fire suppression (whether through park management or the elimination of aboriginal burning) has operated through these three modes of aspen–fire interaction to contribute to the aspen decline in the northern range. First, the predator-swamping hypothesis would seem to have been falsified by the 1988 fires and the Romme et al. (1995) and Ripple and Larsen (2000) observations of total sucker removal in this century. The phenomenon might have prevailed in prehistory when elk were present in low numbers and aspen was abundant. But it could not occur under the realities of the park conditions of the early 1990s and quite possibly through the entire period of high elk numbers since 1885, absent the short period of population reduction. Second, there is evidence that fire stimulates sucker growth in some clones but not invariably (see Gruell and Loope 1974; Renkin and Despain 1996). Aspen has regenerated in exclosures in the northern range in the absence of fire (Houston 1982; Kay 1990). Hence fire is neither essential to aspen regrowth, at least on some sites, nor does it stimulate growth in some situations. In any event, fire again could not have fostered aspen recovery or stemmed its decline with the large elk numbers in this century. This leaves the question of conifer invasion. There is valid evidence that some aspen stands have been invaded by conifers in the northern range (see Houston 1982:plates 34, 36; Kay 1990:figures 16, 19), and wider evidence of increased conifer densities over much of the park (Houston 1982; Renkin and Despain 1996; Meagher and Houston 1998). The questions then arise as to its prevalence, and its proportionate role in the sum total of factors reducing the northern range aspen. The need is for measurements of the area dominated by aspen at park establishment that has been invaded by conifers to varying degrees. As

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first approximations, measurements of cover for the two types within aspen clones could serve as metrics. Two studies of conifer invasion, in the northern range have been reported. Houston (1982:414–15) examined 175 aspen stands in the northern range inside the park, and 28 outside in the course of a 1970 range reconnaissance. He commented that “About 70% of 203 stands showed substantial invasion by conifers.” He, as did Kay (1990:77), pointed out that he sampled stands as he encountered them, and they were probably biased toward stands along roadsides, hence were not a random sample of the northern range. More important, it is not clear what he meant by “substantial invasion.” A clone can have significant numbers of conifers growing within it and still function ecologically as an aspen stand. Kay (1990:63–68) examined a random sample of 194 aspen stands inside the park and 74 outside in the Eagle Creek area north of the park’s north entrance. He found that 79% of the park stands and 91% of the Eagle Creek stands had 10% or less conifer canopy cover. It was not clear what the ranges of canopy cover percentages were in the 21% of park stands and 9% of Eagle Creek stands with >10% conifer cover. But clearly the fraction of stands in which conifers had developed to the point of largely dominating them was quite small. In the park, 49% of the stands and in Eagle Creek 57% contained no conifers. In total, some conifer invasion has occurred, but it is not possible to place a value on it. The evidence suggests that it has not been a major factor in the northern range aspen decline. Kay concedes that any aspen clones that had been completely usurped by conifers since park establishment would not have been evident to him and would not therefore have figured into his calculations.

Senescence, Disease, Fungi, Beaver As discussed, aspen trees die of a number of causes (Krebill 1972; Jones and Schier 1985; Mueggler 1987), and if nothing else of old age by virtue of their relatively short life spans. Thus aspen ramets of all ages present at the time of park establishment would have largely died out over the 132–year park history even without a major elk population increase. This turnover would have been hastened in the vicinity of streams and ponds by beaver cutting. Indeed, Warren (1926) was so concerned about the number of aspen being cut by beaver in the 1920s that he advocated a beaver-control program. Fires also hasten the process; and elk in large numbers hasten it by bark stripping and girdling the younger trees and by providing entry of lethal fungi into the more mature trees by biting off pieces of bark (Krebill 1972; Hart 1986). Thus it is entirely likely that most of the trees alive in 1872 have died by now. The trees that are present today apparently established during the late 1800s and early 1900s. They have reached their typical longevity but have managed to survive so far. Bishop et al. (1997:51) scoff at Kay’s estimate of 95% decline in park aspen. But if the reference is to trees (as in Kay 1990:206), this percentage could approach reality as discussed above. Many of the surviving park aspen


clones that still have trees are characterized by a scatter of a few survivors, a few dead trees still standing, and a litter of dead logs on the ground that once constituted the trees of more robust stands (see figure 6.6).

Juvenility, Secondary Compounds, and Leader Browse Until publication of the Romme et al. (1995) study, park research and publications struggled to find a rationale for an apparent contradiction between the assumptions of the natural-regulation hypothesis and certain lines of evidence. Houston (1982) surmised that the northern herd had numbered on the order of 12,000–15,000 in prehistory, and that has since been taken as a virtual foregone conclusion in park thinking, publications, and discourse. At the same time, the prevailing view, based in part on the photographic record, was that aspen was a vigorous if not necessarily dominant part of the vegetation in prehistory. The question needing reconciliation was how a vigorous aspen vegetation could coexist with an elk herd of that magnitude in the 1800s when aspen was obviously decimated by an equivalent number of elk in the 1900s. Thus, Bishop et al. (1997:56) comment, “There is considerable uncertainty over why that browsing has a different influence now than it has had historically.” Resolution of this dilemma has been a major impetus to the proposals mentioned of causal factors other than elk: climate change, fire suppression, etc. If somehow the environmental conditions in the two centuries were different, perhaps that could explain a surmised difference between the two periods in elk–aspen interaction. One hypothesis without supporting evidence and one set of observations that has never been published, documented, or well described by the observer, have been proposed as possible explanations of these surmised century differences and have been cited repeatedly in park publications (Renkin and Despain 1996; Bishop et al. 1997; Huff and Varley 1999). Despain (1990) observed that aspen suckers in their first year of life in some cases grow as much as 2 m in height. He also noted that investigations elsewhere had measured high levels of secondary compounds in first-year aspen ramets (see Bryant et al. 1992). He speculated that under favorable conditions (e.g., climatic conditions that perhaps prevailed in the 1800s), chemically protected ramets might grow swiftly in 2 or 3 years, during which time they were not browsed because of the protection, and at the end of which their leaders would have grown beyond elk reach. There are several problems with this hypothesis. First no secondary-compound measurements have been made in the northern range aspens. It is thus not known what the levels are or how they compare in well-defended and poorly defended ramets. Second, there are no observations to show that the defenses are strong enough to deter browsing by nutritionally deprived animals. Third what constitutes “good conditions” for significant defense and rapid growth has not been determined, nor has whether the effect exists at all. It remains speculative. Fourth, it is difficult to accept the idea that favorable conditions, if they

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exist, would only occur in the 1800s and never in the next century. Finally, a 3year escape for a ramet does not make it invulnerable to elk attack. Highly palatable and probably nutritious, young aspen have their bark readily removed by elk (see figure 6.3), and there are numerous reports of elk pushing down young trees to browsing levels. In total, the evidence available at this point does not support this hypothesis. The other set of observations has been alluded to briefly in a number of park publications. D. G. Despain split a sample of 50 aspen logs from the northern range to examine their internal morphology (Despain 1990:99; Renkin and Despain 1996; Bishop et al. 1997). He found that the terminal buds had been browsed off in 49 of these trees in their first year or 2 of life. But the plants had continued to grow, eventually maturing to tree stature with the mass of the boles encircling and encasing the browsed tips of the original main stems. The inferred browsing history could only be disclosed by splitting open the logs. The authors inferred from these observations that the browsing of 49 out of 50 terminal buds implies browsing intensity similar to levels experienced in the twentieth century (see Barmore’s 75% shoot utilization and the Romme et al. and Ripple and Larsen total sucker removal). Yet the plants in Despain’s sample continued growth to tree stature, whereas the Romme et al. and Ripple and Larsen data indicate minimal recruitment of trees since 1890–1900 and none since the 1920s. Renkin and Despain (1996) conclude: “It appears that aspen utilization levels being experienced today are . . . similar to those experienced a century ago. The determining mechanism . . . that allowed for aspen to grow beyond the browse influence is not exerting the same influence today.” But because the plants continued growth to tree stature, lateral stems must have assumed dominance and formed the boles of the trees (Keigley and Frisina 1998). And the browsing pressure must not have been sufficient to prevent these stems from becoming trees. R. B. Keigley (personal communication, September 4, 1998) reports that this is common growth form in northern Rocky Mountain aspen, and that a variety of forces—other vertebrate herbivores, insects, storm damage—can injure terminal buds. No ages have been reported for these trees, but Renkin and Despain comment that they were “>80 years in age.” But because the Romme et al. and Ripple and Larsen data imply virtually no tree establishment since 1890–1900, it is reasonable to assume that Despain’s trees began life before 1890 when the northern herd was at low levels and browsing pressure was light.

General Discussion on Causation Prior to adoption of the natural-regulation policy (Anonymous 1967a) and enunciation of the natural-regulation ecological hypothesis (Houston 1971), virtually all investigators implicated elk browsing as the only significant factor in aspen decline. This of course was the rationale for the herd-reduction policy. The natural-regulation hypothesis, posed when elk numbers were still low from the herd reduction, contained the postulate that the elk would limit their


own numbers without extreme impact on the northern range ecosystem. Subsequent research generated two successive lines of inference, somewhat contradictory, that were consistent with this postulate. One line implied that there was little change in the system. Thus, “In general the studies to date show surprisingly light biotic effects overall and suggest a relatively stable ungulatevegetation system” (Houston 1982:135). “In general, the photographs revealed that most of the range sites that were considered over-grazed looked the same way in the 1870s and 1880s. The range conditions that so alarmed managers in the 1930s and later were actually range conditions that had been there all along” (Despain et al. 1986:87). The second, more recent line of inference tacitly acknowledges that there have been significant vegetation changes, but dilutes or discounts elk influence by implicating other factors (e.g., climate change, fire suppression, succession, variations in production of secondary compounds, etc.). In a succinct and balanced treatment, Barmore (1980:371–76) changed his views in the 1970s, hypothesizing a complex of causes. Houston (1982) found correlated changes in other factors (climate, fire, succession) and inferred causation. Other authors have followed suit (Despain et al. 1986; Romme et al. 1995; Bishop et al. 1997). But there is little if any evidence of significant effects by most of these factors. Nevertheless they are hypothesized repeatedly to the point where they are treated in park circles and public-information releases as established fact. The result has been to discount elk effects to varying degrees anywhere from considering them only one among a complex of roughly equivalent factors to nearly discounting them entirely. Thus, Houston (1982:127) wrote: These data seemed consistent with the hypothesis that the decrease in aspen was largely the result of fire suppression . . . in a changing climate. . . . Herbivores doubtless affected the rate of postfire succession but were not the driving force that determined the direction and timing of events. In a balanced treatment posing several factors, and acknowledging that the influence of fire suppression is a hypothesis, Despain et al. (1986) nevertheless concluded that “something in addition to elk pressure must be contributing to the decline of aspen” (p. 104). Romme et al. (1995) had firm evidence only of intensive elk browsing obtained during their field studies, yet invoked a complex of factors—“numbers of elk, beaver, and moose were low, fires had occurred recently, climatic conditions were moist, and wolves were present”—to explain their interpretation of contemporary aspen age distribution and an inferred recruitment pulse in 1870–90. “This combination has not occurred since then” nor by implication had it occurred in the decades prior to 1870. “A single-factor approach to understanding and managing aspen dynamics in YNP is not sufficient. . . . We are not able to say confidently which [of the proposed factors] are of greatest importance.”

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Bishop et al. (1997) promptly adopted the Romme et al. scenario but were not so reticent in interpreting the latter authors’ data or adopting single-factor causation: it may be safe to assume that aspen only rarely meet circumstances in Yellowstone that allow them to escape browsing and grow to tree-size . . . these aspen studies . . . provide circumstantial support for historical analysis . . . indicating that ungulates were abundant in the Yellowstone area before 1870. (p. 53) Yet these authors retreated 25 pages later to the earlier park position: “The changes in northern range vegetation that have caused observers to raise alarms for decades are either not the result of elk ‘overpopulation’ or are part of longterm processes we do not fully understand” (p. 78). The prompt adoption of the Romme et al. scenario by park officials (see Bishop et al. 1997; Schullery 1997a, b; Huff and Varley 1999) is explained by its resolution of a second conundrum (the first was significant elk effect or not) with which park personnel had struggled for some time. Contrary to authors of the previous half century, Houston (1982) had concluded that elk were numerous in the area prior to park establishment. Yet Houston and his successors believed that aspen had been a significant (if not major) component of the vegetation, thus coexisting with elk in ways that it could not in the twentieth century. This dilemma produced such statements as the Despain et al. (1986) “something in addition” quote and Bishop et al. (1995:56) “browsing has a different influence now than it has had historically.” The Romme et al. interpretation that northern range aspen had only established in the 1870–90 market-hunting period when elk numbers were sharply reduced provided reconciliation. The evidence all along had been convincing that elk browsing had been the major deterrent to aspen regeneration in the twentieth century, even if previous authors had declined to concede it. The Romme et al. interpretation that aspen regeneration had been rare before 1870 was now consistent with Houston’s claim that elk had been abundant prior to park formation. The dilemma was now apparently resolved. However, there is persuasive evidence that mature aspen trees were a significant component of the northern range vegetation in the early decades of the 1800s, contrary to the Romme et al. interpretation. This is consistent with the weight of the evidence presented in chapter 3 that elk were present in low densities prior to park establishment. Bolstered by the accumulated evidence including that of Romme et al. and Ripple and Larsen, I agree that elk browsing has been the dominant force preventing aspen regeneration and inducing its decline in the twentieth century. I do not suggest that other factors may not have had at least some influence. Fire suppression, beaver, and perhaps other factors may have played a minor role. But the evidence convincingly points to elk browsing, occasioned by a vastly increased elk population since park establishment. This pattern is repeating itself in numerous areas of the West, in many cases demonstrated experimentally with large exclosures (see Baker et al. 1997 for Rocky


Mountain National Park; Kay 1998 for two national forests in Utah; Shepperd and Fairweather 1994 for Arizona). Klein et al. (2002:67) comment in the NRC report: “All data sources indicate that the abundance of large aspen in northern Yellowstone has declined during the twentieth century, and most indicate that this decline is due primarily to ungulate browsing.” Finally, I wish to respond to a comment in an otherwise excellent paper by Coughenour and Singer (1996). Noting that woody plant browse is at present a minor component of the northern range winter elk diet, and speaking in the context of the degree to which forage depletion limits the numbers of the northern herd, these authors commented that elk abundance and woody browse on the northern range are “largely decoupled.” However, the paucity of browse in the northern range elk diet is almost certainly due in part to the fact that most of it has been eliminated by the elk. Aspen is highly palatable. Kittams (1948) quoted Skinner’s early comment that aspen was an important dietary component for elk at all times of the year. Thus if elk are “decoupled” from aspen at present, this was less the case 100 years ago. More important, herbivory is a two-way interaction. The evidence indicates that elk browsing has been the main cause of aspen decline in this century and continues to prevent its recovery by browsing the annual growth shoots. From this perspective of herbivory, elk are still tightly “coupled” to aspen.

THE ASPEN WOODLAND ECOLOGICAL SYSTEM Understory Vegetation The discussion so far in this chapter has addressed only the biology of aspen, the species, and the factors inducing its decline. Yet aspen is only the dominant plant species in an ecological subsystem within the northern range ecosystem and as such assumes its ecological importance in the structure and function it provides to the subsystem and ultimately the entire ecosystem. This interactive role, except as a food resource for elk, and as both food and building material for beaver, has received only a fraction of the research attention that the species itself has received. Kay (1990) conducted extensive research on the understory vegetation in northern range aspen stands. He measured aspen understory vegetation inside and outside 7 exclosures in YNP, 6 in the Bridger-Teton and Gallatin National Forests within the Greater Yellowstone Ecosystem (GYE), and 1 near Pinedale, Wyoming. He measured percentage cover of each species in his sampling plots. His results from his tables 63–67 are summarized here in table 6.1. Two, marked inside-outside differences, were consistent at both the park and nonpark exclosures: The exclosure interiors had much higher shrub canopy coverages than the exteriors, whereas the latter had much higher grass cover than the interiors. The differences in both vegetation types were highly significant at

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Table 6.1 Percentage Cover and Species Densities of Shrubs, Forbs, and Grasses Inside and Outside 14 Exclosures in YNP Region as Measured by Kay (1990) Mean Percentage Cover and Species Densities Shrubs

Forbs

Grasses

Exclosures

X|% Cover

X| No. Spp.

X|% Cover

X| No. Spp.

X|% Cover

X| No. Spp.

7 YNP, Inside 7 YNP, Outside 7 Nonpark, Inside 7 Nonpark, Outside

46.3 10.1** 91.0 33.9**

5.3 3.7 9.6 8.6

29.7 22.6 20.1 33.3*

2.4 2.7 4.7 5.6

22.1 65.1** 8.6 31.9**

2.3 2.9 2.6 3.6

Note: Park: Total of means spp. inside 10.0, outside 9.3. Nonpark: Total of means spp. inside 16.9, outside 17.8. *Significantly different at 0.05. **Significantly different at 0.01.

p < 0.01; and they were roughly inversely correlated: the nonpark exclosure interiors had the highest mean shrub canopy cover at 91.0% and lowest mean grass canopy cover at 8.6%. The exteriors of the park exclosures had the lowest mean shrub canopy cover at 10.1% and the highest grass canopy cover at 65.1%. Forb cover was slightly though significantly (p < 0.05) higher inside park exclosures than outside, but the reverse was true in nonpark areas. Species densities were not consistently different, statistically, between the interiors and exteriors for any of the vegetation classes, but there were marked differences in the relative proportions of the three vegetation types. At the park exclosures, timothy (Phleum pratense) and bluegrass (Poa pratensis) were the dominant grasses both inside and outside. But because the grasses were a much larger component of the vegetation outside the exclosures, the mean canopy cover for each of the species outside the exclosures was approximately 3 times the cover inside. These two species were inconspicuous or absent at all of the nonpark exclosures except one in the Gallatin National Forest. These values, especially including those for the exclosure exteriors, can be tentatively interpreted in terms of elk effects. In the park, with its extremely heavy elk density, shrubs and forbs are reduced to low components of the understory vegetation, which is converted to a dominance of grasses (refer to figure 6.4). At the exteriors of the nonpark exclosures, there is apparently some elk pressure, judging by the inside-outside differences in shrubs and grasses, but not as heavy as in the park and permitting higher shrub and forb cover and lower grass cover than the exteriors of the park exclosures. Hence a gradient of elk pressure, from high to low, might be from park exclosure exteriors to nonpark exteriors to exclosure interiors of both areas. One other indication of the heavy park pres-


sure appears to be the abundance of exotic timothy and bluegrass outside the park exclosures and their near absence outside the nonpark exclosures. Kay (1990) also surveyed understories in aspen stands in three areas of the northern range and in the Eagle Creek area just outside the park north boundary. The latter area receives some winter elk use depending on the size of the northern herd (see figure 6.8), but nowhere near the consistent, annual pressure sustained by the northern range inside the park boundary at all levels of the elk population. Kay sampled aspen understories with 2 × 30 m transects, 74 in Eagle Creek and 194 inside the Park. His results, summarized from his table 11 are: Inside the park, the mean percent canopy cover for shrubs, forbs, and grasses was 10.0%, 32.5%, and 58.8% respectively, all different from each other at p < 0.01. In Eagle Creek, the corresponding percentages were 16.9%, 63.9%, and 18.1%, with the forb percentage different from shrubs and grasses at p < 0.01 but the latter two not significantly different. And the differences between Eagle Creek and the park in each of the three vegetation classes are statistically significant. These results are consistent with the exclosure evidence. Under the complete protection of exclosures, aspen understory vegetation is dominated by shrubs, usually of 5–10 species, and ranging from 46% to 91% cover (table 6.1). This pattern can be seen today by casual inspection of the dense shrub understories of the Mammoth, Junction Butte, Lamar West, Lamar East, and Range Plot 10 exclosures. Forbs are the intermediate type with cover ranging from 20% to 30%, whereas graminoids (grasses and sedges) are the least abundant class at 9–22%. At the other extreme of herbivorous pressure inside the northern range, graminoids are the dominant vegetation type at 59–65% cover and dominated by the two nonnative species P. pratense and P. pratensis. The latter result is not new, Houston (1973) having commented more than a decade before Kay’s work on the dominance of Phleum in the aspen understories. Shrub canopy cover inside the park has been reduced to about 10% (table 6.1). Forb cover remains intermediate between the other two classes (but see figure 6.9). Under intermediate levels of elk use (i.e., outside nonpark exclosures), the shrub component, though less than in exclosures, still exceeds the small amount on the northern range. The grass component is less than that outside the park exclosures, and the amount of nonnative grasses is also far below. The net effect of the heavy elk pressure within the park is to sharply reduce the structural and habitat diversity of the aspen understories by largely eliminating the shrub and forb stratum and replacing it with a graminoid vegetation dominated by two nonnative grass species.

Associated Fauna A major long-range effect of the heavy northern range elk browsing on the aspen woodlands is to reduce its habitat structural diversity from between 3 and 5 strata to 2. Initially, the younger aspen age classes that form on the outsides

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Figure 6.9 Dense forb understory in Eagle Creek outside park’s northern entrance. Compare with figures 6.4 and 6.6. Photographed July 19, 1986.

of the clones as the latter expand (figure 6.1) are browsed out, reducing the number of canopy strata from at least two to one. The understory shrub strata are browsed out, leaving only an herbaceous, largely grass stratum. This leaves two strata: a mature-tree-canopy stratum—which persists through the lifetime of the trees but thins, becomes patchy, and contracts as the trees die out—and the herb stratum (figure 6.6). When the trees reach the limits of their longevity and die out totally, the site is converted to the single herbaceous stratum if continued browsing of the aspen suckers prevents any recruitment of trees (figure 6.2) and certainly if the clone dies out. I am not aware of any research on the effects of these structural changes in aspen woodlands in the northern range on the vertebrate fauna. But a considerable amount has been done elsewhere in the West. Hansen (1997) studied a range of factors, including natural disturbance and logging, that affect avian diversity in the GYE. He concluded that “cottonwood, aspen, and willow have about twice the bird species richness and abundance of any of the lodgepole pine, sage, or grassland cover types.” In a survey of the avifauna in aspen and cottonwood woodlands in the National Elk Refuge south of YNP, Dobkin et al. (2002) observed 76 species “closely associated” with these vegetation types. Woodlands that lack recruitment and are structurally simplified had only “spare numbers of a small subset of these species.” Turchin et al. (1995) studied 14 aspen stands and their associated bird and plant communities in Rocky Mountain National Park in 1993.


They summarized a number of studies showing that aspen forests support greater bird species richness and total abundance than other North American montane habitats. Aspen had higher species richness than surrounding conifer sites, with eight species occurring in aspen that were not in the conifers. They concluded, “the amount of understory vegetation appeared to be the main determinant of bird species richness.” No complete system study has been conducted on the northern range that would allow calculation of total faunal diversity and how that would change with elimination of the aspen woodland subsystem. The closest approach was the 1976–78 research on a montane sere in northern Utah by James A. MacMahon and co-workers. These investigators conducted complete ecosystem studies of four seral stages similar to those of the northern range: meadow, aspen, fir forest, and spruce forest. These give at least a rough indication of what the elimination of aspen implies for the structure and function of the northern range system. Smith and MacMahon (1981) observed 43 breeding species in the four seral stages. Eight species, largely cavity nesters and edge species, were unique to the aspen. Elimination of aspen would thus have reduced the avian species richness of the combined vegetation types by 8/43 × 100 = 19%. The avifauna of Yellowstone is quite similar to that of the northern Utah area, and the disappearance of aspen is in all probability reducing the northern range avian diversity by a similar order of magnitude. Mean annual number of breeding birds in the Smith and MacMahon (1981) study were 147/km2 in meadow, 783/km2 in aspen, and 905/km2 in fir and spruce combined. Elimination of aspen, and conversion to the meadow type would have reduced the number of breeding birds in the area that had been aspen by (783 – 147)/783 × 100 = 81% In an area the size of the northern range, with similar proportions in meadow and conifers (Houston 1982:86), aspen on the order of 10,000 ha in 1872, and with breeding bird densities in the 3 habitats similar to those of Smith and MacMahon, conversion of aspen to meadow would reduce the total number of breeding birds on the order of 10%. In the same research, Anderson et al. (1980) conducted similar studies on the herbivorous small mammals along the sere. They observed 5 species in the meadow type, 7 in aspen, and 9 in the coniferous. Because all of the species in aspen also occurred in the coniferous types, conversion of aspen to meadow would not result in the disappearance of any mammalian species. But that conversion would eliminate two species from the areas that had been in aspen, leaving the five species characteristic of meadow and thus reducing slightly the average species density for the entire system. In these same studies, Waagen (1979) observed greater species diversity and biomass of ground-stratum spiders in aspen than in meadow. He attributed this to the greater number of habitat niches in the woodland vegetation. Although the group has not published comprehensive evidence on the abundance and diversity of other components of the arthropod fauna, the more abundant spider

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fauna in the aspen type than in the meadow probably indicates a more abundant prey base. Bailey and Whitham (2002) observed 69% lower arthropod richness and 72% lower abundance in Arizona aspen stands that had been severely burned and heavily browsed by elk than in stands that had been moderately burned and browsed. Beetle (1997) studied snails at eight sites in northern range aspen stands. She identified 21 land and 2 freshwater species, with the species per grove typically numbering 3–5. She observed 8–11 species in “damp” sites with favorable plant litter, soil, and moisture. She commented that the “extensive gnawing of tree trunks and grazing of aspen suckers by elk results in the decadence of aspen groves. This eliminates habitat for the mollusks.” Beetle pointed out that land snails convert leaf litter and fallen logs into soil nutrients, and in turn provide food for some species of small mammals and birds. In total, elimination of aspen woodland, the most diverse of the northern range subsystems, which at park establishment probably occupied on the order of 10% of the northern range area but significantly more of its diversity, has materially impoverished the northern range biotically. The result is a reduction in landscape, structural, and species diversity.

Influences on Upland System Structure II: The Sagebrush-Steppe Subsystem

7 The difficulty lies, not so much in developing new ideas, as in escaping from old ones. —John Maynard Keynes

INTRODUCTION The areas assigned by Houston (1982:86) to “upland steppe,” “shrub steppe,” and “wet meadows” combine to 43,000 ha, or 53% of the Park portion of the northern range. Grassland and sagebrush combine in “complex continua.” Smith et al. (1915) said the winter range “vegetation is generally characterized by sagebrush and its accompanying plants.” I am calling this combined type the sagebrush steppe. It is the most extensive vegetation type on the northern range and by far the major source of ungulate forage. The park portion of the northern range (refer to figure 1.1) slopes from 2,244 m elevation at the northeast park entrance westward to 1,899 m at the Mammoth headquarters, a highway distance of 47 km, then drops abruptly to 1,616 m in 5 km to the northern entrance at Gardiner, MT (Houston 1982). Thus 95% of the park winter range, and that much or more of ungulate winter forage production, is at elevations of 1,899 m and higher. Only 5% is located in the lower elevations at the northwest extreme of the range (figure 1.1), including 1,600 ha of land in Montana that were added to the park in 1932, and is called the Boundary Line Area, or BLA. In the pages that follow I refer to these different areas as the higher elevation and the Gardiner or lower-elevation ranges. The elevational gradient is accompanied by east–west gradients of increasing temperatures and declining 91

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precipitation (Houston 1982). Winters are more severe and vegetation growing conditions more favorable toward the eastern portions of the range. In the early weeks of winter elk are distributed widely across the east–west expanse of the winter range (Craighead et al. 1971; Houston 1982). As winter progresses and conditions become more severe, much of the population shifts westward down the elevational gradient to less severe conditions. But elk hunters just outside the park boundary at the northwesternmost extent of winter range within the park, of which the elk are aware, pose a barrier through most of the winter to easy northwestward movement down the Yellowstone River valley (figure 1.1). The barrier is not totally impermeable— increasing numbers of animals venture northward as far as Dome Mountain despite the risk. But the net effect of the hunting has been to cause animals to concentrate at the northwest end of the winter range inside the park. The result is uneven population distribution through much of the winter and consequent uneven distribution of browsing pressure on sagebrush. In response, the trends in sagebrush have differed between the higher and lower portions of the winter range. As with aspen, there are differences of opinion over the effects of elk on this type. Houston (1982:appendix VI) summarized numerous studies and measurement initiatives begun in the 1930s, but most were continued for only short periods of years. The congressional appropriation in 1986 to research the question “Is the northern range overgrazed?” stimulated a flurry of research, but this was also of short duration. As in the case of aspen, the research on the sagebrush steppe was begun over a half century after the park was established and nearly as long after a burgeoning elk herd had begun impacting the type. Hence park research began on what must have been a long-affected and probably altered system.

VEGETATION STRUCTURE Sagebrush Sagebrush Biology Although the sagebrush steppe is, like the aspen woodland, a system of interacting species, it is useful to understand the behavior of the components. The northern range sagebrush steppe is dominated by big sagebrush (Artemisia tridentata) (Wambolt and Sherwood 1999): mountain big sagebrush (A.t. vaseyana) occupying the higher elevation ranges, and therefore 95% of the total, and Wyoming big sagebrush in the lower elevation area. Description of big sagebrush ecological behavior may be colored at times by tendencies toward negative, value-laden stereotypes. The species has increased at times and in places in the West as a result of cattle grazing and at the expense of grasses. Hence it tends to be viewed negatively by the livestock industry as


inferior forage and by environmentalists, who look on it as a symptom of disturbance and land misuse. Houston (1982:368) recognized this in commenting on earlier range surveys based on “The assumptions (e.g., the presence of big sagebrush in grasslands was always interpreted as an unnatural retrogressed condition due to grazing)” and considered these assumptions “inappropriate for native ungulates in a national park.” And he comments elsewhere (p. 357): “A consideration of species abundance and composition suggests that vegetation sampled by these transects represents examples of climatic climax plant communities. This interpretation would apply to most of the sage communities on the area.” In fact, when protected from browsing, big sagebrush can become a dominant vegetation on the northern range (figure 7.1), indicating that it is a well-adapted, effective competitor in the intact, undisturbed vegetation of the area. It thus becomes a climax dominant. Moreover, it has high nutritional value for herbivores (Welch and McArthur 1979), and over much of the West is a winter dietary staple for mule deer (Odocoileus hemionus) and pronghorn antelope (Antilocapra americana).

Relationships between Elk Numbers and Sagebrush Abundance In the case of aspen, it was possible to derive quantitative estimates of the extent of its decline by extrapolating from the 1930 estimate and Barmore’s calculations of total aspen area in the

HISTORIC AND PHOTOGRAPHIC EVIDENCE.

Figure 7.1 South side of Mammoth exclosure showing inside-outside difference in sagebrush density and shrub stature. Photos taken July 16, 1997.

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1960s. A photographic record allowed projection backward from 1930. And the tree-aging studies made it possible to detect the sharp decline in tree establishment around the turn of the century. Although Houston (1982:86) reports the total area of “shrub steppe,” there are no range-wide estimates of the trends in areal abundance of sagebrush itself. Some inferences can be drawn from the photographic record and historic accounts, which give indications of the species’ status through park history. Some of the same authors who commented on aspen effects described impacts on sagebrush in the 1920s and 1930s. Rush (1932) stated on the basis of his 1928–29 studies: “All browse species are heavily overgrazed and will eventually disappear from the range unless improvement is shown in the next few years. Even the sagebrush on the Mammoth to Gardiner section is nearly all dead.” Wright et al. (1933): “The present available winter range shows unmistakable signs of overgrazing. [Based on surveys begun in the late 1920s.]” Wright and Thompson (1935): Carcasses, but not of elk; just sagebrush remains of the once luxuriantly sage-covered hills of northern Yellowstone. [Mount Everts, May 23, 1932, p. 111] Sagebrush graveyard. Sagebrush killed by overbrowsing. Characteristic of much of the Yellowstone elk winter range . . . September 17, 1933, near Black Tail Deer Creek. [Caption for their figure 27] See also figure 7.2. Grimm (1939) remarked on the basis of his research in the park: “increasing areas of dead sagebrush may be seen in section [Gardiner region] and it is obvious that the existence of this valuable shrub is threatened. [In a sample of 22 100-square foot plots, 65% of sagebrush plants were dead.]” In an effort to conduct a semi-quantitative, chronological evaluation of the photographic record, I examined 70 photographs with sagebrush vegetation, 33 higher-elevation and 37 lower-range shots. These include 2 photographs of Wright and Thompson (1935), 2 paired photographs of Houston (1982), and 12 repeat photographs of Meagher and Houston (1998), all for the higherelevation range (i.e., from Mammoth eastward). I examined 7 paired photographs of Kittams (1948), 3 paired Houston photos (1982), 1 pair of Phillips (1963), a single shot by Wright and Thompson (1935), and 4 repeat photographs of Meagher and Houston (1998), all for the lower range. I assigned a visual rating to the abundance of sagebrush in each photo: 3 if judged abundant, 2 for moderate, and 1 for scarce, absent, or severely impacted. These are summarized chronologically in figure 7.3 and partitioned separately for higher elevation and lower range. These are small samples and the ratings are subjective. But given these caveats, the 1871–1918 ratings are largely abundant to moderate, both in the higher elevation and lower ranges. This suggests that sagebrush was a major component of the sagebrush-steppe portion of the northern range in the early years of park history.


Figure 7.2 “The right side of the picture shows elk winter range within Yellowstone National Park. The fence is the boundary. The left side of the picture shows rank growth of sagebrush just outside the park. Here we can compare the original and the present state of the range.” Wright and Thompson’s (1935) figure 36, taken June 1, 1932, near Gardiner.

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Figure 7.3 Rankings of northern range sagebrush abundance as judged visually from 1871– 1991 photographs. Each number ranks one photo: 3 represents sagebrush as abundant; 2 moderately abundant; and 1 scarce, absent, or severely impacted. Numbers are placed on dates of photographs.

The ratings for 1922–60 are predominantly 1s, implying sharp reduction from the early conditions. This is also the period of alarmed reports, quoted previously, by park biologists expressing concern for the ungulate impacts on sagebrush and for the future of the species in the northern range. Houston’s and Meagher and Houston’s photos taken in the early 1970s indicate recovery to moderate or abundant levels at the higher elevations but no recovery and possibly further decline in the lower range. THE EXCLOSURE EVIDENCE. Park personnel constructed 8 2.1–ha exclosures across the northern range in 1957 and 1962. From east to west, these are: 2 in the Lamar River valley, 1 at Junction Butte, 2 above Blacktail Creek, and 1 at Mammoth Hot Springs, all in the higher-elevation ranges. Two were constructed near Gardiner in the lower-elevation ranges. Sagebrush vegetation was measured inside and outside these exclosures by park personnel with paired permanent belt transects in 1958 and 1962, 1967, 1974, 1981, 1986, and 1990, except that the 1988 fires swept through the Blacktail exclosures, and no 1990 measurements were made there.


These measurements have been analyzed and published by park personnel Singer and Renkin (1995), Norland and Reardon (1996), and Reardon (1996), aggregating the data in slightly different combinations of exclosures and years and reporting somewhat different vegetation parameters. Norland and Reardon’s and Reardon’s excellent reports provide the most continuous analyses. I have placed particular weight on Reardon’s (1996) tabulations. He combined all shrubs, but because sagebrush is by far the most abundant, it is reasonable to assume that his results portray sagebrush trends. The results need to be interpreted in two dimensions: (1) how the vegetation changed over time and whether this coincided with the changing trajectory of the northern herd; and (2) spatially how the sagebrush subjected to browsing outside the exclosures compared to that inside the refugia. All measurements showed low sagebrush canopy cover both inside and outside all exclosures in 1958–62 (figure 7.4). This was, of course, the point at which the exclosures had just been constructed, and the studies were measuring the same vegetation inside and outside. All of the park studies showed increase in sagebrush canopy cover between 1962 and 1990 inside both upper- and lower-elevation exclosures and to a lesser degree outside the upper-elevation ones (figure 7.4). The increases began between 1962 and 1967 and continued through 1990. Cover measurements outside the lower (Gardiner) exclosures declined continuously from 1958 to 1990. The trend in densities of sagebrush plants over time in the park studies are not as consistent as the cover trends. Singer and Renkin’s (1995) table 4 data show density decline both inside and outside the higher exclosures between 1958–62 and 1990, increase inside the lower ones, and decline on the outside of the latter. But their 1986–87 results in their table 6 cannot be compared with those in their table 4 because the scale is not shown in the latter. In both the Norland and Reardon and the Singer and Renkin results, densities in all browsed transects declined between 1958–62 and 1990 except that Norland and Reardon’s measurements show initial decline, then slight increase by 1990. Both studies showed density increase inside the Gardiner exclosures. However, where Norland and Reardon’s measurements showed essentially no change inside the Junction Butte and Blacktail exclosures but decrease at Mammoth, Singer and Renkin’s combined values for the higher-elevation exclosures showed decline. Similarly, both park studies showed densities inside the Gardiner exclosures in the last measurements in 1990 to be markedly higher than those exposed to browsing on the outside. Both studies found densities higher on the outside at the higher-elevation exclosures: Singer and Renkin’s combined values, and 3 out of 4 of Norland and Reardon’s individual sites. Singer and Renkin (1995) found shrub heights to be uniformly low in 1958–62 (means = 0.7–1.9 cm) both inside and outside all exclosures. They had increased in all areas by 1986–87 and 1990, but in all cases heights were greater inside the exclosures. In the most recent and extensive studies, Wambolt and Sherwood (1999) established a stratified sampling procedure at 19 “sites” in 1994. At each site

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Figure 7.4 Trends in shrub cover inside and outside exclosures. Data from Reardon (1996).

they matched 10 2 × 30 m belt transects on similar slopes, exposures, and edaphic conditions both inside and outside the exclosures. Thus total transects were 20 × 19 = 380, 220 at the higher elevations (Blacktail not included) and 160 at the lower. (Note: The number of park-established transects, each roughly similar in area [Singer and Renkin 1995] to Wambolt and Sherwood’s, was 16.) The latter authors measured canopy cover and estimated annual production per plant with a model (Wambolt et al. 1994) designed to make such calculations and density of plants with minimum canopy cover ≥15 cm. Sagebrush canopy cover at all 19 sites was higher inside the exclosures, with 18 significant at p < 0.01 and 1 at p < 0.03 (table 7.1). Similarly, densities of shrubs >15 cm were higher inside the exclosures than outside. Production per shrub was higher inside the higher-elevation exclosures than outside. The authors did not calculate production at the eight sites of the two Gardiner exclosures because the plants had been browsed to such small stature on the outside that their model could not estimate the values. The mean production per sagebrush plant for the higher-elevation sites, estimated here from Wambolt and Sherwood’s table 4, was 147.0 g inside the exclosures, 78.1 g outside, a 1.88× difference.


Table 7.1 Mean Percent Cover, Plant Density, and Production per Plant of Sagebrush at 11 Higher Elevation Sites (4 Exclosures) and 8 Gardiner Sites (2 Exclosures) from Wambolt and Sherwood (1999) Higher Elevationa

Mean % cover Mean shrubs/60 m2 Gm. Prod’n./Shrub

Gardinera

Insideb

Outsideb Insideb

Outsideb

24.4 41.4 147.0

10.7 25.7 78.1

0.8 1.1 —

7.8 15.5 —

aMeans

per site calculated from Wambolt and Sherwood’s (1999) tables 2, 3, and 4. bThese are shrubs with a minimum canopy of 15 cm.

The authors commented that the production per unit area could be calculated by multiplying per shrub production by the shrub densities. My calculation of the inside and outside mean densities of 41.4 and 25.7 shrubs/60 m2 (table 7.1) places the mean density difference at 1.61×. Thus per unit area sagebrush production inside the exclosures is 3.03 × that on the outside: 1.88 × 1.61 = 3.03. Because Wambolt and Sherwood could not estimate per plant production outside the Gardiner exclosures, similar calculations cannot be made. But with sagebrush shrub density outside the Gardiner exclosures at 1.1/60 m2 = 0.02 shrubs per m2, and shrubs so small that Wambolt and Sherwood’s model could not estimate production, sagebrush production is all but nonexistent as a significant source of ungulate winter forage (figure 7.5) in that portion of the winter range. Working out of Wambolt’s lab, Rens (2001) made sagebrush measurements in 1998 at the Blacktail exclosures to document recovery from the 1988 fires. These showed 1998 canopy cover inside the exclosures to be 2 to 3 times that of the outside measurements: Percentage cover, matched transects Inside Outside Exclosures Exclosures 16.3% 7.9% 20.3% 6.3% 24.8% 6.3% 18.1% 11.4% 20.5% 11.9% In 1993, Wambolt and Hoffman (2001) aged dendrochronologically 20 large (>22 cm across the widest portion of the crown) and 20 small (2,500 m) plateau that constitutes roughly 90% of the park area. The several conifer types that he lists (p. 86) for the northern range (see figure 1.1)—which is 9% of the total park area—aggregate to 33,700 ha or 41% of the park portion of the northern range. Conifers occur as mature forests on the north-facing slopes descending from the high plateau and on the tops and north faces of the small mountain ranges within the northern range. They occur in varying densities and abundance along some stream bottoms and as small thickets or even isolated trees on favorable sites in the open terrain of the northern range. Houston (1982) subdivided the northern range coniferous vegetation into four subtypes: • Douglas fir (Pseudotsuga menziesii), 16,100 ha • Lodgepole pine (Pinus contorta), 10,600 ha

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Influences on Upland System Structure III: Conifers and Deciduous Shrubs

• Engelmann spruce (Picea engelmannii)/subalpine fir (Abies lasiocarpa), 6,600 ha • Whitebark pine (Pinus albicaulis)—400 ha Since Houston wrote, Despain (1990) has identified whitebark pine as a high-elevation species, and the closely related relative occupying limited area in the northern range is now considered to be limber pine (Pinus flexilis). While referring in several places to Rocky Mountain juniper (Juniperus scopulorum) on the northern range, Houston never gave any estimate of the area it occupied. Mature coniferous vegetation is notoriously lacking in understory vegetation that could provide winter forage for ungulates. Yet younger trees within reach are browsed despite the strong chemical defenses that place them low on the palatability rankings. All of the species on the northern range show signs of browsing use. Barmore (1980) and Houston (1982:139) recorded numbers of elk seen in the major vegetation types as they traveled through the northern range. Approximately 10% of 59,784 feeding elk were recorded in the vegetation type “other,” which was made up largely of coniferous growth. They pointed out that the types, particularly “other,” were not sampled in proportion to their areas. Yet observation of only 10% of the animals in a type occupying 41% of the area suggests that the conifers are used far less than the other types proportionately. As with other components of the vegetation, early park biologists commented on severe browsing impacts on conifers. There is evidence of some reduction in conifers by elk use, but there is also evidence that the area of the northern range occupied by conifers has increased since park establishment. Because there has been no systematic sampling of the distribution and abundance of conifers, especially periodic measurements over time, it is impossible to determine what the net trend has been. However, some inferences can be drawn from the available evidence about the chronology of browsing pressure, and these can be related to the elk population trajectory. The result is another body of evidence that takes its place alongside the evidence for the other ecosystem components being analyzed herein, which together provide an extensive picture of elk impacts on the northern range ecosystem. The signature indication of browsing pressure on conifers is the browsing highline. Highlines on all coniferous species are as ubiquitous on the northern range today (figure 8.1) and as characteristic of its condition, as chewing-induced, bark scarring on aspen (refer to figure 6.8). The question of when conifer highlines appeared on the northern range bears on the debate over elk numbers in prehistory and early park years. As with other components of the ecosystem, there are different interpretations of the evidence and different sources of evidence cited. The only early evidence (i.e., prepark and first two decades after establishment) is the photographic record. Several factors kill the lower branches of conifers and therefore must be considered in interpreting plant structure in photographs. One is ungulate

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Figure 8.1 Browsing highline on Engelmann spruce in Yellowstone’s northern range. By virtue of its stiff, prickly needles and heavy defensive chemistry, spruce is considered to be at the bottom of the browsing palatability scale. Photograph by Charles E. Kay, July 15, 1986.

browsing. Another is low-intensity ground fires as Kay (1990:238) points out. Yet another is self-pruning. As a conifer stand closes in and begins shading the lower branches, they die. And in Yellowstone, the hot water and hydrogen sulfide releases from the thermal areas often kill surrounding vegetation, including coniferous foliage. In relatively open conifer stands in which self-pruning has not yet become a significant influence, the height separating live branches and those killed by ground fire tends to be variable and somewhat irregular. Highlines created by browsing tend to be uniformly parallel to the ground and at a consistent height of ~ 3 m in the case of elk (see figure 8.1). They are often most clearly shown in photographs of trees on the skylines of mountain tops and slopes. Highlines on slopes are parallel with the slopes reflecting the fact that animals browse at exactly the same vertical heights above where they are standing, irrespective of their position on the slopes.

Chronological Changes in Browsing Pressure Historical Evidence EARLY PARK YEARS, 1871–99. Houston (1982) and Cole (1983) concluded that they could see highlines on conifers in early (i.e., 1871–93) park photographs.


Kay (1990:237–48) scrutinized these along with about 50,000 pictures in several photographic archives and commented at length that there was not evidence of early highlining. I comment here only on the two scenes in which Houston inferred highlines—on scenes in Kittams (1948) that Kay did not cite and on those in Meagher and Houston (1998), published since Kay’s work, including scenes in which the authors concluded that there were indications of highlines. Houston (1982) opined that the trees in his plates 34 and 38, 1885 photographs by J. P. Iddings, showed highlining. Plate 34 is a scene near Tower Junction. The trees at the left of the photograph have branches to the ground. The tree on the left of the central three has suggestion of a highline, but the other two in the trio have branches to the ground. Moreover, as mentioned in chapter 7, this site appears to have been burned in the recent past as suggested by the absence of sagebrush, the thick bunches of young aspen in the center, and the dead tree trunks in the upper right. The appearance of the conifers contrasts sharply with those in Houston’s 1970 retake of the site, which are clearly highlined. Houston similarly inferred the presence of highlining in his plate 38. But here again the judgment is problematical. The two trees farthest on the right are somewhat suggestive. But the several trees in the left half of the photo appear to have branches to the ground. Once again the contrast is stark between these and the clearly highlined trees in Houston’s 1971 retake of the site. Meagher and Houston (1998:245) infer evidence of highlining from their plates 27, 34, 40, 52, 74, 80, 81, and 88. Their plate 74 is the same as Houston’s plate 38 just discussed and needs no further comment. Of the remaining seven locations, only plate 52 is a photo of the northern range, and hence the series has questionable relevance to the issue at hand. The first scenes in plates 27 and 88 were taken in 1900 and 1931, respectively, and therefore are not chronologically relevant at this point in the discussion. With those reservations, my inferences from this series of photographs are as follows. The single northern range photo, plate 52.1, was taken in 1885 by Iddings near Tower Junction. One tree in the middle ground may have some missing lower branches, but there is no clear highline. The bottom of the second tree is behind the ridge, but the portion that shows clearly has no highline. The only other trees, on this primarily open-ground photo with few trees on which to judge, are the small conifers in the distance of the left center. They appear to be intact and lacking any highline. From the remaining photos of the high-elevation ranges cited by these authors, irrespective of date, I conclude the following: 1. Plate 27.1, Madison Junction, 1900. The authors infer highlining on trees in the central background. But these are in a mature stand where selfpruning may be occurring. The effect is not clear on the trees in the right half of the photo. There clearly is no highlining on the open grown trees in the foreground. By 1900, the elk population had risen to extremely high levels. Once again, this is not northern range.

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2. Plate 34.1, Midway Geyser Basin, 1871. Not only do I not see evidence of highlining, I consider that the trees do not show such evidence. Here again, there is sharp contrast with the 1971 and 1990 retakes, which show classical highlining. The few gnarled plants in the upper center of the first photos are growing in a thermal area and doubtless struggling with their environment. 3. Plate 80.1, Hayden Valley, 1881. I do not see convincing evidence of highlining, and the authors themselves comment, “In 1881 those trees present appear to have had denser basal branches.” 4. Plate 81.1, Hayden Valley, 1871. Again a thermal area, with equivocal evidence of highlining. The authors attribute what they infer to be highlining to bison rubbing, not elk. 5. Plate 88.1, again a 1931 photo and not relevant at this point in the discussion. 6. Plate 40.1, Upper Geyser Basin, 1885. This is the one early photo with definite appearance of highlining. It is a curious spot, an open area with a hotel that is obviously a tourist attraction. One wonders if the browsing might have been from horses tied or grazed in the area. A nearby similar site, the Old Faithful area in the following plate (41.1), shows numerous young conifers in the center of the picture that are clearly not highlined. Once again, these are not northern range areas. These authors searched early photographs for evidence of the existence of highlining. As commented, most of the scenes were not of the northern range and they were somewhat intermixed chronologically. Moreover, there was no comparable effort to determine whether there were indications of the absence of highlining. I have attempted what I suggest is a more objective, systematic, and somewhat quantitative approach. First, I have restricted my review to northern range scenes in Kittams’ (1948), Houston’s (1982), and Meagher and Houston’s (1998) photo sets. Kittams’ and Houston’s are all of the northern range. I considered Meagher and Houston’s northern range scenes to be plates 1–8, 10–11, 44–62, 67, and 72–74 based on their maps and those largely below 2,134 m (7,000 feet). I scanned the photos and judged whether a significant number of conifers in each were (a) highlined, (b) the evidence was uncertain, (c) apparently not highlined, and (d) too far in the background or positioned in the landscape in ways that did not provide a basis for judgment. I grouped them in four chronological blocks: prior to 1900, 1900 to the 1940s (Kittams’s retakes were largely in 1948), 1970–73 (the years of Houston retakes), and 1990–92 (years of retakes by Houston and Meagher). I summed the numbers in each browsing category in each chronological block to attempt a provisional quantitative indication (the numbers are small), their change over time, and their prevalence across the northern range. I list my conclusions on the pre-1900 scenes next and then sequentially the other chronological blocks.


The only plates in Houston’s (1982) photo set of the early northern range with conifers of sufficient size and positioning to allow reasonably dependable judgment are his plates 34.1 and 38.1. I discussed these already and concluded there was no convincing evidence of highlining. In the Meagher and Houston set, plates 4.1, 8.1, 45.1–49.1, 51.1, 54.1– 57.1, and 67.1 were all taken in 1900 or later. Plates 1.1–3.1, 5.1–7.1, and 58.1– 61.1 do not have enough trees large enough or positioned in ways to allow judgment. My inferences on the remaining plates are as follows. 1. Plate 10.1, 1885. The trees on the left are in a mature, dense conifer stand where some self-pruning may be occurring. The younger, open grown trees on the right are clearly not highlined. 2. Plate 11.1, 1885. This is a somewhat hazy print, but the trees in the foreground and upper right do not appear to be highlined. 3. Plate 44.1, 1885. Cannot judge for most of the photo, except possibly for the trees in the open space on the lower right, which do not appear to be highlined. 4. Plate 50.1, 1885. Difficult to judge, but there does not appear to be any evidence of highlining. 5. Plates 52.1 and 53.1, 1885. Discussed already. No clear evidence of highlining. 6. Plate 62.1, 1885. Some evidence of self-pruning in the mature stand at left center, but trees in lower center and lower right clearly have branches to the ground. 7. Plate 72.1, 1885. No highlines, young conifers with branches to the ground. 8. Plate 73.1, 1871. No evidence of highlining. 9. Plate 74.1, 1885. Same as Houston’s Plate 38 discussed above. Kittams (1948) provides two pre-1900 photos that show conifers that permit inferences about highlining. 1. Figure 7, 1887. These are Douglas firs on a hillside with no clear highline, except possibly for one small tree left of center. 2. Figure 21, 1887. The taller Douglas fir at left center has probably selfpruned. The rotund younger tree at left could be experiencing the beginning of browsing on the lower branches. The tree at right appears unbrowsed. In sum, this is a sample of 13 scenes, widely distributed over the northern range and none later than 1887. None shows convincing evidence of pronounced highlining, if any at all, especially in contrast to the scenes discussed next. With one exception, I have concluded the same on those high-elevation scenes already discussed, in which Meagher and Houston infer highlining. The most dramatic photograph bearing on the highline question is an 1871 photograph, the year before the park was established, by W. H. Jackson of the Mammoth Terraces (shown here as figure 8.2a). With branches virtually covering

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Figure 8.2 Coniferous vegetation below Mammoth Terraces in the heart of the lower winter range. (top) 1871 photo by W. H. Jackson. Species are probably Rocky Mountain juniper and limber pine, the surviving species today. Note dense, unbrowsed growth down to the ground. (bottom) Same area showing highlined junipers as photographed by Charles E. Kay, August 26, 1989. 130


the surrounding ground, the trees (judging by the species present today) are probably Rocky Mountain juniper and limber pine. This area is in the heart of the lower northern range. Reports of highlines on conifers by park investigators began at about the same time as reported impacts on other components of the vegetation. Smith et al. (1915), on the basis of their 1914 survey, commented “In places also the twigs and small branches of Douglas spruce [sic], Engelmann spruce, Alpine fir, and lodgepole pine are also trimmed up as high as the elk can reach during deep snows.” Skinner (1927) showed a 1920 photograph (his figure 44) of conifer highlines, and commented (1928), “Elk even stood on their hind legs to reach and pull down the lowest branches of limber pines . . . red cedar . . . and cottonwoods.” Wright and Thompson (1935) showed four 1932 photographs (their figures 31, 32, 33, 36) of highlined junipers (see figure 8.3) and Douglas firs near Mammoth and the Gardner River. They commented that the seedlings were browsed and hence the species could not replace themselves. Shafer (2001) shows a Ben Thompson 1932 photo of a highlined Douglas fir near Mammoth Springs.” Grimm (1939) commented on juniper highlines, as did Cahalane (1943), who remarked, “However, much of the edible portions of shrubs and trees, especially of fir, juniper . . . were already stripped as high as elk could reach . . . Reproduction of these species . . . became seriously inhibited.” Commenting on a 1916 photograph, Kittams (1948:29) stated: “A prominent browse line on trees . . . indicates heavy browse use prior to 1916. . . . Of

1900–48.

Figure 8.3 Browsing highline in Rocky Mountain juniper on the northern range in 1932. This is figure 36 in Wright and Thompson (1935).

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significance is the lack of reproduction. With juniper . . . lack of successful reproduction over a period of several decades is significant.” Meagher and Houston (1998) show 11 northern range scenes from 1900 to 1929. I cannot judge 6 of the scenes (4.1, 8.1, 45.1, 46.5, 55.1, and 67.1). Of the others, 2 do not appear to show highlining (48.1, 54.1)and 3 do (49.1, 56.1, 57.1). Among Kittams’s (1948) scenes, I am unable to judge 7 (figures 9, 11, 15, 25, 35, 37, 39) of the early shots before his 1948 retakes; of those on which I can draw inference, 8 appear to show highlining (1, 5, 13, 23, 27, 29, 31, 33) and 2 do not (17, 19). Of his 1948 retakes, all 16 on which judgments can be made show evidence of highlining, and no judgment can be made on 4. Thus highlining on the northern range clearly appeared by the 1910s–1920s. The sample is small for the upper northern range and not amenable to judging prevalence. But Kittams’s series implies that it was the general condition in conifers by no later than the 1920s and certainly present earlier on the lower northern range. By Kittams’s 1948 retakes, again largely on the lower northern range, it was essentially universal. 1970S–1990S. Among Meagher and Houston’s (1998) 1970–72 retakes, I could not judge on 18 of the northern range shots whether there was or was not evidence of highlining. But in 12 (1.2, 3.2, 10.2, 45.2, 49.2, 53.2, 54.2, 56.2, 57.2, 72.2, 75.2, 74.2) I did infer evidence of highlining, and in 4 (8.2, 11.2, 44.1, 62.2) I could not detect such evidence. Thus in 16 photos, largely of the upper northern range, highlining appeared to be present in 75% of the total. In these authors’ 1990–92 retakes, the numbers were approximately the same. I formed conclusions again on 16; 13 showed evidence of highlining, and 3 did not. CONCLUSIONS. I have examined 13 pre-1900 northern range photographs that had images of conifers that allowed judgment. None had persuasive evidence of highlining anything like that shown in later photographs and characteristic of northern range conifers today. The most probable inference is a low-density elk population, which in addition did not winter in the area for decades prior to the photography. Of 5 1900–29 northern range scenes in Meagher and Houston, 2 do not appear to show highlining, and 3 do. The 2 that do not were taken in 1900 and 1918. Of 10 Kittams photos in the same decades, 8 appear to show highlining and 2 do not. The latter 2 were taken in 1927. Of his 16 retakes in 1948 that allow judgment, all show highlining. Of 16 photos taken each in the early 1970s (Houston) and early 1990s (Meagher and Houston), the proportion showing highlining is approximately 75%. Thus the evidence of highlining appeared in the early 1900s along with the rise in the elk population. The high percentage of later photos showing highlining indicates that it became a general pattern across the northern range. If there is any bias, it is toward conservative estimates because it is more likely that the condition would be missed in photographs than appearing to be present when


it was not. There is some indication that the condition became common on the lower northern range earlier than on the upper range. There is no basis for estimating the percentage of trees that are highlined. Subjectively, one gets the impression that it is present in most with branches low enough to be browsed. Highlining is present in all 6 coniferous species on the northern range. But there is no way of knowing at this point whether the condition appeared earlier in the more palatable species (e.g., junipers) and progressed down the palatability scale to the least palatable (e.g., Engelmann spruce) or whether it appeared simultaneously on all species. It evidently appeared at about the same time heavy browsing impacts were reported on palatable deciduous species, such as aspen (chapter 6) and willow (chapter 10). Although Houston (1982:129) questioned the observations of earlier biologists that browsing was preventing reproductive replacement, there does appear to be evidence that it has occurred.

Lessons from Plant Architecture A highlined tree is a plant architectural type that conveys a history of the browsing pressures to which it has been subjected during its lifetime. Keigley (1997a; Keigley and Frisina 1998) has devised a method for dating the browsing intensity and variations therein that a plant experiences. I elaborate his methodology fully in chapter 10, but I use only a part of his procedure next in drawing some chronological inferences from the photographic record of conifers. The full procedure requires dendrochronological aging of the plants in question, and this is obviously not possible for photographs. But limited inferences can be drawn from photographs with partial use of the methodology, and these supplement the chronological inferences already drawn on the appearance of highlining. The discussion that follows is keyed to figure 8.4. Four age classes of plants can be identified in this small cluster of conifers, in descending order of age: (a) the oldest, tallest and largest tree on the right; (b) 3 intermediate-aged trees: 1 in the center and 2 to the left of the oldest tree; (c) 5 young trees: 2 in the center, 2 in front of, and 1 to the right of the oldest tree; and (d) a number of short, hedged plants across the center and in the foreground of the photo. There are some browsed willows in the picture, but these are not under consideration here. Tree age classes a through c are all highlined at the same height, which is the height to which elk can browse. Some of the plants in age class d are dead, but some have a small amount of living foliage near the ground. That tissue is doubtless protected from browsing by winter snow. The space between the tops of these plants and the highline is the browse zone, between approximately 0.3– 0.4 m and ~3 m. No conifer foliage grows and survives in the browse zone, and it must follow that the tops of age classes a, b, and c had grown above the top of the zone before browsing reached sufficient intensity to remove all of the foliage within the zone. Thus the plants in these age classes have experienced at least two browsing intensities during their lives: an early absence or light intensity that enabled

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Figure 8.4 Browsing effects on four age classes of Engelmann spruce in Yellowstone’s northern range. See text for interpretation of these plant architectures in reconstructing browsing chronology. Photograph taken July 17, 1997.

them to grow through and develop foliage above the zone, and a recent heavy intensity that removed their foliage within the zone. Some of the plants (e.g., the two bare, apparently dead stems at left center) in age class d also apparently experienced two browsing intensities. They began growth and attained some height during an early period of light or no browsing. But they never reached sufficient height to escape the browse zone. When browsing began and intensified, their central stems were fully exposed, and they were browsed back to the snow-protection level. The youngest plants in age class d were never able to extend significantly above the snowline. Hence these plants must have lived throughout a recent period of heavy browsing. In sum, these architectural interpretations of the plants in figure 8.4 convey the same story as the photographic record: an earlier period of light or no browsing followed by a period of heavy browsing. The older plants (age classes a–c) have lived through both periods and are now continuing height growth and lateral growth above the highline. A small number, perhaps only two, of the age class d plants (those with the dead central stems) also experienced the two periods. But they were never able to grow to sufficient height to escape the browse zone before the period of heavy browsing set in. The youngest plants in age class d have lived only in the recent period of heavy browsing. Some of the hedged plants are dead, probably because the continued removal of photosynthetic tissue prevented support of stem and root metabolism.


General Synthesis on the Coniferous Vegetation The principles I have deduced from figure 8.4, based on Keigley’s methodology, must apply generally to highlined plants, including the larger trees of the northern range. Such plants must have begun life and grown up beyond maximum potential browsing heights during periods of light or no browsing. Highlines were produced during a later period of intensive browsing that removed the lower foliage as high as the animals could reach. These deductions are consistent with the chronology of highline evidence inferred from the photographic record. Both sources of evidence are consistent with the conclusions of chapters 2–5 that the northern herd existed at low densities, and migrated out of what is now the northern range during winter, prior to park establishment and its early decades. That low population was followed by sharp increase and high numbers from the 1890s through about the 1950s. An extensive systematic sampling of conifers across the northern range with Keigley’s full methodology, including coring near the ground and above the highline, could answer a number of questions raised by scrutiny of the photographic record. The methodology could provide insights into the spatial and chronological development of highlining. Keigley (personal communication, 1994) did core 12–14 released-type Douglas firs on the northern range. He found that the vertical stems were released between 1963–75. This was the period during which the northern herd was at low levels following reductions and is another source of evidence—along with that on aspen and shrub-steppe already cited, and on other system components to be discussed in later chapters—on system response during the short period of herd reduction. Keigley’s conclusions are based on a small sample, and the inferences are tentative, of course. The question could be addressed more adequately with an extensive, well-designed sampling effort. Additional questions arise from the photographic record that could be answered with an adequate sampling scheme. Meagher and Houston’s (1998) plates 53.1 and 72.1 are 1885 northern range scenes. Houston’s retakes of these scenes (plates 53.2 and 72.2) are dated 1970 and 1973 and show new conifers, generally highlined, that were not evident as young trees in the original photos. These new trees presumably became established and grew through the browse zone somewhere between 1885 and the second decade of the 1900s when highlining was becoming widespread, a 25- to 30-year period. Coring the trees would test this presumption. More generally, Houston (1982) infers convincingly from his early 1970s photographs that conifer stands increased in density between the dates of the early photos and his retakes and spread into previously open areas. He concludes (p. 91) that forest cover increased primarily on the north slopes, colonized burns, and invaded bunchgrass communities and aspen. The increase was less extensive on south slopes and valley bottoms, Houston estimating (p. 92) that it occurred over about 5% of the area. (This increase occurred during a period of slightly increased temperatures and decreased precipitation to which he and other park biologists have attributed the decline of other vegetation components.)

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An extensive sampling effort would provide an understanding of the chronology and spatial extent of these extensions. Did they occur largely prior to the rise of the elk herd and has browsing pressure largely stopped the spread, as might be inferred from figure 8.4? The Meagher and Houston (1998) photos of the early 1990s suggest that there was not a significant further advance after the shots taken 20 years earlier (see Meagher and Houston’s plates 7.3, 10.3, 11.3, 44.3, 45.3, 47.3, 49.3, 50.3, 52.3, 53.3, 55.3, 56.3, 57.3, 58.3, 72.3, 73.3, and 74.3) although the trees in the earlier scenes matured. Is browsing now an impediment to conifer reproduction, as it is to aspen (chapter 6)? And when the present conifers die of old age, will this vegetation type change to shrub steppe, as is occurring with aspen? If they did change, and browsing pressures continued at current and recent levels, the result would be part of a massive stepwise alteration of the northern range vegetation structure. For all intents and purposes, willow and other riparian woody species have already approached elimination (chapter 10). Aspen has decreased sharply, and the remaining trees are at the limits of their longevity. Most can be expected to die out in the next decade or two (refer to figure 6.6). Conifers are longer lived, but without reproductive replacement they, too, could disappear as a type, at least in portions of the northern range heavily browsed. This could be expected to take a century or two, but could be accelerated by fire. Although the 1988 fires released massive reproductive replacement of lodgepole pine in the park’s higher elevations, some northern range areas burned by the fire show no replacement (figure 8.5). Will wolf effects change these prospects? Thus, it is unfortunate that almost no research effort has been directed to this component that occupies 41% of the area of the northern range. Numerous broader ecological questions arise. The pre-1970 increase in coniferous forest in all likelihood reduced herbaceous forage production for ungulates but may have fostered increase in avian and mammalian boreal species. Inability to answer these questions leaves a major gap in our understanding of the northern range ecology. Keigley, park biologist before his transfer to the Biological Resources Division of the U.S. Geological Survey in 1993, requested approval for an extensive study of the northern range woody vegetation, including the conifers. That approval was not granted.

DECIDUOUS SHRUBS AND SMALL TREES Introduction The northern range vegetation contains, depending on location, over 20 broadleaved shrubs and small trees. They grow on a variety of sites—woodland understory, riparian zones, and open, semi-arid situations. They do not generally assume dominance in the vegetation, growing as subsidiary species within the


Figure 8.5 Engelmann spruce stand on the northern range west of Tower Junction burned by the 1988 fires. With no replacement by young trees in the understory, this stand may convert to shrub steppe.

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major vegetation types. A number produce berries used for food by birds and both species of bears. A partial list of berry-producing species is: • • • • • • • •

Amelanchier alnifolia, serviceberry Cornus stolonifera, red-osier dogwood Lonicera involucrata, black twinberry Prunus virginiana, chokecherry Ribes montigenum, mountain gooseberry Rosa woodsii, wild rose Shepherdia canadensis, buffaloberry Symphoricarpos albus, common snowberry

Other common species of shrubs or small trees include: • • • • •

Acer glabrum, Rocky Mountain maple Alnus incana, mountain alder Betula occidentalis, river birch Purshia tridentata, bitterbrush Rhus trilobata, smooth sumac

Most of these species are highly palatable winter forage for browsing ungulates. Early reports suggest that they were common in the park area, as they are today in the region outside the park. As with other components of the vegetation, park biologists were reporting heavy browsing impacts on these by the 1920s. Very little research has been conducted on these species in the park. But the limited information available indicates what is now becoming a repeating pattern among all components of the northern range ecosystem.

Chronological Changes Kay (1990, 1995) searched early historical accounts of the vegetation in the YNP area. He cites an early account by Rev. Samuel Parker who traveled through Jackson Hole, Wyoming in 1835 and found an abundance of serviceberry and other shrub fruits along the Hoback River south of Yellowstone. Haines (1965) cited the journal of the 1869 Cook-Folsom-Peterson expedition, which encountered Native Americans who were gathering and drying large amounts of chokecherries on Tom Miner Creek just north of what is now the park. Haines quotes the journal: “Here we found a wickiup inhabited by two old squaws who were engaged in gathering and drying chokecherries . . . they had two or three bushels drying in the sun.” Kay (1995) cites the account of the Washburn expedition of 1870, which reported that near what is now YNP: “we crossed a small stream bordered by black cherry trees, many of the smaller ones broken down by bears, of which animal we found many signs.” Haines (1977:104) describes Norris’s early visit to the park area accompanied by Frederic Bottler who, 6 months earlier, had been mauled by a family of grizzly bears with which he had disputed “possession of a berry patch.” Else-


where (p. 111) he describes Truman Everts, a member of the 1870 Washburn expedition, falling ill from having “gorged himself on wayside berries.” Schullery and Whittlesey (1992:1–59) quote W. R. Raymond’s description of his 1871 trip to the park area in which he comments that “bears frequent the service-berry thickets.” The first accounts of heavy browsing impacts on these species were by Smith et al. (1915): “Mountain maple, service berry, wild rose, snow-berry, fly honeysuckle, and many other shrubs are eaten during the winter. If there is an abundance of food merely the buds and tips of twigs are taken, but in case of scarcity the stems are eaten down to mere stumps.” Rush (1932:65) observed during his 1928–29 studies that “all browse species are heavily overgrazed and will eventually disappear from the range unless improvement is shown in the next few years.” Wright and Thompson (1935) commented that Purshia is being killed out and in the caption of their figure 30 remarked, “Every available form of browse is reduced to a small percentage of its normal productive capacity.” As with other components of the vegetation, these species were a prominent part of the system before and at the time of park establishment in 1872. But by 1914 they were showing signs of heavy browsing, and this condition became more pronounced in the following decades.

Exclosure Evidence Kay (1990, 1995) measured the stature and berry production of three shrub species—serviceberry, chokecherry, and buffaloberry—inside and outside five exclosures in the Greater Yellowstone Ecosystem: Mammoth and Lamar West in YNP, Uhl Hill in Teton National Park south of YNP, East Elk Refuge in the National Elk Refuge just outside of Jackson (Wyoming), and Camp Creek near the confluence of the Hoback and Snake Rivers south of the city of Jackson. All areas are subject to heavy elk use. Canopy cover of each species varied between 6.5% and 18.5% inside the exclosures, between 0.4% and 2.0% on the outside. In 6 comparisons, the differences were statistically significant at p < 0.01 in 4, not different in 2. Mean shrub heights ranged between 86.6 cm and 281.9 cm inside, between 19.8 cm and 124.5 cm outside. All eight comparisons were different at p < 0.01. Kay generalized that “nearly all” persisting serviceberry and chokecherry plants on the YNP northern range are less than 0.3 m tall. Stands with numbers of plants are maintained in stunted form (figure 8.6) similar to the stature of shrub aspen (figure 6.1). Kay measured the number of berries produced per plant among the three species inside and outside the five exclosures. In four of the exclosures, mean berry production per plant was (Kay 1990:184): • Buffaloberry: 1,191.46 inside, 2.5 outside • Serviceberry: 104.68–1,333.07 inside, 0–0.07 outside • Chokecherry: 65.08–2,121.78 inside, 0 in all outside measurements

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Figure 8.6 A chokecherry stand near Tower Junction maintained in hedged condition by annual browsing and unable to produce fruit. The species normally grows to small tree stature. Compare with the similar growth form of aspen in figure 6.1. Photograph by Charles E. Kay, August 7, 1986.

The East Elk Refuge inside measurements were high like the others: 1,300.20 for serviceberry, 1,517.00 for chokecherry. The outside measurements, although two orders of magnitude lower than the inside measurements, were slightly higher than the outside measurements at the other exclosures (13.20 and 33.40 for serviceberry and chokecherry). Kay attributed these higher measurements to the fact that the terrain around this exclosure induced formation of a snowbank, which provided the plants a measure of protection from browsing. Hence, although they were browsed, the impact was not as extreme as at the other sites. All 8 inside-outside differences were statistically significant at p < 0.01. Kay pointed out that the stems flower and bear fruit on the growth segment preceding the current year’s growth. With each year’s new stems browsed off, no segments can mature to the second year of growth, flower, and bear fruit. He concluded that heavy hedging plus inability to reproduce explained the decline and scarcity of these species in the northern range. Thus there is a clear parallel with aspen and conifers. Prior to park establishment and in the latter 1800s, these species were a significant component of the northern range, and they produced a sizable berry crop for avian and mammalian inhabitants of the region, including humans. During the 1900s, a greatly increased elk herd for all intents and purposes eliminated this constituent of the system and a valuable food source for the fauna. The result is another aspect of reduced biodiversity.

Influences on Upland System Structure IV: The Ungulate Guild

9 Your elk herd is headed for extinction, and so also are many other forms of wildlife that are affected adversely by the elks [sic] overbrowsing of the grass ranges. —A. Starker Leopold

INTRODUCTION As large, highly mobile herbivores with a broad feeding niche that includes both woody and herbaceous vegetation, elk are potentially formidable competitors with other herbivorous species. That potential is magnified by their tendency to increase to high densities. And the competition potential is not confined only to herbivorous species through food competition. The removal of habitat can affect species in other trophic levels that use the vegetation. We have seen examples of this in both the aspen and sagebrush-steppe subsystems and will see additional cases in the riparian. Interspecific competition needs to be defined here to ensure common understanding. It occurs when two species use the same resource. If, in using it, one or both reduce its abundance to the point where one or both species cannot fulfill its (their) needs, and the shortage restrains their normal biological function below what it would be in the absence of the other species, competition occurs. Restraint occurs at the level of the individual— nutritional or habitat shortage—but is of greatest ecological interest when expressed at the levels of populations, communities, and ecosystems. Interspecific competition can be asymmetric, affecting one species more than another. It can involve more than two species. The effect on any given species can operate over a range of intensities from slight to extreme reduction in its 141

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mean, steady-state population level, to complete exclusion, with a resulting range of effects on community structure, biodiversity, and ecosystem function. Competition may result from long-term change in abundance of a resource. Two species may progressively reduce a resource over a period of time, gradually intensifying the shortage, increasing the restraint, and affecting their abundances. Or it may occur repeatedly in time without progressive depletion of the continuing productivity of a resource. This would occur if the competitors consumed current, annual vegetative production for each growing season to the point of shortage without progressive reduction of a plant population or community and its annual production. In either case, the competition and effects on the competitors take place in immediate interaction within the current time scale. The effects can fluctuate in time with variations in the abundance of a resource, as distinguished from relatively constant effects (see Wiens 1977). Examples would be year-to-year variations in vegetative production associated with variations in weather or variations in the availability of a resource caused by varying snow depths. The result would be temporal variations around some longterm mean effect. Competition can be difficult to demonstrate conclusively. Needed evidence includes: 1. Evidence of mutual use of a resource by two or more species. 2. Evidence that the resource is depleted by the use. 3. Measures of the impact of each species on the other in terms of the shortage created for each species by the other. Measures of resource-use overlap alone are not sufficient and must be transformed into impact measures that include per capita use of the resource and aggregate use by a population. The proportion of a given resource in a species’ diet alone is not a valid indicator of its impact, as discussed in chapter 7. A given plant species may be a minor dietary fraction. But if the consumer is a large animal, and that small fraction constitutes a significant absolute per capita amount, and furthermore if there are large numbers of the animals, the aggregate consumption can result in significant depletion and impact on another species. Classical competition models equate the impact of one species on another to the product of the first species’ per capita impact coefficient and the number of animals in its population. 4. Evidence of a negative effect on the biological function, preferably demographic, of one or more of the putative competitors. Experimental evidence is always desirable, if not from advertent experiment, at least from “natural” experiments. Thus, as Pianka (1981) points out, the population response of one putative competitor to changes in the numbers of another can serve as such a quasi-experiment. Two species

Influences on Upland System Structure IV: The Ungulate Guild

may use the same resource, but if in using it there is no effect on either’s biological function, there is no competition. The purpose of this chapter is to ascertain whether there is evidence of competition during park history between elk and the other upland ungulates using the same northern range vegetation. Houston (1982) posed competitive exclusion of sympatric herbivores as one falsification criterion for the natural-regulation hypothesis. But this is too stringent. The populations of other species could be suppressed without being excluded, and that suppression would constitute competition, in the process significantly altering the northern range ecosystem. In the next chapter I will examine evidence for riparian ungulate species, which along with the upland forms, constitute a seven-species guild using the northern range.

ELK INTRASPECIFIC COMPETITION: HARBINGER OF EFFECTS ON OTHER SPECIES Park publications have stated that “None of the traditional measures of animal fitness, including winter fat, winter mortality of prime-age animals, birth weights, or antler size, indicate any nutritional problem in the Northern Yellowstone ungulate herds.” (Anonymous 1992). Aside from the inconsistency of such statements with the basic tenet of the natural-regulation hypothesis—that limitation of the herd would in part be achieved through density-dependent constraint on population growth through intraspecific competition for forage, and the fact that such constraint would have to occur through effects on the animals’ nutritional plane transmitted to birth rates and death rates—there is clear evidence to the contrary. It should be clear from the last 3 chapters that the northern herd has progressively and profoundly reduced the woody vegetation on which it browses. And I have not yet discussed the riparian vegetation. This depletion is probably one reason for the small fraction of browse in the elk diet (Singer and Renkin 1995; Coughenour and Singer 1996; DelGiudice et al. 2001a). The evidence reviewed in chapter 7 pointed to significant depletion of the herbaceous vegetation by the time of the herd reductions. It apparently recovered to some degree during the 1960s and 1970s but may have started to decline again after 1981 (see figure 7.8). In the absence of herbaceous vegetation measurements at or before 1900, it is not possible to state whether the recovery of the 1960s and 1970s returned the range to anywhere near pristine conditions. The nutritional evidence is more conclusive, thanks to the excellent research of G. D. DelGiudice and co-workers. DelGiudice et al. (1991a) collected femur marrow and urine excreted onto the snow near 4 calves, 5 cows, and 1 bull that had died shortly before the collection period of January 27–February 16, 1989. All 10 elk had femur marrow fat 2,200 m) than in “height-suppressed” willows at lower elevations (2 mm in diameter constituting the remainder. The average rock-fragment cover for the entire data set is 20%. However, the percentage cover of surface rock fragments is not unequivocal evidence of sheet erosion. In part, the fragments on the surface come from the parent material from which the soil is formed. Concentration at the surface could occur either from sheet erosion or from frost heaving of parent-material fragments in the soil profile. To distinguish between the two, it would be necessary to compare number or weight of rock fragments per unit of soil volume at varying depths in the soil profile with their abundance at the surface. No such comparison has been made. Thus the prevalence of rock fragments on the surface of northern range soils may or may not indicate sheet erosion.

Experimental Evidence Lane (1990, summarized in Lane and Montagne 1996) conducted an experiment designed to determine whether elk use affects erosion. He compared soil chemistry, bulk density, and infiltration rate inside and outside the eight large exclosures described in previous chapters. In a simulated rainfall experiment, the authors also compared surface runoff and sediment yield inside and outside five of the exclosures.

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The authors concluded that chemical properties and nutrient levels were not widely affected by grazing versus nongrazing. However, differences in physical properties were found. Bulk density of the surface 5 cm was greater outside 7 of the 8 exclosures, although differences were statistically significant (p < 0.10, n = 4) at only 4 of the sites. At 6 sites, infiltration was greater inside the exclosures than outside, but the differences were statistically significant (p < 0.10, n = 3) at only 1 site. At the other 2 sites, infiltration was greater outside the exclosure, but differences were not significant (p > 0.10). Percent clay differed inside versus outside the exclosures (Lane 1990), thereby confounding the ability to discriminate between grazed versus ungrazed conditions. Lane (1990) conducted a simulated rainfall experiment that provides the only direct evidence of grazing effects on erosion rates. The experiment employed three treatments: (1) vegetation was left intact, (2) vegetation was clipped and clippings were left on the ground, and (3) the vegetation was clipped and clippings were removed. Each treatment was repeated 3 times inside and 3 times outside each of 5 exclosures, and both surface runoff and sediment yield were measured following simulated rainfall. Lane’s (1990) treatments simulated different seasonal conditions. During summer, elk typically move from the winter range in northern YNP to summer range in central YNP. Because of the lower numbers of elk, grasses on the winter range are largely ungrazed during the summer months. As elk move onto the winter range during fall and winter, the grasses are grazed. By May, winter range grasses are cropped close to ground level. Given this pattern of elk use, Treatment 1 mimicked conditions during the growing season. Treatment 3 mimicked conditions during spring, just prior to the growing season (and during the period of maximum expected runoff). Lane (1990) found that sediment yields were greater outside than inside two exclosures (p = 0.10); both cases involved Treatment 3. In 11 of the remaining cases, mean sediment yield was greater outside the exclosures, but differences were not significant at the p = 0.10 level. In two cases, sediment yield was lower outside the exclosures; again, the differences were not significant. Of the factors considered by Lane (1990), surface runoff and sediment yield provide the most direct tests of a possible relationship between elk and erosion. In most instances, Lane failed to detect a statistically significant difference in runoff or sediment yield between grazed and ungrazed sites. However, this may reflect low statistical power or use of inappropriate statistical methods rather than the absence of actual effects. In all cases, Lane used t-tests to evaluate the null hypothesis of no difference between the grazed and ungrazed conditions. These tests were employed on a site-by-site basis and can be expected to exhibit low statistical power due to minimal replication (n = 3 each for the grazed and ungrazed treatments at each site). More important, the relevant question is whether grazing generally influences surface runoff or sediment yields over the northern range as a whole. Use of site-by-site t-tests is inappropriate for testing an overall effect among multiple sites (see Zar 1984:162).

Influences on Ecosystem Function I: Erosion

We used randomization tests (Manly 1991) to reevaluate Lane’s (1990) data, which he provided, testing 2 1-tailed null hypotheses for the northern winter range as a whole: (1) mean surface runoff in the presence of grazing is less than or equal to mean surface runoff in the absence of grazing, and (2) mean sediment yield in the presence of grazing is less than or equal to mean sediment yield in the absence of grazing. For these tests, we assumed Lane’s study sites to be representative of conditions on the northern range. We ran tests separately for each of Lane’s three treatments. Two distinctly different approaches to randomization testing were used with the analyses conducted by K. A. Keating (USGS, Northern Rocky Mountain Science Center, Bozeman, MT). First because pronounced differences in background values for runoff and sediment yield were evident among sites, a stratified randomization procedure was used to control for intersite differences. In this test, mean runoff and mean sediment yield were first calculated, by treatment and grazing status, over all exclosures. Overall differences, D, between grazed and ungrazed means were then determined as 5

D=

RE

∑∑ X

5

G,E,I

E=1 I =1

5

RE

− ∑ ∑ XU,E,I E=1 I =1 R

∑∑i E=1 I =1

where E is the particular site (i.e., exclosure), i indexes replicates from the site, RE is the total number of replicates for the Eth site, G and U indicate data from “grazed” and “ungrazed” treatments at the Eth site, and xG/U,E,I is the value measured by Lane for the parameter being considered (runoff or sediment yield). In most cases, RE = 3. However, for the surface runoff data, two records from the Gardiner 2 exclosure (one for a grazed and one from an ungrazed treatment) were identical. Because this was clearly due to a data-entry error, these data were excluded from further analyses; thus, RE = 2 for the runoff data from that site. The expected value of D under the null hypothesis, D*, was determined by randomly assigning data to the grazed or ungrazed categories prior to calculating D. Rather than pooling data from all sites, they were randomized on a siteby-site basis. Thus, randomization of the data was stratified by site to control for intersite differences. This process was repeated 10,000 times, and the resulting distribution of D* values was examined to estimate the probability of observing values as large as or larger than D. Theoretically, in using the stratified randomization approach, data from a site with very high measured values (relative to the other sites) might still exert undue influence under some conditions. To assess whether results were robust to the details of the randomization procedure, we also used a nonstratified randomization procedure. In this second test, intersite differences were controlled using a data transformation. Measured values were divided by the mean value for that parameter at that site, with the mean calculated with data from both grazed and ungrazed treatments at that site. Observed differences between mean

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transformed values for the grazed and ungrazed treatments, Dt, were calculated in the same manner as for the stratified test. The expected value of Dt under the null hypothesis, Dt*, was determined by combining all transformed data from all sites, then randomly dividing those data among the grazed and ungrazed data pools prior to calculating D. Again, this process was repeated 10,000 times, and the resulting distribution of Dt* values was examined to estimate the probability of observing values as large as or larger than Dt. Results of the randomization tests (tables 11.1 and 11.2) were quite different from those reported by Lane (1990). Regardless of treatment, surface runoff and sediment yield were both significantly greater where grazing was present than where grazing was absent. Results did not vary greatly with the particular randomization procedure that was used, indicating they were not unduly influenced by aberrant data from any particular site. Contrary to Lane (1990) and Lane and Montagne (1996), we conclude that soil erosion on the northern Yellowstone winter range significantly increases when exposed to heavy ungulate grazing. The Lane and Montagne experiments may not reflect the full level of influence the elk had earlier in the twentieth century. An armored soil surface would be less erodible than a finer textured soil. The exclosures used in the experiments were constructed 30 years after the initial reports of sheet erosion, and some 70 years after the elk herd had increased to very high levels. The erosion measurable today may be a fraction of that occurring at the time of Rush and Wright and Thompson and even less than that in the decades preceding them.

Discussion A number of observers have inferred elk-induced erosion from landscape features on the northern range. In many cases the evidence is thin, equivocal, or

Table 11.1 Probabilities (p-values) that the Higher Runoff Rates Measured outside Five Exclosures (Grazed) Differed by Chance from the Betas Measured inside Exclosures (Ungrazed) Based on Reanalysis of Data Presented in Lane (1990) Treatmenta

Testb

D (or Dt)

95% CL from D* (or Dt*)

One-Tailed p-Value

1 1 2 2 3 3

S T S T S T

349.714 0.535 1119.000 0.922 700.071 0.913

–341.571 to 344.286 –0.631 to 0.624 –669.571 to 663.714 –0.505 to 0.496 –619.500 to 615.500 –0.563 to 0.569

0.021 0.051 0.051 0.001 0.011 0.001

a

Treatment: 1 is vegetation intact, 2 vegetation clipped and left on soil surface, 3 vegetation clipped and removed. b Test S is results from stratified test, T is results using transformed data.


Table 11.2 Probabilities (p-values) that the Higher Sediment Yields Measured outside Five Exclosures (Grazed) Differed by Chance from the Rates Measured inside the Exclosures (Ungrazed) Based on Reanalysis of Data Presented in Lane (1990) Treatmenta

Testb

D (or Dt)

95% CL from D* (or Dt*)

One-Tailed p-Value

1 1 2 2 3 3

S T S T S T

87.827 0.561 128.439 0.553 147.918 0.780

–90.721 to 90.706 –0.658 to 0.650 –126.093 to 125.456 –0.477 to 0.474 –158.044 to 153.886 –0.555 to 0.551

0.033 0.048 0.018 0.011 0.033 0.003

aTreatment:

1 is vegetation intact, 2 vegetation clipped and left on soil surface, 3 vegetation clipped and removed. bTest S is results from stratified test, T is results using transformed data.

nonxistent, and the park has understandably declined to agree with these inferences. One example is gully erosion. Sparsely vegetated gulches and gullies are some of the most striking erosional features on the northern range. A photo caption in Wright and Thompson (1935) described the erosion of topsoil from one such gulch. These authors did not attribute formation of the gulch to elk. Rather, they observed that overgrazing by elk had led to the loss of topsoil from the sides of the gulch. Still, Mount Everts currently experiences a high level of elk use, and questions have been raised as to the potential erosional effect that elk may have on the feature. The gullies develop on the west flank of Mount Everts (figure 11.2). At the top of the west-facing flank is a vertical cliff ranging up to about 25 m high; in places the cliff is absent. Below the cliff is a very steep, sparsely vegetated slope formed from Cretaceous shales and sandstones. This slope descends some 250 m over a horizontal distance of about 0.5 km. The flank of Mount Everts then grades into somewhat less steep alluvial fans that descend some 200 m over a horizontal distance of about 0.7 km. Gullies extend from the base of the cliff to the Gardner River below. As they pass through the alluvial fans, some gullies are deeply incised, ranging from 8 m deep at the top to 1–2 m deep at the bottom of the fan. Schmitt et al. (n.d.) examined 4 questions: 1. What are the primary processes by which sediment is transported from Mount Everts to the Gardner River channel? 2. How do these processes vary temporally? 3. What are the controls on sediment erosion, transport, and delivery? 4. Is there evidence that heavy grazing is contributing significantly to the delivery of sediment to the Gardner River channel?

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Figure 11.2 Gullies are a prominent feature on the flank of Mount Everts. Junipers in the foreground have been highlined.

Schmitt et al. found that the dominant sediment transport processes across the Mount Everts alluvial fans were debris flow, hyperconcentrated flow, and sheetflood. The debris flow and hyperconcentrated flow events occurred in active channels, whereas unconfined sheetflow occurred on the fan surfaces. All but one alluvial fan was truncated at the distal margin by the Gardner River. Although debris-flow levees have developed along the margins of gullies, Schmitt et al. concluded that the majority of debris-flow events do not spill out of the incised channels. In these cases, the sediment flows directly into the Gardner River and leaves no stratigraphic record. Using a combination of sedimentologic and dendrochronologic techniques, Schmitt et al. (n.d.) dated debris-flow deposits. They concluded that the processes that contributed to the construction of the fan over the past several hundred years are the same as those operating in recent time. From this, one could infer that there has been no change in processes due to changes in elk numbers that have occurred since the establishment of YNP. But as already noted, the authors found that most debris-flow events flow directly into the Gardner River and leave no stratigraphic record. Thus, debrisflow sediments do not provide an effective record from which to reconstruct possible elk influence. The authors observed that grass and shrubs trapped sheetflood sediment before it entered the Gardner River. If grazing by elk has had an influence on


erosion, that influence should be recorded in the sheetflood deposits. To date, no research has examined sheetflood sediments to determine if there has been a change in the rate at which they have accumulated. Due to its geologic and geomorphic characteristics, Mount Everts is a highly erosive feature. Geologic factors likely confound any attempt to identify whatever influence elk may have on erosion processes. Stream turbidity is a northern range phenomenon that has been another dimension of the erosion question. During spring meltout and after summer thunderstorms, northern range streams carry high sediment loads. Because the turbidity affects the quality of sportfishing, the condition has attracted considerable attention. Because a large segment of the public has come to associate erosion with elk, the turbidity has often been blamed on heavy grazing by elk. To determine the cause of that turbidity, Shovic et al. (1996) inventoried major erosional features in the northern range and in parts of the adjacent Gallatin National Forest. The authors classified erosion potential into three levels: high, moderate, and low. High erosion potential was defined to exist where significant sediment is produced from a major site at a 1- to 10-year frequency. Sites with high erosion potential have active rills, gullies, landslides, or evidence of sheet erosion. Sites with high erosion potential do not have significant erosion pavement or armoring by gravel or larger materials. Moderate erosion potential was defined to exist when noticeable erosion occurred on a small proportion of a site or significant erosion only occurred during extreme events. Low erosion potential was defined to exist when there was no visible active erosion; on low erosion-potential sites, relic erosion features may be present, but current transport mechanisms are ineffective in delivering material to perennial streams. This study concluded that within the northern range, high erosion potential sites were primarily limited to high-elevation drainages of the Lamar River and Soda Butte Creek. These upper drainages are steep, and the slopes are cut through erodible volcanic rock. The steep shale and mudstone scarps of Mount Everts were also identified as highly erosive. The erosive character of these sites was attributed to slope steepness and the erodible character of the shale and mudstone, a finding consistent with that of Schmitt et al. (n.d.). The principal objective of the Shovic et al. study was to identify the major sources of sediment load that enter the Yellowstone River. The authors concluded that current sediment input to the Lamar and Yellowstone Rivers is primarily due to geologic factors. By implication, elk have little effect. The studies by Shovic (1996) and Shovic et al. (1996) focused on current processes and conditions. The research did not apply that evidence to the reconstruction of historic rates and processes. The Schmitt et al. (n.d.) study examined the geomorphic processes at work on Mount Everts and drew inferences as to processes that existed in the past.

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Thus, some of the erosional features on the northern range that have been alleged to be induced by heavy elk use are so confounded by other geomorphic processes that any elk effect, if present, cannot be discerned with the research to date. But the observations of early investigators, our reanalysis of Lane’s (1990) results, and evidence presented in the next chapter make a persuasive case for accelerated erosion on the northern range since rise of the elk population during park history. That conclusion accords with Packer’s (1963) studies of erosion on the winter range of the Gallatin elk herd just outside the northwest corner of YNP. In 2 years of observation, Packer measured higher soil bulk density and erosion yield outside exclosures than inside. The conclusion also accords with Singer’s (1995) observation of significantly more bare ground outside the exclosures than inside and significantly more ground cover by lichens and mosses inside the exclosures than outside. It also accords with the observations of Frank et al. (1994) and Augustine and Frank (2001), cited in chapter 14, of accelerated downslope movement of soil nitrogen. Given the massive range ecology literature on the effects of domestic ungulates on soil erosion, it would be surprising if there were no effect on the northern range. Another northern range analog with livestock effects is the horizontal terracing on hillslopes formed by animal movements (Figures 6.5, 7.6, 11.3). How much topsoil has been lost from the northern range since park establishment cannot be determined from existing information.

Figure 11.3 Terraces running parallel with the slope are common on northern range hill slopes (see also figures 6.5, 7.6). Elsewhere, such terraces are common with heavy livestock use. Such terraces on the northern range are attributed to heavy elk use.


STREAM-BANK DESTABILIZATION Fluvial-Geomorphologic Evidence In a study to examine the contribution of stream-bank sloughing on turbidity of northern range streams, Rosgen (1993) studied stream morphology and processes in upper and lower reaches of the Lamar River and its tributaries Calfee Creek, Soda Butte Creek, and Slough Creek. Stream reaches were classified into eight types based on degree of entrenchment, width-to-depth ratio, and sinuosity. Those general types were further subdivided based on slope and particle sizes of sediment comprising the channel bed and banks. Stability was evaluated from morphological features, such as downcut channels and abandoned floodplains. Annual rates of stream-bank erosion were measured by inserting 3-ft (0.91 m) long erosion pins horizontally into the stream banks. Rosgen indicated that significant volumes of sediment are delivered to channels from steep erodible terrain in the upper reaches of the Lamar River drainage, a finding consistent with Shovic et al. (1996). The stability of a given stream segment is measured by the fate of sediment as it transits the segment. In stable stream segments, sediment passes through the segment with little or no change in stream pattern, dimension, grade, or profile, and the morphology of unstable stream segments changes in response. Streams high in the northern range above elk wintering grounds had relatively small width-to-depth ratios and stable channel dimensions. Reaches of the same stream type at lower elevations within elk wintering areas were less stable. These differences were correlated with presence or absence, respectively, of robust, woody stream-bank vegetation. Rosgen pointed out that this was a common pattern elsewhere in the West shown by research on streams impacted by livestock use. In one study, a soil column with 16–18% root volume was 20,000 times more resistant to detachment/erosion than a column of soil with no roots. On Soda Butte Creek, Rosgen showed the influence of vegetation on bank erodibility by comparing annual erosion rate of a C4 stream type segment in an upstream reach, where deep snow prevented elk from eliminating tall willow, to the annual erosion rate of a C4 stream segment in a downstream reach, where tall willow had been eliminated by elk. Where tall willows were present, banks receded 0.1 ft between 1989 to 1990; where tall willows were absent, more than 3.0 ft were eroded (figure 11.4). Rosgen noted that photos presented in Kay (1990), taken in the late 1800s and early 1900s, showed robust willow growth along the Lamar River and its tributaries (refer to figure 10.1). By the middle 1900s, retakes of these sites showed willow largely eliminated. He concluded that the changes were not attributable to climate but rather to browsing by an enlarged elk herd. Without the extensive occurrence of woody roots to stabilize the banks, erosion rates had accelerated.

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Figure 11.4 Bank sloughing on Slough Creek. Photographed July 17, 1997.

Bishop et al. (1997) generally criticized Rosgen’s work. But the authors did not direct their comments to Rosgen’s research design and evidence, including the erosion pins and 30× higher bank-sloughing rate in the Slough Creek stretches that had no woody vegetation. Rather they criticized his “erroneous conviction” that “large herds of elk did not inhabit the greater Yellowstone ecosystem until the 1900s.” Beschta (1999) reviewed geomorphic changes associated with the Lamar River and Slough Creek, using aerial photographs taken in 1954, 1971–72, and 1991. He emphasized 3 points. First, the loss of riparian woody species along riparian areas of the Lamar River and tributary streams has been extensive and was well under way before 1954. He inferred from ground-based photographs that the loss resulted from heavy ungulate browsing. Second, Beschta emphasized that the loss of woody species destabilized the Lamar River and major tributaries and initiated a major period of accelerated stream-bank erosion, increased channel widths, reduced sinuosity, increased amounts of braided channels, deposition of coarse sediments over former finetextured floodplain surfaces, and other disturbance features. Beschta noted that the Lamar River has, in several locations, moved almost completely across the width of the valley in less than 50 years (figure 11.5). This movement was considered high and believed to result from the major perturbation caused by the loss of woody plants within the floodplain. Finally, Beschta concluded that restoration of woody plants to pre-1900 conditions would not be possible, given the levels of ungulate herbivory that have occurred over many decades and continue to occur.


Figure 11.5 Upper Lamar River photographed in 1999. Compare with figure 10.1.

In 1993, R. B. Keigley found evidence suggesting that there has been a significant change in sedimentological processes along a segment of Soda Butte Creek. The area is located about 2 km east of the confluence of Soda Butte Creek and the Lamar River. In that vicinity, Meyer et al. (1995) mapped and dated six fluvial terraces. The oldest terrace lay about 8 m above Soda Butte Creek and consisted of late Pinedale glacial outwash (Pierce 1974). Meyer et al. (1995) reported that Soda Butte Creek had downcut to the modern floodplain level in the late nineteenth and twentieth centuries. At the area Keigley examined, Soda Butte Creek flows through a constriction formed by volcanic bedrock onto a floodplain that is confined by fluvial terraces to the south and to the north by morainal deposits. In 1986, a stream gauge (USGS Station no. 06187950) was installed at the downstream end of the bedrock constriction. Relic channels arc across the modern floodplain, which consists of two general kinds of deposits: one in which the surface layer is dominated by cobbles and one in which the surface is dominated by sand (figure 11.6). The cobble-dominated deposit occurs within the central region of the floodplain. The sand-dominated deposit occurs along the margin of the modern floodplain. In places the sand ranges up to 2 m thick. In the sand-dominated region, cobbles are absent on the surface. Two relic channels are shown in figure 11.6. Dead roots, presumably willow, 2–3 cm in diameter were present in the southernmost relic channel. Live willows with roots of this size were not present in 1993. During flood stage, Soda Butte Creek carries a heavy sediment load. The near absence of sand in the cobble-covered central region indicates that most of

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Figure 11.6 Map of Soda Butte Creek floodplain. The central region is composed mainly of cobbles. Sand-dominated deposits occur along the periphery.

the sediment is transported through this segment of the floodplain. Debris deposited by recent floods was limited to within a few meters of the low-water shoreline of Soda Butte Creek (see dashed lines in figure 11.6). Based on the distribution of sand and flood debris, the extreme margin of overbank flow lies well within the central cobble-dominated zone. In contrast, sand deposits up to 2 m thick along the periphery of the floodplain are evidence that in the past, sediment was deposited during overbank flow (see figure 11.6). Evidently Soda Butte Creek has downcut since the sand terraces were deposited.

Lichenometric Evidence Distribution of the lichen Xanthoria elegans provides additional evidence of downcutting. Xanthoria occurs in many ecological settings, including those associated with stream banks. Adjacent to streams, Xanthoria grows to a lower limit where it often forms a distinct orange band on rocks. In a study of Xanthoria at nine USGS stream gauge stations, the lower limit of such bands was found to correspond with river-stage height present in late June (Keigley 1997c). In the vicinity of the Soda Butte Creek gauge station, there is a well-defined band on the bedrock walls. Below the band on the bedrock walls, isolated lichen thalli


extend an additional 80 cm. The vertical distribution of the lichens appears to have been produced in two stages. First, the well-defined band of Xanthoria was established over a period in which late June stage height remained relatively constant. Then, in response to a reduction in stage height, Xanthoria colonized a zone 80 cm below the previous lower lichen limit. The relic (well-defined) lichen limit and the top of the sandy terrace corresponded respectively to 6.6 and 7.2 ft (2.0 and 2.2 m) on the staff gauge. Upstream from the study area, peak flood-stage height exceeded that lower lichen limit by about 0.3 m. If floodwater exceeded the relic lower lichen limit by 0.3 m, stage height in the study area in late June would have been about 2.32 m, a height sufficient to inundate the sand terraces. Peak flood occurs in early June; peak flood-stage height would have been even higher. Photographs show that willow covered the Soda Butte Creek floodplain in 1895 (figure 11.7, top). By 1948, elk had eliminated those tall willows (figure 11.7, bottom), which would have been a powerful stabilizing influence. The change in sedimentological processes is probably related to the elimination of willow from the floodplain.

CONCLUSIONS Based on their observations, park biologists expressed alarm in the 1930s over what they perceived to be serious elk-induced soil erosion. This view was maintained by their successors for approximately 40 years, who also showed evidence of the disappearance of stream-bank vegetation (see Kittams 1948). Based primarily on an interpretation of photographs, Houston (1982) concluded that accelerated erosion had not occurred. We have examined the erosion-related evidence and have, along with W. L. Hamilton in chapter 12, reanalyzed much of the data. Some of the results— for example, the gully erosion and turbidity evidence—do not provide answers. But several sources point to accelerated erosion during park history: the eyewitness accounts of Rush and Wright and Thompson; the surface rock-fragment evidence (although not unequivocally); the changes in stream-channel morphology; and convincingly, reanalysis of the pond sedimentation and Lane studies. In aggregate, evidence points persuasively to accelerated erosion during park history, and no significant body of evidence contradicts that conclusion. This evidence, except for the fluvial-geomorphological and sedimentation data, is conservative in that past erosion has probably caused the northern range to become less sensitive to erosion influences. The soils that exist today are in all probability substantially different from those that existed prior to the establishment of YNP. They have been subjected to more than a century of heavy use. Research should be undertaken to identify how existing ecological conditions differ from those that would have prevailed if the elk population had not increased to the high levels of the twentieth century.

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Figure 11.7 (top) A Haynes 1895 photograph showing willows (indicated by arrow) across the width of the Soda Butte Creek floodplain. (bottom) Kittams (1948) photograph showing absence of willows. Kittams attributed the loss of willows to heavy elk use.

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Influences on Ecosystem Function II: Historical Perturbations in Small Lake Basins WAYNE L. HAMILTON

12 Our notions of law and harmony are commonly confined to those instances which we detect; but the harmony which results from a far greater number of seemingly conflicting, but really concurring, laws, which we have not detected, is still more wonderful. —Henry David Thoreau

INTRODUCTION For decades the question of equilibrium between elk and the range in northern Yellowstone has occupied park personnel. After adoption of the natural regulation policy in 1967, the northern herd rose rapidly, and outside observers began noting effects on riparian habitat and range that resembled those that had earlier convinced the Park Service to control elk (and bison) numbers artificially. Park Service biologist W. H. Kittams began working in Yellowstone just after World War II. Kittams (1948) rephotographed portions of the northern range where older file photographs illustrated the preexisting condition of vegetation. The comparison led him to observe: “That willows once flourished in many of the moist sites within the northern range is borne out by several photographs. And, in many cases, the decline and inability to recover can be attributed to browsing by wild animals.” Kittams considered and rejected a climatic cause— including stream entrenchment and water table lowering—of willow decline. Later Kittams’s work focused on specific study plots, using photo comparison along with quantification of vegetation removal and soil loss attributed to elk. He observed (Kittams 1952): 231

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Wind prevents snow from accumulating on many exposures; hence it is instrumental in making forage available in late winter. When serial portions of the plants are removed and finally many of the plants die, wind erosion carries away much of the good surface soil and further aggravates the condition. . . . On many steep slopes the final damaging erosion is primarily due to water. Thin vegetative cover and compacted soil fail to keep the surface water where it falls. Instead, much of the water runs off and carries a heavy load of soil [with it]. Kittams (1952) described the progression of soil loss attributed to overpopulation of elk as follows: The one saving feature of most sites which become snow-free is that an erosion pavement—surface of rocks—finally is reached and greatly retards the action of wind and water. This, of course, is no correction of the problem; it simply emphasizes the need for preservation and improvement of vegetation before the pavement is established. These were the impacts recorded by park biologists at a time when the northern elk herd had declined from about 11,000 in the mid-1930s to approximately 9,000 animals (Houston 1982). As the son of a park official in the 1950s and seasonal employee, I was privy to these views. Herd control measures then reduced the population to under 4,000 by the late 1960s. With cessation of herd reduction, the herd size then returned by the late 1970s to numbers recorded in the 1930s, rising even higher in the 1980s. When I returned to work in Yellowstone in 1980 more than 20 years later, I began to hear concerns expressed that ungulate numbers had risen so high that range conditions were again deteriorating, as I had witnessed in the 1950s prior to the control measures. However, the concerns were now being expressed not by my new colleagues but by area ranchers, anglers, state soil conservation professionals, and some nearby university scientists. As the new coordinator of physical science research in the park, I convinced my supervisor in the mid1980s that sediment studies could be applied to such questions. I recommended palynological, paleontological, and sedimentological investigation of dated sediment cores as a means of quantifying effects of the changing ungulate influence in closed sedimentary basins. The result was a contract, the design of which I played no part, between the park and the Limnological Research Center at the University of Minnesota. The seminal publication was that of Engstrom et al. (1991). It reported a puzzling “asynchroneity of the stratigraphic changes among the lakes” and concluded that “the sedimentary record does not support the hypothesis that ungulate grazing has had a strong direct and indirect effect on the vegetation and soil stability in the lake catchments or on the water quality of the lakes.” Their findings were contrary to my direct observation that elk activity was more intense in the 1980s and 1990s than during the 1950s in the vicinity of

Influences on Ecosystem Function II: Historical Perturbations in Small Lake Basins

some of the lakes they studied. They implied that a threefold increase in elk numbers over the 30–year interval had produced no discernible effect on sedimentation. Was it conceivable that Kittams had misinterpreted his observations, leading the Park Service to wrongly implicate elk in causing erosion on the northern range? I carefully examined the data Engstrom et al. had used to draw their conclusions and published a brief comment in the Journal of Paleolimnology (Hamilton 1994a). In the interest of caution, propriety, and brevity, my comment considered only 1 of the several parameters reported by Engstrom et al.: total sediment. Observing that they had not incorporated site parameters in a synoptic assessment, I ordered the total sedimentation rate data by elevation and saw that rate changes exhibited a coherence that seemed to explain the asynchroneity they had observed. My comment noted that “the asynchroneity of the stratigraphic changes among the lakes” could be interpreted as a time-transgressive change in total sediment accumulation episodes from lower to higher elevation catchments over the northern range. I then proposed that available dendrochronological data and elk population numbers “suggest that ungulate foraging patterns might explain the time transgressive relationship with elevation,” and I suggested that “Reinterpretation of (the Engstrom et al.) sediment data suggests that reanalysis might lead to a different conclusion.” However, the authors did not review their other data synoptically to experimentally assess the merit of my suggestion and only restated their original interpretation (Engstrom et al. 1994). Hence I have found it desirable to conduct that review herein. Other details of the Engstrom et al. (1994) reply will be cited next in this more thorough analysis of their core data.

ANALYSIS OF THE ENGSTROM ET AL. (1991) EXPERIMENTAL DESIGN In their introduction, Engstrom et al. outline their objective and the hypothesis to be tested as follows: The objective of the present study is to examine the landscape history recorded in sediments of small lakes in the northern range in order to unravel the environmental perturbations of the last 150 years. Of particular interest are changes that took place after the Park was established in 1872. A simple model that isolates elk as the sole agent of environmental change in the northern range would predict the following: Overpopulation of elk should be recorded in lake sediments by (1) an increase in silt derived from eroded hillslopes or from trampled lake margins; (2) a decrease in the pollen of aspen, willow, alder, and birch; (3) an increase in the pollen of weedy plants associated with soil disturbance, including ragweed and various chenopods; and (4) an increase

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in indicators of eutrophic conditions in the lake, caused by increased manuring near the lake shore. A map of the study area is shown in figure 12.1. Their introduction describes variables other than elk numbers that could confound their analysis, including other ungulates, climate, fire, and human activities. There is no recognition in their Introduction and Methods sections, in the case of discharging lakes, of the potential influence on sediment parameters of pond outlet adjustments that lower or raise lake levels over time. Outlet and discharge characterization are described for only 3 of the 8 lakes in the Results section. Moreover, there is no consideration of the possible variables associated with the spatial distribution of the lakes spanning 90 km over an elevational range of 390 m in basins developed on different soil and lithological types and groundwater and surface-water recharge regimes. And there is no statement of how the authors would distinguish between the influences of elk numbers and climatic variations. Core segments from eight lakes were aged by the Pb-210 dating method, and sediments were analyzed for the following: relative abundance of pollen genera and diatom species; sediment density and chemistry, including loss on ignition for combustible organic and carbonate portions; and elemental analysis for some important cations. Both biogenic and allogenic silica were determined. Analytical error data are shown only for sedimentation rates, not as dating error, which must be calculated by the reader.

Figure 12.1 Map of the north portion of Yellowstone National Park showing major rivers and streams within the northern range. Lakes sampled by Engstrom et al. (1991) are indicated within circle symbols, and names are abbreviated as Middle Rainbow (MR), Big Slide (BS), Floating Island (FI), Buffalo Ford (BF), Big Trumpeter (BT), Slough (S), Foster (F), and Buck (B). Landslide deposits mapped by Prostka et al. (1975) and Pierce (1973) are approximately delineated by stippled areas. The 1932 boundary adjustment is shown north of Middle Rainbow lake.

Influences on Ecosystem Function II: Historical Perturbations in Small Lake Basins

Their Geochemical Interpretation section adds an important restriction to their test of the elk hypothesis. They state that even if sediment measurements show a large increase or decrease in deposition rate of silt and organic matter over time, these cannot be attributed to increase or decrease in ungulate activity around the pond basins, as stated in the introductory model, unless a change in chemical composition of the clastic and organic sediment is also simultaneously recorded. If such a change in elemental ratios, implying a new source, is not noted, the authors suggest that the sediments are probably being “focused” (redistributed within the pond basin or moved from the shallows to the deep-water coring sites). This introduces a conflict with the stated hypothesis, because its first element specifies “an increase in silt . . . from eroded hillslopes or from trampled lake margins” (emphasis added). More important, it is not clear how the new source requirement would apply if an increasing or decreasing number of elk were wading in, manuring in, foraging near, and utilizing salt effloresences along the banks of the lakes as they so commonly do. The authors did not explain how they proposed to distinguish between elk as a focusing agent and other agencies such as lake-level change or the action of wind and waves. Neither did they suggest an assessment of the heterogeneity of soil composition and lithology away from the shoreline in lake catchments as a test in applying this strict criterion. There is no statistical testing of their data against the elk population hypothesis or alternate causative variables. Their decision to present sedimentation rate data for total sediment only prevents more detailed elucidation of the components of the sedimentation. For example, to understand how sedimentation rate of allogenic silica (silt) changes over time, one must derive numerical values graphically from their allogenic silica concentration plots and multiply those values by similarly derived values from total sediment deposition-rate plots. The report is well illustrated graphically, and if one does not consider the limitations just stated and goes directly to the figures, one can easily conclude that several of the requirements of the elk hypothesis have been met and several have not. The conclusion that “the sedimentary record does not support the hypothesis that ungulate grazing has had a strong direct and indirect effect on the vegetation and soil stability [note the exclusion of the shore area] in the lake catchments or on the water quality of the lakes” led me to suppose that the elemental concentration data (their figures 10, 11, and 12) implied deposition-rate plots supporting the null hypothesis. The Engstrom et al. (1994) reply to my comment (Hamilton 1994a) criticized not my comments on their experimental design but my use of their data. They stated, “He also sidesteps the problem of inferring basin-wide fluxes (such as erosion) from single sediment cores. . . . Hamilton arbitrarily selected one (core from Big Slide Lake) as representative of the entire basin.” My rationale in using the deep water core should have been obvious because it was the only one comparable to the other seven deep-water cores.

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REANALYSIS OF ENGSTROM ET AL. (1991) DATA, 1872–CA. 1987 Sedimentation-Rate Changes Total Sedimentation In their figure 6, the deep-water cores at 6 of the 8 lakes exhibit increases in bulk sedimentation rate from the time of park establishment until the time of sampling (not stated, but implied by Wright 1987). Numerical values scaled from their plots (shown here in table 12.1), when all ponds are considered, indicate a mean increase of 2.4× (surface sedimentation rate divided by 1872 rate). The decrease shown at Big Trumpeter Lake contrasts with an increase for that lake illustrated in Wright (1987). A more cautious interpretation would exclude two lakes owing to road construction along the shore (Floating Island) and evidence of sediment desiccation during the drought of the 1930s (Big Trumpeter). The mean increase with this refinement is a factor of 2.9×, with only one lake (Foster) showing a decrease among the six. This result appears to support the elk population hypothesis (Table 12.1). It was this bulk sediment record that I assessed previously (Hamilton 1994a), showing that intervals of accelerating deposition rate at catchments in the central and eastern portions of their study area corresponded to growth suppres-

Table 12.1 Lake Water Chemistrya Pond

Date

T

pH

AlkT

Cl

K

Mg

Na

KN

NN

Cond.

16 19 17

7.5 9.0 8.2 8.8

138 67 176 120

0 0 1.8 —

2.7 1.3 3.9 —

4.3 2.9 28 —

18.6 26 4.9 —

.25 .97 .31 —

.01 3× increases in organic carbon and biogenic silica deposition in the lakes since 1872, increased eutrophication will persist as the likely result of large elk numbers. In short, until the question of elk impact on the northern range is explored both incisively and holistically, management by the NPS will continue to attract criticism by those who are aware of the reversal of the agency’s view of range condition that occurred in the late 1960s. As the data show, the condition of the northern range with respect to erosion and willow abundance did not abruptly improve—or even level off.

Influences on Ecosystem Function III: Bioenergetics

13 In this era of heightened environmental concern, it is essential that scientific knowledge form the foundation for any meaningful effort to preserve ecological resources. —Richard W. Sellars

PERSPECTIVE Characterizing the bioenergetics of ecological systems in terms of energy flow through trophic levels and food chains or webs has been a useful conceptual tool in ecology. Its use as an analytical tool with production at one trophic level modeled as some function of the production of lower levels has been somewhat less general. It has been most applicable in aquatic systems where all or a major part of each trophic level is a food resource for and a potential restraint on the next higher level. It has been less useful in terrestrial systems, particularly forests, where herbivores consume a minor fraction (e.g., 5 cm in diameter. He estimated the ages of ~497 narrowleaf cottonwoods. None of these had developed during the past ~60 years.

Summation of First-Order Vegetation Effects I have now reviewed the evidence available on northern range vegetation in four sequential, chronological periods: prior to and in the early years following park establishment; approximately 1884 to 1959, when the northern herd numbered more than 6,000, in some years 3 to 4 times this number or more; 1959–70, the period of extreme herd reduction and first few years of recovery; and 1971 to early 2000s when the herd had returned to high numbers. Except in 1 or 2 cases of vegetative components in 1 or 2 periods, there is evidence of change in each of the 6 vegetative components in all of the 4 time periods. In some cases the evidence is fragmentary. In most cases it is circumstantial. But the sequence in essence constitutes two pseudo-replicates each of two treatments for each component. The inside-outside exclosure comparisons and inside-outside park comparisons constitute further pseudo-replication. In all cases the patterns are consistent with what would be expected of large variations in elk herbivorous pressure and with the extensive literature on herbivory, including the immense range-ecology literature, which has been largely ignored in park publications. As discussed in chapter 1, the northern range may be entering a fifth replicate in response to wolf effects. Singer et al. (2001a) observed increase in willow heights, possibly in response to changes in elk distribution driven by wolf hunting. Ripple and Beschta (2003) observed differences between 1995–96 and 2001–2003 photographs in the heights of cottonwoods in “low-risk” and “highrisk” sites along Soda Butte Creek and the Lamar River. Cottonwood heights increased on low-risk sites but not on high-risk ones. And Ripple and Beschta (2004) photographed willow regrowth on Blacktail Creek. The evidence overwhelmingly indicates alteration of all components of the vegetation when the northern herd increases to levels above 6,000–10,000. The effects range from nearly complete elimination of some components, as with aspen woodland, to altered structure as with sagebrush-steppe and conifers. With pronounced changes occurring in all components of the vegetation in response to different levels of elk numbers, it is to be expected that the other sectors of the ecosystem that interact with the vegetation would also vary. Though most of these sectors have had little or no research attention, there is evidence of response in those that have been studied, particularly the other ungulate

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species. I now summarize that evidence that has been presented in previous chapters, grouped according to the services provided by the vegetation.

Second-Order Effects via the Vegetation Reduced Food Resources for Herbivores I discussed at considerable length in chapter 9 an apparent population decline in three species of upland ungulates during park history as elk increased an estimated threefold by the 1990s and possibly four- to sevenfold by the early 1900s. Bison may be an exception, possibly increasing during this period when freed of the same constraints as those limiting the elk in prehistory. Collectively, the declining upland species are two browsing and one grazing form. Pronghorn, for which sagebrush is a winter staple, formerly occurred in large numbers in the northwestern area of the northern range where sagebrush has been profoundly reduced (figures 7.5, 7.6). I suggested pronghorn decline on the order of 90%. Mule deer appear to have declined initially to the 1940s to 1960s (figure 9.5) when they wintered inside the park, but then recovered numbers during the 1980s and ’90s when they began wintering outside the park and evading elk competition. Bighorns are grazers, probably experiencing competition with elk and bison for a heavily usurped herbaceous vegetation. I suggested ~ 90% reduction during park history. They showed some evidence of population recovery during the elk population reduction (figures 9.1), but declined again in the latter twentieth century. In chapter 10 I discussed two browsing riparian species that probably declined in response to the sharp reduction in the riparian zone. Whitetailed deer disappeared completely as northern range residents by the 1920s. Moose largely disappeared from the northern range following their late appearance in the park early in the twentieth century. Singer et al. (1998a) agreed that this species’ decline was probably driven by elk competition. The virtual disappearance of beaver from the northern range may be partially the result of food deprivation. The willows and aspens that are their primary food sources are also their major building materials. When removed from or near aquatic sites, the animals are unable to build the dams that form the ponds that provide colony habitats. Beyond ungulates and beaver, there has been no work examining the effects of vegetation alteration on the food sources of small mammals and invertebrates. There almost certainly have been extensive influences.

Habitat Reduction The changing vegetative structure undoubtedly has altered habitat for the many species that use the vegetation for that purpose. But there have only been a handful of studies in the northern range to document this effect. I cited several studies in other areas of the Greater Yellowstone Ecosystem in chapters 6, 7, and 10 that doubtless reflect what is occurring on the northern range.

Synthesis

These studies have shown several avian species that are unique to aspen and disappear with elimination of the woodland. Other species decline as the structural diversity of the vegetative type is reduced (see Dobkin et al. 2002). Similar effects occur in cottonwood stands. Jackson (1992, 1993) documented declines in avian numbers and species as browsing intensity increased in northern range willows. Berger et al. (2001) observed the same effect in Teton National Park willows in response to moose browsing. I am not aware of any studies that have measured effects on the avifauna of structural modifications in the northern range sagebrush-steppe. But in chapter 7 I cited work elsewhere that indicated such effects. One can see an extreme example by observing the avian activity inside the Gardiner exclosures and the virtual absence of such activity outside (figure 7.5). As discussed in chapter 10, Debinski (1994, 1996) discussed the tenuous position of the Yellowstone checkerspot butterfly in the northern range riparian zone. She also observed (Debinski and Brussard 1994; Debinski et al. 1999) correlations between butterfly and avian diversity suggesting that browsing impacts on riparian vegetation has the same effects on butterfly diversity as on avian diversity. I cited G. E. Beetle’s (1997) work on snails in aspen stands in chapter 6, with habitat reduced to what she called “decadent” stands for the 8– 11 snail species she observed. Beyond these few studies, I am not aware of any research on the hundreds of northern range invertebrate species that would give clues to their population trends during park history and the possible elk effects on those trends. Nor am I aware of any studies on the small mammals, although research elsewhere has explored their relationships to sagebrush-steppe vegetative structure. I also do not know of any northern range studies on the effects of beaver pond disappearance on the aquatic fauna of this type.

First- and Second-Order Effects on Carnivores As discussed in chapter 9, the northern range has long had higher coyote densities than most of the northern portion of the western United States. Their conspicuous presence prompted Murie’s (1940) study of the species in the 1930s. Murie observed that elk carrion was a major source of winter food for the northern range animals. Knowlton (1972; Knowlton et al. 1999) has concluded that availability of winter food is a major determinant of coyote abundance. Hence the high northern range elk population has in all probability been a major determinant of the high coyote densities. The high coyote population, in turn, has influenced other system components. In chapter 9 I cited the O’Gara (1968) and Barmore (1980) observations of high pronghorn fawn mortality exacted by coyotes. Thus the large elk population applies the dual pressures on pronghorns of forage competition, and what Holt (1977) calls “apparent competition” by supporting significant coyote predation. The high coyote population may also suppress populations of other mediumsized carnivores (Wagner 1988). Gese et al. (1996) have observed aggressive

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encounters between coyotes and red fox (Vulpes vulpes) on the northern range, a relationship that has been observed elsewhere (Gosselink et al. 2003). There is also reason to infer dual effects on grizzly bear. Kay (1990) has pointed out that riparian zones in the eastern slopes of the Rocky Mountain system are important travel lanes and food sources for the species and postulated that the zones’ impoverishment on the northern range has been detrimental to grizzlies. To this I would add the heavy impacts on berry-bearing shrubs, discussed in chapter 8. But other authors (Mattson et al. 1991; Green et al. 1997) have pointed out the heavy use of elk carcasses and calf kills by grizzlies emerging from hibernation in spring and the likely benefits of these. All of these interactions are subject to change, depending on the effects that wolves will have on both elk and coyotes. This will be touched on in the next section.

EFFECTS ON SYSTEM PROCESSES Ecosystem function is driven by myriad processes acting over a range of phenomenological integration: physiological, population, community, and system among the biota and chemical and physical processes among the abiotic components. Because the purpose here is to synthesize northern range ecosystem structure and function, I will comment here only on system-level processes. There have been only a few studies of northern range system processes, and these have spanned only 1 to 2 years. Hence there is no basis for correlating variations in process measurements with the phases of elk abundance. Some tentative inferences can be drawn from periodic standing-crop measurements over a period of time. Evaluating how primary production of the northern range system today compares with the level prior to 1872 requires a combination of production values for the several vegetation types and for these two time periods. There are no data for such a comparison. So only a crude perusal of numbers, mostly from standing-crop measurements, can be brought to bear on the question. I cited available data on herbaceous primary production on the northern range in chapter 13. All are from either 1- or 2-year studies. But one can perhaps infer production trends from the standing-crop measurements shown in figure 7.8. These showed low levels of herbaceous standing crop in 1958, 1962, and 1967, the first two measured at the end of a seven-decade period with a large elk population. The measurements increased between 1967 and 1974, the latter few years with reduced herd size. They approximately leveled off from 1974 to 1981, then declined in 1986 and 1990 when the northern herd had returned to high levels. Thus, to the extent that standing-crop measurements at the end of the growing seasons are indices of herbaceous annual production, the latter have varied with changing herd size.

Synthesis

But as I have discussed, there is some evidence that the herbaceous vegetation is influenced by sagebrush competition, and that the herbaceous trends may in part be a function of sympatric sagebrush trends as well as elk grazing. Thus, trends in herbaceous vegetation are probably determined by these two influences, and one cannot meaningfully speculate on how primary production in this type today compares with its level prior to park establishment. It certainly did not experience the grazing pressure that it had during the twentieth century. I reviewed the work of D. A. Frank in chapter 13 and found the evidence for overcompensation mixed at best, and not persuasive that it is occurring in the northern range grasses. In total, it seems unlikely that herbaceous production today is higher than it was in 1872. Despite sagebrush regrowth since 1962 (figure 7.4), the inside-outside exclosure comparisons indicated that browsing constrained the recovery through 1990. And Wambolt and Sherwood’s (1999) extensive measurements in 1994 showed sagebrush outside the exclosures producing at one-third the inside levels per unit area. An 80% reduction in aspen clearly implies an equivalent decline in production from 1872 levels by a species that formerly occupied on the order of 10% of the northern range. Woody riparian vegetation has evidently declined by a comparable order of magnitude, although it is a smaller component of the vegetation, perhaps originally occupying no more than 1%. A major contributor to northern range primary production is the conifers, for which, to my knowledge, there has been no research on the northern range. There have been three documented significant changes in this type during park history. One was its spread of ~5% reported by Houston (1982). The second was the 1988 fire. Although there are estimates of the total proportion of the park’s area that was burned, I am not aware of any similar value for the northern range, at least for the coniferous vegetation. The third change has been the ubiquitous highlining during the twentieth century. Thus, whether there have been changes in total coniferous production, what is within reach of browsing ungulates and smaller animals has decreased. In total, there is no way of calculating primary production for the entire northern range vegetation today or how that would compare with production levels at the time of park establishment. But with the clear decline in aspen and riparian production, evident reduction in sagebrush, either similar or reduced production in herbaceous vegetation, and significant decline in coniferous growth within reach of ungulates, the total primary production available to these and smaller animals is in all probability reduced from the levels present in 1872. Whatever the trends in primary production during park history, it seems clear that the speculative estimate of a 2.3× increase in ungulate biomass since park establishment (table 9.2) implies an increase in ungulate consumption and secondary production of similar magnitude. Except for the near disappearance of northern range beaver and decline of some avian species, nothing can be said about the remainder of the fauna.

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Coughenour (1991) commented, on the basis of inside-outside exclosure measurements, that grazing diverts herbaceous plant material into ungulate production. The material would otherwise become litter—presumably much of it did prior to park establishment—and courses through the detritivory channels of the ecosystem. If only a limited amount of research has been conducted on northern range bioenergetics, even less has been done on nutrient cycling and that largely the work of D. A. Frank and co-workers on nitrogen. These authors make a convincing case for ungulates’ role in accelerating the nitrogen cycle on the northern range, and Coughenour (1991) measured higher nitrogen content in vegetation outside exclosures than inside. The result may be more rapid nitrogen cycling on the northern range during the twentieth century than before 1872, when few ungulates wintered there. But the input-output processes have not been systematically studied to work out the system budget. There is evidence of accelerated surface erosion since park establishment and the destruction of microphytic crusts. The latter could reduce nitrogen fixation. In total no mass balance can be calculated at this time. Hamilton, Keigley, and I discussed in chapters 11 and 12 the evidence indicating that surface erosion has increased during park history. Although some of this could be the result of hoof impacts, it is in all probability induced at least in part by vegetation alteration and released constraints on soil movement. Singer (1995) measured 2 to 8 times more bare ground outside exclosures than inside. We also discussed evidence of stream-bank destabilization that resulted from the widespread elimination of riparian vegetation. The fluvial geomorphology of the Soda Butte and Lamar Rivers today are quite different from what is shown in early photographs.

EFFECTS ON ECOSYSTEM STRUCTURE AND FUNCTION Concepts Clarified With essentially all of the measured components of the northern range ecosystem responding to variations in the elk population, it must follow that the synthetic characteristics of the entire system are also responding. I will examine those responses in this section. But before doing so, I need to clarify concepts to be discussed to ensure clear communication. I use the term ecosystem in the standard textbook form: a group of interacting components, biotic and abiotic, that collectively perform one or more functions that the individual parts could not perform separately. The concept has the additional implication that if one part of the system is changed, the other parts, and hence the system, are changed. Ecosystems have spatial limits, usually arbitrarily set for purposes of study or management. I have accepted the borders of the northern range (figure 1.1)

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as the limits of the northern range ecosystem. These delimit the area in which the northern herd winters. There are no closed ecosystems. All have inputs and outputs of energy and material that cross or penetrate the limits. Even the Earth is not a closed system. But contrary to Chase’s (1986) rejection of the concept because no ecosystems are closed, the existence of inputs and outputs does not invalidate ecosystem analysis as long as they are measured and included in any mass-balance formulations. Ecosystem analysis investigates processes that involve entire systems, as opposed to the functioning of the individual parts. Such processes include acquisition, containment, and exchange of energy and material. They have had very limited research attention on the northern range. However, energy and materials are acquired, contained, and exchanged by the components that have received research attention. Hence system processes can be inferred to some degree from the component dynamics, which also provide the system’s physical structure, and the constraints on both structure and function. In the sections that follow, I will attempt to infer structure and function of the northern range ecosystem in terms of the components discussed up to this point in the book.

The Northern Range Ecosystem Structure A Hierarchical Structure As stated at the beginning of this chapter, O’Neill et al. (1986) and Allen and Hoekstra (1992) have generalized ecosystems as hierarchies of functional components fundamentally structured by systems of constraints. Though O’Neill et al. considered that there may be several different forms of hierarchies, they emphasized a dual structure: the ecosystem is a dual organization arising from the structural constraints that operate on organisms and functional constraints that operate on processes. . . . Once the constraints are lost, the hierarchical organization is lost. When a system goes unstable, it is the normal functioning of unconstrained components that tears the system apart. (O’Neill et al. 1986:210–11) Allen and Hoekstra (1992:34, 90) further emphasize the important role of pathways: frequency and constraints are the most important criteria for ordering levels. Upper levels constrain lower levels by behaving at a lower frequency. . . . We define the parts and explanatory principles of ecosystems as pathways of processes and fluxes between organisms and their environment . . . the critical parts are the pathways that may involve organisms, not the organisms themselves. The northern range system can be conceptualized according to these criteria into a hierarchy. But for illustrative purposes, the concept must be simplified to

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the few, major driving forces affecting the system’s structure and function. Smith et al. (2003) have characterized the northern range biota with a network of organisms connected by trophic or competitive links that is highly instructive in representing the complexity of the system. But as mentioned in chapter 10, Holling (1995) generalizes that ecology is moving toward a paradigm in which most of the functioning of complex ecosystems is driven by a small number of its components and processes. This appears to be the case with the northern range. The northern range ungulate guild is a case in point. Technically, all five upland ungulate species apply herbivorous pressure to the vegetation, and each could be represented as a single block in the flow of influence in a hierarchy. But as discussed in chapter 9, elk constituted 84% of upland ungulate numbers, made up 91% of ungulate biomass during the 1990s (table 9.2), and are the major herbivorous force affecting the vegetation. Moreover, the major interaction between the four nonelk species with the vegetation is not their first-order impact on it but its usurpation by elk and the consequent second-order competitive pressure on them by the elk. The position of bison, one of the four, may be intermediate between the extremes of elk impact on the vegetation and the other three species heavily influenced by elk competition. Hence, I have simplified representation of the system into a hierarchy comprised of the few major driving forces affecting its structure and function subject to the following provision: 1. Elk are positioned by themselves in the hierarchy for the above reasons. 2. The O’Neill et al. (1986) and Allen and Hoekstra (1992) constraints and the latter authors’ pathways are in some cases one and the same. 3. The magnitudes of the interactions, or in the above authors’ lexicon the strength of the constraints, have varied between the different phases of elk abundance. Hence the structure of the system has changed, and the chronology of the hierarchy must be specified. Consequently, the evidence points to the northern range ecosystem represented by a trophic hierarchy shown in figure 15.1. Prior to park establishment the top predators—humans, wolves, cougars, grizzly bears, coyotes—and winter migration out of what is now the northern range exerted the top-level constraints and prevented heavy ungulate pressures on the vegetation. Kay (1990, 1994a, 1998) has stressed the importance of aboriginal hunting and identified Native Americans as a keystone species. Though the indices of wolf abundance discussed in chapter 3 pointed to low wolf numbers prior to 1872, they were not necessarily inconsequential predators on elk and other ungulates. The latter were also present in low numbers. Hence the prey:predator ratios may have been low enough to allow the wolves to exert a significant predation rate. The predatory constraint on the elk herd and other ungulates minimized their impacts on the vegetation and allowed it to develop the profuse growth shown in the early photographs and commented on by the early observers. At

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Figure 15.1 Hierarchical structure of Yellowstone’s northern range ecosystem at park establishment and in the 1990s. Rectangles are major system components; ovals are constraints and/or pathways as described by O’Neill et al. (1986) and Allen and Hoeckstra (1992). Arrows represent the directions of influence.

that point, the major constraints operating on the vegetation probably were competition functioning within localized conditions of climate, topography, soils, and periodic fires. The full expression of the vegetation then provided ample food resources for the ungulates, beaver, and other herbivores; habitat for birds, small mammals, and invertebrates as discussed in chapters 7, 8, and 10; and constraints on watershed and fluvial processes as discussed in chapters 11 and 12. The entire pattern is consistent with the quote from Allen and Hoekstra (1992): The structure and function of the entire hierarchy is maintained by the constraints at the top level. With removal of the top predators—transfer of Native Americans away from the area, control of the large carnivores, and protection from Euro-American hunting by the park boundaries—the top-level constraints were removed. Along with cessation of winter migrations, the number of elk wintering on what became the northern range increased to several times prepark numbers. It then exerted the heavy pressures on the vegetation discussed in previous chapters, and the elk and other components of the system became limited by the reduced vegetation. Kay (1998) characterized these changes as conversion from top-down to bottom-up limitation, or as change from a predator-limited to resource-limited system. Again it coincides with the Allen and Hoeckstra (1992) quote on what happens to a system when the top-level constraints are removed.

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In an analogous model, Ripple and Beschta (2004) represent the changes in the northern range system as a trophic cascade occasioned by the removal of wolves. They suggest that wolves, and associated predation risk, can structure the system. They emphasize system responses (vegetation recovery, increase in beaver) in areas from which elk have moved to avoid predation risk by the presence of wolves: “Can predation risk structure ecosystems? Our answer . . . [on the basis of evidence and theory] is yes.” This is indeed part of the story. But the evidence also points to a likely role of aboriginal hunting, perhaps a major one. Wolves were still present, though being killed, through the period between park establishment and the 1920s when the elk herd burgeoned. Constraints on aspen ramet development began in the 1890s, and severe impacts on other components of the vegetation were evident at least by 1914. Moreover, as the evidence shows, it is the number of elk and the duration of use that are the main determinants of the degree of impact. Changing the distribution of the animals may provide system releases in areas they avoid, but this only shifts the distribution of the herbivorous pressure on the system as a whole. That pressure is not reduced unless elk numbers are reduced.

Declining Biodiversity The changes in the northern range ecosystem since park establishment can also be characterized as progressive reduction in biodiversity. Boyce’s (1998) comments on diversity seriously oversimplify this aspect of changes in the system. He is correct that Chadde and Kay (1988) found higher species richness outside northern range willow exclosures than inside. And this is probably an example of other range ecology findings that grazing mediates competition between plant species and permits coexistence of more species, much as Paine’s (1966) classic starfish- predation study showed maintenance of diversity in intertidal invertebrates. But the pattern has not been consistent on the northern range: Singer (1995) and Reardon (1996) did not find inside-outside differences in herbaceous species composition at the large exclosures. Beyond these cases, diversity is a far more complex matter. It can be measured at a range of biotic levels—genetic, species, communities as in the above cases, and landscapes—and spatial scales. The northern range is a tapestry of subsystems: aspen woodland, sagebrushsteppe, coniferous stands, riparian zones. Some of the more mobile animal species use several or all of these, variously for feeding, cover, reproduction, and movement. The landscape thus provides what a number of authors are calling “habitat complementation” (Pulliam 1988; Pulliam and Danielson 1991; Dunning et al. 1992). Changes in their abundance and proximity within the animals’ mobility ranges may foreclose the northern range for some species. Kay (1989, 1990) points out that riparian strips are important travel lanes for grizzly bears. Their near elimination from the northern range may have contributed to the species’ decline in that portion of the park.

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More directly, the widespread elimination of aspen, willows, and cottonwoods constitutes a reduction in habitat and landscape diversity. The almost certain decline in avian and insect species obligate to these types, and likely decline of more generalist forms using them as discussed in chapters 6 and 10, must reduce the area’s species diversity. Two documented mammalian cases are the disappearance of white-tailed deer and virtual disappearance of beaver. Boyce’s disclaimer that these are of no consequence because they are abundant outside the park is not relevant to the current question of diversity trends on the northern range. A more subtle scale is the situation of low-mobility animal species existing in meta-populations within fragmented landscapes. An entire symposium (Ehrenfeld 1995) discusses species persistence in fragmenting landscapes where habitat patch sizes decline, interpatch distances increase, and corridors between patches disappear. Local extinction probabilities increase in arthropods, plants, amphibians, birds, and small mammals (Fahrig and Merriam 1995), and repopulation probabilities decline, with increasing landscape fragmentation. These questions have received no systematic research attention on the northern range, the few exceptions being Jackson’s (1992, 1993) work on willow avifauna, Debinski’s (1994, 1996) on butterflies, and Romme’s (see Romme and Knight 1982) and Turner’s (Turner et al. 1994) work on disturbance effects on landscape dynamics. The fact that the few studies that have been conducted have shown the expected effects on species diversity suggests that they may be common in the northern range biota. The NRC review (Klein et al. 2002) made a number of statements about the northern range relevant to this discussion on changing biodiversity that need to be addressed. They are not helpful in and of themselves, but they prompt more meaningful conceptualization of the changes that have taken place. Thus the review stated that: “Yellowstone is not in ecological trouble . . . not on the verge of crossing some ecological threshold beyond which conditions might be irreversible . . . has not been associated with ecological disaster. . . dramatic ecological change does not appear to be imminent.” Without ecological definition, these terms are not meaningful and thus do not convey the reality of what has occurred in the northern range ecosystem. What constitutes ecological trouble, ecological disaster, dramatic ecological change, or thresholds in terms of the system’s components and processes is not clear or probably conceptualized. What has occurred is a progressive biotic impoverishment in the sense of a continuum. In that perspective, what is a disaster? Has a threshold been crossed when a species (e.g., whitetailed deer) is eliminated? Or two species (whitetails and beaver)? Or if bighorn and pronghorn also disappear? Or if a habitat is eliminated? Aspen and its obligate fauna, now reduced by ~80%, could disappear with continuation of the trends of the 1990s. Has an ecological threshold, beyond which conditions might be irreversible, been crossed with loss of several centimeters of topsoil? Or with alteration of the fluvial geomorphology of the Lamar River? The reality is a progressive reduction of biodiversity and change

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in system structure. Given a hypothetically long enough period of time with continuation of trends of the 1990s, the system could be markedly changed. But some form of ecosystem would continue to function. Would the change then have been “dramatic,” or a “disaster,” or would a threshold have been passed? In a similar case of ambiguity, Boyce (1998) misquoted Wagner and Kay (1993) in ascribing to them the statement that “Yellowstone will somehow self destruct.” If trends into the 1990s were to continue, the northern range might be expected to continue to lose species—pronghorn and bighorn might be next— and landscape diversity. But some form of ecosystem would continue to persist. Wagner and Kay never called this self-destruction. And if we designate the prepark state as the “natural” (i.e., without Euro-American influence) state, the system today is well out of the range of natural variation. Finally, the point needs to be made that all of the changes that have occurred cannot be attributed to the natural-regulation policy. Much (perhaps most) of the change had occurred by 1967 after ~75 years of use by an enlarged elk herd. The effects of the policy have been to resume the changes underway before the herd reductions and probably to accentuate them further.

Alternate Steady States? The question arises as to whether the northern range ecosystem has occupied one or more stable states. Most authors define stability as the tendency for a system to return to some “persistent configuration” following perturbation away from that configuration (see May 1973, Case 2000). Recent literature has questioned the reality of stability and its related concept equilibrium in natural systems, often citing its treatment by Botkin (1990). Part of the debate is semantic, confused on the question of time scale. As O’Neill et al. (1986), Allen and Hoeckstra (1992), and Wagner et al. (1995a) point out, equilibrium is a matter of time scale. On geological, evolutionary, or climatic time scales or during continuing human disturbance there are no equilibria. But for periods of time scaled in years or decades, many systems remain in roughly constant ranges of values over time. A related problem is confusion with the condition of stationarity, the total absence of change. Natural systems in variable environments fluctuate over time and may rarely, if ever, settle precisely on single equilibrium values. But negative feedback processes place probability limits on and restrict the range of values within which a system fluctuates. They exert continuing pressures to move systems back toward (if not precisely to) mathematical equilibrium points. The density-dependent pressures operating on the northern elk herd function in this manner, as discussed in chapter 2. Equilibrium of natural systems in varying environments is thus absence of net or mean trend over specified periods of time, maintained within roughly constant limits of variation. The state of the northern range system was clearly different prior to park establishment from its state during the 1900s. Whether that state for some decades or centuries prior to 1872 could be considered stable obviously can only be speculated on. There were climatic variations including the Little Ice Age and

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the Medieval warming period. Yet temperature proxies for the past ~400 years do not show any secular trends or extreme anomalies until the mid- to latter nineteenth century, initiating the contemporary warming period (Mann et al. 1999; Crowley and Lowery 2000). Kay (1990) has pointed out the relative scarcity of ungulate remains down through the western North American archaeological columns for what must be centuries into the past suggesting low ungulate densities through that period. And the northern range system changes associated with the 1959–70 low elk numbers, slight and short-lived though they were, tended to move the northern range back toward the prepark state. Whether the prepark system existed in some roughly stable state, it was clearly different from its state during the 1900s. The change was elicited, the earlier state destabilized, by establishment of the park and associated management actions that removed the earlier constraints. The change is symbolized by the shift from left to right in figure 15.1. Moreover the evidence does not suggest any stability during the 1900s. The system has undergone continuing change during the period.

What for the Future? What the future, perhaps 50–75 years, holds in store for the northern range can, of course, only be speculated on. If one were to predict only on the basis of trends up to the mid-1990s, no significant changes in the northern range environmental circumstances, and continuation of current policies, one could reasonably project continuation of the trends of the preceding 20–30 years or indeed most of the twentieth century. The northern herd would probably continue for a time to fluctuate around an equilibrium value of 15,000–20,000 censused animals, but this might decline later into the period. Aspen and cottonwoods could well disappear, as would most willows. Sagebrush would likely decline further with grasses becoming dominant steppe vegetation. Nonnative plant species would probably increase and faunal diversity would surely decline. But three recent changes in environmental circumstances may well change this scenario or may already be doing so. The first is the 41% expansion of the winter range that began in 1988, and the movement of, on average, one-third of the herd into this extended area. This could be lightening the elk pressure on the original 108,553 ha somewhat. But two-thirds of 16,500 censused animals were sufficient to alter the vegetation in the ways discussed in previous chapters on the original 108,553 ha. Hence, the relief might not be great. The expanded range could allow some herd increase, as the Taper and Gogan (2002) calculations suggest. In that case the long-term trend could be eventual depletion of the new 52,663 ha, and resumption of the trends in the entire 152,663 ha that prevailed before 1988 on the 108,553. A second change is the wolf reintroduction. Smith et al. (2003) provide an excellent overview of the growth and distribution of the wolf population since release in 1995, and early indications of possible northern range changes. How

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many (if any) significant changes occur will depend on how much effect the wolves have on elk and other components of the system. The wolf population may have stabilized at approximately 14 packs and 132 animals in the park and 8 packs and 77 wolves in the northern range. Hence, whatever effects occur are likely to result initially from the current population size. Recent modeling exercises (Eberhardt et al. 2003) project potential effects ranging from slight (~ 10%) reduction in the elk herd to almost complete elimination of both elk and wolves, depending on the equations and parameter values used. If the wolves succeed in sharply reducing the elk herd, it is likely that the wolf population would decline as well. In chapter 3, I cited evidence presented by Kay (1990) and Schullery and Whittlesey (1992) indicating that wolf numbers were low at the time of park establishment, an indication of low ungulate numbers. There is still no unequivocal evidence, as this is written in 2004, of significant wolf effects on the northern herd. Calves have been 43% of wolf-killed elk (Smith et al. 2003), and winter-end calf:cow ratios in the late 1990s and early 2000s have been half or less the values in preceding years according to T. O. Lemke (personal communication, July 18, 2001). The herd has declined in the early 2000s, but as discussed in chapter 1, these have been drought years, and Merrill and Boyce (1991) and Coughenour and Singer (1996) have shown strong correlations between northern herd recruitment and population growth rates and precipitation. Lemke has pointed out to me that calf:cow ratios and elk numbers have been down in these years elsewhere in western Montana, where there are no wolves. Some recent northern range system changes appear to be associated with wolf activity, particularly due to alterations in elk distribution and movement. Several investigators have observed increase in willow growth in the vicinity of wolf dens (Singer et al. 2001a; Smith and Guerny 2002). Ripple et al. (2001) measured greater aspen sucker heights in high wolf-use areas than in low. Fecal pellet group indices showed lower elk use in the high wolf-use areas than in the low. Ripple and Beschta (2003) measured higher browsing intensity and lower heights of cottonwood in open terrain with clear surrounding views than in sites where visibility was blocked by stream banks and vegetation. The authors hypothesized that elk occupy the more open sites which permit visibility of wolves. As stated before, Ripple and Beschta (2004) observed willow regrowth in response to wolf-driven shifts in elk distribution. However, until and unless there is significant reduction of the northern herd, these changes may only represent redistribution of browsing pressure and not actual reduction over the area as a whole. Crabtree and Sheldon (1999) observed a 50% reduction in the northern range coyote population following wolf reintroduction. Wolves have been observed killing coyotes. Smith et al. (2003) reported some increase in pronghorn fawn survival in recent years, possibly associated with the coyote reduction. If these wolf-induced changes continue, and particularly if they become significant, they will in essence move the system back toward the pre-1872 state.

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This will be accomplished by restoration of some of the top-level constraints on the system hierarchy. A third environmental change currently under way is the incipient global warming and climate change. This has enormous implications for the northern range and the entire park. Two major general circulation models (GCMs)—the British Hadley Centre Circulation Model 2 and the Canadian Coupled General Circulation Model 1—project mean annual temperature increase of 3.6°C (6.5°F) and 6.3–6.5°C (11.3–11.7°F) by 2080–2100 in the Intermountain West of the United States (Wagner 2003). They project annual precipitation increases of 54– 184% (Wagner 2003). Two recent analyses of twentieth-century weather records in the region show that these trends had already begun during the 1900s. Annual, average temperatures in the northern Rockies rose 0.6°C (1.1°F) (Baldwin 2003) during the 1900s and annual average monthly minimum temperatures rose 0.859°C (1.55°F) in the same area (T. G. F. Kittel, unpublished data). Annual precipitation in the northern Rockies rose 6% in the region (Baldwin 2003), and summer precipitation rose 29.5% (Kittel et al. 2002). The changes are likely to affect every aspect of the physical and biotic resources of the region. Two Canadian climatologists predict zero snowpacks in the northern Rockies by 2070 (Fyfe and Flato 1999). In fact, the area occupied by glaciers today in Glacier National Park is only a third of the area of glaciers at the time of that park’s establishment in 1910. The glaciers are predicted to disappear completely in about 30 years. That alpine snowpacks are beginning to melt earlier in the spring is reflected in a 10-day advance in the run-off peaks of several western streams (Baldwin et al. 2003). If winter snowpacks no longer form (i.e., if snow changes to rain) at the YNP higher elevations, the ungulates may remain on what is now summer range throughout the year. The immediate effect would be to ease the herbivorous pressures now applied to the northern range. But the same changes would likely occur with all of the YNP elk herds that now summer at the park’s higher elevations and winter at lower areas. At the same time the current constraints of winter weather and winter forage exhaustion would ease and allow population increase that could eventually impact the entire park. The climate changes would also be likely to stimulate wide spatial shifts in vegetation types (Romme and Turner 1991). With temperature increases alone, forested zones would be expected to move upslope and reduce or eliminate alpine zones. The whitebark pine zone, a species important to grizzly bear, would be expected to shrink or be eliminated. With increase in both temperature and precipitation, the coniferous zone would be expected to expand into the lower elevation shrub steppe (Reiners 2003) and significantly change the character of the northern range. In total, what appears to be imminent climate change would be likely to alter the ecology of the northern range profoundly. There is therefore an urgent need for extensive monitoring and well-designed research to provide an understanding of these changes and a knowledge base for future management policies.

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THE NATURAL-REGULATION ECOLOGICAL HYPOTHESIS TESTED I stated in chapter 1 that a secondary purpose of this book (after marshaling the evidence on the truth of the northern range situation) is to test the naturalregulation ecological hypothesis. In fact, it has already been implicitly tested by many of the studies cited in the previous chapters and explicitly tested by several previous authors. Kay’s (1990) massive study falsified most of the projections on vegetation effects and competitive exclusion of other herbivores. In his closing chapter, he concludes, “Since, I can find no evidence to support any of these hypotheses and all of the available data support the opposite conclusions, I feel compelled to reject the entire ‘natural regulation’ paradigm.” Boyce (1991) begins a discussion of natural regulation by stating “existing evidence suggests that each of these premises has been violated, and therefore, one might think that the natural regulation hypothesis should be rejected.” He next comments that each of the “premises”—for example, assuming that an ungulate population would not affect the vegetation—“is inappropriate or misleading.” By some logical turn that I don’t follow, he then concludes that because the premises supporting the hypothesis are inappropriate or misleading, “no basis exists for rejecting the natural-regulation hypothesis.” Boyce complicates the discourse further by adding such value considerations as “how many elk should be in the . . . Park” (emphasis added) and whether the “consequences of high elk numbers appear to be socially unacceptable.” This is an excellent example of the problem discussed in chapter 1, the commingling of the natural-regulation ecological hypothesis, a scientific question, with the naturalregulation management policy, a value question. What Boyce calls premises are in fact the several predictions of the hypothesis and the tenets to be tested. That he considers them “inappropriate” or “misleading” is his concession that they were improbable when they were posed, as judged by the accumulated ecological evidence and theory of the time and the existing evidence on the northern range. Subsequent research has falsified them and the hypothesis as a whole, and has confirmed the a priori probability that he conceded. Singer et al. (1998a) state: “We conclude that the YNP natural-regulation management model was internally inconsistent because . . . [it] predicted almost no [vegetation] changes would occur.” The authors also critique the hypothesis for underestimating the significance of predators. I assume that the intent here is a comment on the ecological hypothesis, but it is another example of failure to make the above distinction. The NRC study (Klein et al. 2002) also implicitly falsified the hypothesis. In the statement “vegetation changes observed in the past 130 years or so appear to have been influenced more by ungulate browsing than by climate change,” The panel in essence challenged the no-vegetation-effect tenets of the hypothesis.

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The wolf reestablishment may provide yet one more test. If the wolves succeed in significantly reducing the northern herd, and the vegetation responds by returning toward pre-1872 conditions, the hypothesis will have been falsified. As already discussed, some of those vegetation changes may already be occurring. I listed the tenets of the hypothesis in chapter 1, and there is no need to repeat them here. On the basis of the evidence presented in the preceding chapters, I agree with the authors that all have now been falsified except the one predicting that the elk herd would equilibrate. But as mentioned in chapter 2, this was essentially a foregone conclusion. Moreover, equilibration occurred at higher levels than Houston’s (1971, 1974) predictions. Furthermore, it has occurred significantly through usurpation of and effects on the forage resource, rather than through some self-regulatory mechanisms that would prevent vegetation effects, as implied by the hypothesis. Contrary to the hypothesis, predation has evidently played a significant role, as Singer et al. (1997b) have shown. Beyond these population questions, the hypothesis had in essence been tested before it was posed in 1971. There was already an abundance of northern range evidence before that date attesting to its improbability. That the ensuing 33 years of research have falsified the hypothesis is thus not surprising although the park position has come around to this realization only gradually and haltingly, and even today reservedly.

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The Role of Science in Policy Process

IV The fundamental purpose of the said parks . . . is to conserve the scenery and the natural and historic objects and the wild life therein and to provide for the enjoyment of the same in such manner and by such means as will leave them unimpaired for the enjoyment of future generations. —National Park Organic Act of 1916 We should be thinking in terms of what will be here for the 22nd century and the 23rd. We will have dishonored our legacy if we are not prepared to protect it, preserve it, and pass it on to succeeding generations. —Robert G. Stanton


Why the Science Missed the Mark

16 General understanding of the system is a necessary base for any type of constructive interaction with the system. The past is full of examples of misunderstandings because of ignorance of a system’s behavior. —Bengt-Owe Jansson and Harald Velner The ability of the bureaucracy . . . to control information and resist change seems to show a level of individual and group ingenuity and persistence that reflects conscious control by dedicated and intelligent individuals as well as the unconscious part of the organization and culture of bureaucracies. —L. H. Gunderson, C. S. Holling, and S. S. Light

INTRODUCTION As stated in the Preface, the second purpose of this book is to explore the relevance of the northern range saga to the broader question of the role of science in policy setting. It arises out of my conviction that effective environmental and natural-resources policies can only be set in an environment of thorough scientific understanding. Science does not set policy but rather is a service to policy setting by enlightening it with facts that clarify the consequences of alternative policy options. Policy setting for public resources is itself a political process, and policies may be set that are contrary to scientific predictions of negative consequences. As Sabatier (1978) states, “No policy decision can be based solely on technical information. Normative elements invariably enter.”

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THE ROLE OF SCIENCE IN POLICY PROCESS

The natural-regulation policy is an excellent example. A half century of research had converged on the view that the northern herd had increased to such high levels that it was significantly altering the northern range ecosystem and that these effects could only be reversed by reducing the herd size. When the natural-regulation policy was adopted in 1967, it was contrary to that science as the quote from the W. H. Barmore memorandum in chapter 1 points out. It was, as most observers including then NPS Director Harzog agree, set in response to political pressure. But if policies are not set in a factual environment, they can only be set on the basis of tradition, speculation, and political forces; there may be little if any prospect of accurately predicting their consequences. Moreover, if there is open, scientific understanding surrounding policy decisions, and they are made despite indications of unfavorable consequences, it becomes incumbent on the process to explain the rationales for the decisions, and there can be no pseudoscientific rationales for the actions. Thus the critical role of accurate and objective research is to provide a factual environment surrounding policy process. Clearly, I find the weight of the evidence since 1970 failing to support much (if not most), of the park-supported scientific inference on the northern range issue. Given the overall importance of science in policy process, the fact that the northern range science is invoked to support park policy, and the expenditure of millions of taxpayer dollars on the northern range research, it is desirable to explore why the science has missed the mark. Aldo Leopold (personal communication) once commented that a profession is a group of people who demand higher standards of themselves than do their clients. To philosopher Kristin Shrader-Frechette (1993), scientists have an ethical obligation to call attention to faulty scientific practice. But first it is useful to sketch a brief history of the structure of northern range research. Despite the recommendations of 12 reviews, the NPS has never developed a formal research arm (Risser et al. 1992) as did the USDA Forest Service and the U.S. Fish and Wildlife Service early in their histories. Prior to 1993, research in the Park Service was administered out of 10 regional offices or in some individual parks. An NRC review stated, “there is not one NPS science program, but 10 separate programs, each different in form or function.” Why the NPS failed to respond to these recommendations has been discussed at length by Risser et al. (1992), Wright (1992), Wagner et al. (1995a), and Sellars (1997) and need not be repeated here. Yellowstone is one of the parks that has administered its own research. Its first biologist, M. P. Skinner, began studies on the northern range in the 1920s and was publishing in the technical literature before the end of that decade. The park continued with one or two scientists for the next few decades, gradually increasing to three and a supervisory research biologist by the 1960s. Thus northern range research up through the 1970s and early 1980s was conducted largely by NPS scientists based in and supervised by the park with some participation by researchers from other federal agencies. As an example, John and Frank Craighead, operating out of the Cooperative Wildlife Research Unit administered


by the U.S. Fish and Wildlife Service at (then) Montana State University, conducted research on grizzly bears between 1959 and 1970. In 1986, park research entered a new phase when Congress appropriated major funding for it to research the question “Is the northern range overgrazed?” New scientists were added including transfers from other parks. At the same time, major funding was outsourced to university researchers. By 1989, Singer reported that “40 separate research projects” were addressing the overgrazing question (Singer 1989). This expanded research effort continued with support provided to study the effects of the 1988 fires and subsequently preparation of a massive environmental impact statement on the proposed wolf reintroduction. By 1992 there were more than a dozen park scientists plus support staff. In October 1993, park research underwent an unsettling change (Wagner 1999a). Newly appointed Secretary of the Interior Bruce Babbitt removed research from the Fish and Wildlife Service and NPS and combined it into a new, independent Interior agency, the National Biological Survey (NBS). After a change of name, it finally became the Biological Resources Division (BRD) posted in the U.S. Geological Survey. A major purpose of the move was to distance research from management and policy setting and enable it to function in a more objective environment uninfluenced by pressures to conform to policy. The structure was to provide a research service to the parent agencies and at the same time reduce duplication that might have resulted from separate research programs in the individual parks, wildlife refuges, and other agency functions. Former Yellowstone biologists moved to regional research centers in Bozeman, Montana, Fort Collins, Colorado, and in some cases more distant locations. Although some continued to carry out northern range research, others diverged into studies unrelated to YNP. Movement of the scientists into NBS and ultimately BRD deprived the parks of the technical advisory and consultative roles they had played in management and policy matters while in the parks. It also weakened the parks’ influence over BRD research agendas. Many observers predicted that the parks would replace (“backfill”) the departed scientists with their own new personnel. This has been Yellowstone’s pattern, first calling the new hires resource managers, and renaming the research director the director of the Yellowstone Center for Resources. But with the change of presidential administrations and Department of the Interior secretaries in 2000, the park is now simply rebuilding a large in-house research program. Along with private funding, the park is developing a Yellowstone Ecological Research Center, including construction of a major research building. At the same time, a considerable amount of research is being carried out by extramural academic researchers who seek funding from other granting agencies but with logistic support from park personnel and facilities. Thus the northern range research has always been autonomous and administered at the park level in an agency that has not had a strong research commitment or central direction. The research has gone through shifting emphases,

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personnel, and funding; it has been associated with changing management policies. All of this could hardly have contributed to program stability.

WHY THE PARK MISSED THE MARK ON THE NATURAL-REGULATION HYPOTHESIS Problematic Research Procedures The failure of park-supported science to converge on a northern range paradigm that is in concert with the prevailing evidence stems from a complex of causes. One is improper scientific procedures, an example being the selective use of evidence and refusal to consider contrary views of outside professionals. As Wagner et al. (1995a:51–52) pointed out there were numerous critics in the scientific community of the natural-regulation hypothesis in the 1960s and 1970s. But Chase (1986) most forcefully pointed out that Houston (1971, 1982) dismissed a century of observers’ accounts and a half century of research in propounding his natural-regulation hypothesis and the synthesis of his book. Another case of the selective use of evidence has been the park investigators’ refusal to consider the inferences of low elk numbers in prehistory and of elk migration out of what is now the northern range in winter. They have ignored the accounts of early observers and park officials, and summarily rejected Kay’s (1990) and other evidence, discussed in chapter 3, of low elk numbers prior to park establishment. In this same genre has been the post-1970 succession of park attitudes on elk effects. The first position was that there had been no significant change in the northern range during the 1900s (Houston 1982; Despain et al. 1986). When outside critics pointed to obvious evidence of change, the park then posed explanations other than elk grazing and browsing for the change (e.g., climate change, fire suppression, decline of beaver). When outsiders then pointed to evidence of browsing effects, the park posed hypotheses designed somehow to discount elk effects and avoid a straightforward and obvious implication of an elk herd vastly increased above low numbers prior to 1872. Rather, they proposed increased susceptibility to browsing due to climate change and/or reduced production of secondary compounds. Scientists will of course differ on the interpretation of evidence. But objective researchers weigh all the evidence, consider alternative inferences, and explain why they opt for any one. Park science since 1970 has largely taken as an a priori premise the Houston (1982) interpretation of high elk numbers in prehistory and fashioned its inferences to conform to this premise. A second form of problematic science has been the tendency to pose hypotheses when there is no supporting evidence. Conspicuous examples are the hypotheses on bison. I discussed in chapter 9 the park generalization that the bison population was at an upper equilibrium during the 1970s, but then began rapid increase in the 1980s and 1990s when winter roads were groomed to


accommodate snowmobile travel. I showed that the most elementary population analysis of the park’s own census data disclosed a density-dependent pattern of population growth with an average, annual increase of 12% in the 1970s declining to 3% in the 1980s and ’90s. The herd was neither at equilibrium in the 1970s, nor was the increase rate higher in the latter decades. Moreover, subsequent extramural behavior observations showed that the animals were not using the roads to any degree, the imputed cause of the alleged reduction in winter mortality. The park statements had simply been made without recourse to its own data, or without data to support them. Another example has been the treatment of willows. Early photographs showed robust willow growth in moist areas of the northern range when, according to Houston’s paradigm, elk were claimed to be abundant. All observers agree that willows were heavily hedged during the 1900s. Contemporary willows were hypothesized to produce lower levels of secondary compounds than did willows in prehistory. But there was no way of testing the hypothesis because there were obviously no measures of secondary compounds in pre-1872 willows. And there had been no elk feeding trials or field observations to test whether elk, hard-pressed for winter food, were at all deterred from browsing by the secondary compounds contained in the species. Despite the purely hypothetical nature of this idea, park public-information releases touted it as “the most probable” explanation of the difference in willow condition between prehistory and the twentieth century. Science does proceed by formulating and testing hypotheses. But it is often true that numerous hypotheses can be proposed to explain a given phenomenon. If researchers do not refrain from publicizing hypotheses except those for which there is some preliminary evidence or solid implication of theory, the scientific literature can become saturated with unsupported and in most cases erroneous ideas. Top journals will not accept articles with hypotheses that are not supported by at least some evidence. A third problem with park scientific practice has been lack of care in the preparation of some publications. In an article in a mainstream journal, the last name of the second author in the byline was misspelled and the middle initial of the third author was incorrect. There were 31 typos in the 8-page paper, as well as 7 misspellings of author and journal names in the bibliography, and table and figure captions that were either contradictory or not understandable. One wonders where editors and peer reviewers were, and one must conclude that the authors did not read proof. In a manuscript submitted to the editors of a conference proceedings, authors were misquoted; unequivocal statements were made without acknowledging contrary evidence and inferences of other authors; figures were cited in the text as references to statements made that either treated data unrelated to those statements or did not support the statements. In pp. 20–40 as a sample of the 50-page manuscript, 46 of the references cited were not in the bibliography. Again, table and figure titles and captions were unclear or not understandable. Synoptic statistics and graphs summarized data that had not been previously

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published, and hence subjected to peer review, nor were the methodologies in obtaining the data even described. The manuscript was fraught with other problems. There are other less extreme cases. No one is free of errors. But the density of errors in the cited cases indicates an extreme lack of care that makes one automatically skeptical about the care with which original research data are obtained and analyzed, and hence the value of the science.

Research Administration The administration of a program has final responsibility for its quality. The problems inevitably reflect inadequate standards and oversight by the scientists’ supervisors. A 1989 audit by a Department of the Interior assistant inspector general for audits (Bloom 1989) found that no study plans had been prepared for 23 of 41 research projects in Yellowstone. Plans for the remaining 18 were “deficient with respect to content.” The primary mission of an agency research program is to provide it a body of facts with which it can evaluate ongoing and considered management and policy decisions. To serve the agency effectively, the research must be accurate and objective. As controversial as the northern range issue has been, it has surely been the responsibility of YNP’s research supervisors to call in the critics and seriously consider their views and evidence. Although there have been occasional Yellowstone conferences at which both agency and extramural researchers have presented papers on their work, there have been no small, focus-group discussions comprised of individuals with contrary points of view who could discuss and come to some clear understanding of the bases for their differences, if not resolution. Views contrary to park positions have been looked on as threats and avoided. The difficulties with administration of northern range research have not been limited to problems of omission. There is considerable evidence indicating that the park has managed the research to ensure that it did not produce conclusions critical of the natural-regulation policy. The first instance began in the latter 1960s with proclamation of the policy. William J. Barmore Jr. transferred from his position in the USDA Forest Service to YNP, and the NPS in 1962 to become the principal investigator on the northern range. Between 1962 and 1967, he got only excellent performance ratings and was promoted swiftly from GS-7 to GS-12 rank. In 1967, Barmore was reassigned to the office of Natural Science Studies under Supervisory Research Biologist Glen Cole, who was newly assigned to Yellowstone and who authored the naturalregulation policy in the same year (Anonymous 1967a, b). Within his second year, Cole gave Barmore an unsatisfactory performance rating and attempted to have him removed. Cole’s superior, chief scientist for the NPS, did not concur with Cole’s unsatisfactory rating of Barmore, and his position was sustained. Barmore had pointed out to Cole in 1968 that the new policy was contrary to previous research (see quote in chapter 1). He believes that Cole wanted to


replace him with Douglas Houston, who had assisted Cole previously in Teton National Park and who later became author of the natural-regulation ecological hypothesis. Unbeknownst to Barmore, Houston had already been retained and was working on the northern range by 1970 when Barmore left the park for assignment elsewhere. Events surrounding grizzly bear research are well-chronicled cases of park control over its science and scientists. In 1970, John and Frank Craighead, who had conducted 11 years of research on Yellowstone grizzlies and were undoubtedly the most knowledgeable scientists about the species’ ecology in the park, were forbidden to conduct further research there. The Craigheads were concerned over the effects of policy changes on the bear population and challenged the park’s assessment of those effects (Mattson and Craighead 1994; Craighead et al. 1995). In 1983, R. B. Finley Jr. charged the park with refusing to release data on grizzlies, which were needed by a member of the Interagency Grizzly Bear Study Team to examine the status of the population. Commenting on his experience with the team, he stated (Finley 1983): “It quickly became apparent that the team leader was expected to run the team in a way that gave support to the policies of Yellowstone National Park. . . . the Department of Interior was more interested in suppressing disagreement than in finding the truth or saving grizzlies.” In 1993, the data files of Yellowstone bear biologist David Mattson were removed and placed under lock and key and his computer disk was erased by his superior, who also ordered the office secretary to open his mail. Mattson was subsequently transferred to another state. These actions occurred, according to Mattson, because he had disagreed with his superior over inferences drawn from park data on grizzly population trends (Mattson and Craighead 1994; Wilkinson 1998). The controversy over natural regulation has elicited similar administrative and personnel actions. In gathering data for his book that was critical of the policy, Chase (1986) found it necessary to file Freedom of Information Act requests to obtain park information because his previous requests had been ignored. In 1993, park biologist Richard Keigley’s manuscript on northern range cottonwoods was held up for several months until he removed text pointing out the implications of his findings for the policy (Clifford 1993). He was later ordered to discontinue his innovative architecture research (see chapter 10), which promised to reconstruct the history of ungulate use on other northern range woody species. Keigley had forcefully asserted, on the basis of his evidence, that park management policies were significantly affecting park resources, contrary to claims of park administration. Transfer of Yellowstone biologists to the NBS in October 1993 administratively removed them from supervision by the park. But YNP administrators still control park science by discretionary issuance of permits to conduct research in the park. In 1995 Keigley, now an NBS biologist, was denied a permit to conduct research on other woody species in the park. In 1997, he was forbidden to respond to a request from a congressional committee to testify on Yellowstone

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science, and the committee was obliged to subpoena him. Because Keigley was now in BRD in the U.S. Geological Survey (USGS), this denial must have moved up the NPS chain from YNP to the Interior Secretary and back down the USGS chain of command. Keigley’s most recent request for a permit to photograph willows in the park has received no response. In July 1999, a well-known ecologist from a major university told me, “I want to do research in the park, so I don’t dare get them mad at me or they won’t give me a permit.” Another equally noted ecologist from another university, who was conducting research in the park, told me in September 1997, “We have to be careful what we say [in our publications].” Another means of controlling the science output is the issuance of sole-source contracts on northern range research. I have been told of one case by the recipient of such a contract. Another case was reported to me by a midlevel NPS administrator who saw the paperwork. I learned of yet another from a university faculty member who had discussed the matter with the contract recipient. There are rumors of others.

Proximity to the Public Information Program Yellowstone has an aggressive public-information program that reports on park amenities, activities, and developments. Information is extended through a variety of media: semi-technical books; a quarterly journal, Yellowstone Science; popularized annual summaries of the previous years’ events; releases to the news media; brief tabloids distributed to tourists as they enter the park; and naturalist lectures in the interpretive program for tourists. These of course contribute to the park’s important education function. The northern range issue is periodically treated in these outlets by nonscientists whose primary responsibility is to publicize the virtues of the park and its programs. They have neither the scientists’ credentials nor the responsibility to evaluate critically and objectively the strengths and weaknesses of the research methodology and data and inferences drawn therefrom and to report these in carefully qualified statements. Hence the natural-regulation hypothesis and management policy have consistently been portrayed in the public-information outlets in a positive (if not glowing) light. The research and public-information personnel have associated closely, both professionally and personally. In a number of cases they have coauthored publications, and they have regularly reviewed each others’ publications. The net effect, along with the administrative pressures, must be to create a climate that is not conducive to the independence and objectivity needed for quality science.

RESOLUTION In a management agency, research is a service to management and policy setting. Its function should be to provide an accurate and objective environment of


facts within which it can (1) evaluate and predict the consequences of considered and current policy alternatives, (2) assist in the design of management programs employed to achieve policy goals, and (3) evaluate the degree to which existing policy and management are achieving the goals. It also has the responsibility to evaluate itself in terms of the effectiveness with which it is providing this service and the quality of its operations. It seems clear that YNP research since 1970 has not provided an accurate and objective environment for the reasons just discussed. Portrayal of the ecological effects of the several management policies—pre-1872, early park protection, herd control, and natural regulation—over the past 37 years has in each case been erroneous to varying degrees, as the analyses of preceding chapters have shown. The solution must be in separating research from administration by the park, and placing it in an independent administrative structure where it is removed from the policy advocacy of the park administration and public-relations functions. Prior to transfer to the NBS, such a move had been urged by repeated reviews of NPS science. A scathing NRC review (Risser et a1.1992) recommended: “The National Park Service should revise its organizational structure to elevate and give substantial organizational and budgetary autonomy to the science program.” The review further recommended appointment of a chief scientist based in Washington who would report to the NPS director and supervisory scientists in the research administrative line posted in the regional offices. Similar recommendations have also emerged from NPS ranks. Based on his research experience in Mammoth Cave National Park, Alexander (1996) commented: “Perhaps the most important lesson to be learned from the history of NPS-supported research at Mammoth Cave is that research at any park should be supervised by the regional chief scientist, not the park superintendent.” However, the solution has now become bureaucratically complicated by movement of what was the national park research into the BRD of the USGS. What was NPS research is no longer in the NPS. Thus the NPS director has no prerogative of returning the research into and molding a Park Service research arm, even if he or she were convinced of the importance of doing so. Such a move could only be made by the Secretary of the Interior, a political appointee who might or might not have the understanding and commitment to the importance of such a move. Moreover, the personnel backfilling in Yellowstone poses a further obstacle to returning NBS scientists to research in the park, and under the aegis of a new NPS research arm. Such a move would duplicate the new personnel at a time when funding is tight. The net effect of all these moves, along with the park’s discretionary issuance of research permits, has been to strengthen park control over its research. BRD, now marginalized, considers that it is serving “clients” or “customers,” and is therefore in no position to produce research that would be critical of park science and policies. From the standpoint of developing a sound research program

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in the NPS, the transfer to NBS and ultimately BRD must now be looked on as a most unfortunate administrative action. Providing national-park research the independence that it needs to produce objective research that lets the chips fall where they may would take forceful and courageous administrative action at the highest levels in the Department of the Interior. One remaining hope lies with changes in policy-setting procedures, to be discussed in the next chapter.

The Science–Policy Interface

17 I never forget that I live in a house owned by all the American people and that I have been given their trust. —Franklin D. Roosevelt [The Secretary of the Interior will] provide for the preservation, from injury or spoliation, of all timber, mineral deposits, natural curiosities, or wonders within said park, and their retention in their natural condition. —1872 Yellowstone Act

CREDO The domain of this book is the science of the northern range of Yellowstone National Park. I do not intend to advocate a management policy for that system, nor to pass value judgment on the natural-regulation policy. Although I have as a matter of principle repeatedly advocated the separation of science and policy advocacy in natural-resources issues (Wagner 1996b, 1999a,b,c,, 2001), a number of authors have inferred from things I have written that I propose a management policy for the northern range and more generally for national parks. To use an ecological understanding of the structure and function of an ecosystem as a basis for predicting or evaluating the effects of policy options on the system does not of itself constitute policy advocacy. Nor is pointing out the management measures implied by certain policy goals a prescription of those measures (see Huff 1997a). The ecological effects of the naturalregulation policy on the YNP northern range are a scientific question, and analyzing and portraying these effects does not imply value judgment. Similarly, to 317

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critique some of the science used to support the natural-regulation hypothesis, as I have done in the previous chapters, does not imply criticism of the policy. My intentions in this final chapter are to address two aspects of policy that relate to and are clarified by the science of the northern range. One is to explore policy-setting mechanisms that could promote scientific objectivity. The other is to use the understanding that we now have of the northern range to clarify the ecological implications of proposed park goal and management options. Policy setting is a complex sociopolitical procedure that merits analysis and debate in its own right, and I touch only on certain of its aspects that relate to scientific understanding of the northern range.

POLICY SETTING AND SCIENTIFIC FREEDOM What Is Policy? The several dictionaries around me converge on a definition of the word policy as a statement or stated plan of how an organization will operate to achieve some goal. Thus it is a stated plan of a course of action. Much of the discourse on policies for natural resources in national parks focuses on the pros and cons of controlling animal populations, nonnative species, insect outbreaks, and fires; restoring missing species; and use of controlled burns. But these are management procedures. Hence the discourse addresses the means rather than the ends, which are the goals. A number of authors have converged on the view that public resources are managed not for the resources themselves but to satisfy societal values (Hendee 1974; Giles 1978; Kania and Conover 1991; Wagner 1999b, Wagner et al. 1995a; Kennedy and Thomas 1995). This is expressed for Yellowstone by Brussard (1991a): “Before any progress whatsoever can be made, the Greater Yellowstone Ecosystem’s residents must decide what their goals for the area are.” Huff and Varley (1999) make the point very well for the northern range: “The ultimate question that the Park Service will have to answer is whether or not natural regulation [policy] will best serve the local, national, and global constituencies of Yellowstone.” Thus, management goals for public resources are, in the broadest sense, the satisfaction of societal values, and an oversimplified model of the policy-setting process is: Societal Values → Goals → Policies → Management The implications of this model are that societal values are the basis of the whole process. Goals are articulated to satisfy those values. Policies prescribe management programs to achieve the goals, and thereby satisfy the values. Science is not part of the direct causal sequence that sets policy, but it has the indispensable role of factually illuminating every step of the procedure. The


social sciences portray the value profiles of the affected publics and the social, political, and economic implications of alternative goals, policies, and management options. The natural sciences similarly evaluate the biological and physical implications of those options, assist in the design of management programs, and evaluate how well they achieve the goals. Thus, adding the role of science to the above schematic produces:

Policy-Setting Procedures Conducive to Research Objectivity The questions follow as to what procedures for deciding on goals and policies in the parks are most likely to promote research objectivity and who should make those decisions. This is perhaps best discussed in the historical context of policy setting in the natural-resources agencies. The federal land-management agencies were established in the late 1800s and early 1900s with specific charges: the USDA Forest Service to protect and administer use of timber resources on the national forests; the Bureau of Land Management to oversee private livestock grazing and mining on the unappropriated public domain; the Fish and Wildlife Service to protect, control, and facilitate use of wildlife resources not under jurisdiction of the states; and the National Park Service to preserve remnants of the scenic, natural, paleontological, archaeological, and historic heritage of the nation (Nelson 1995a, b). The prevailing political philosophy of the period was progressivism, conspicuously advocated by Theodore Roosevelt (Nelson 1995b). In the natural resources agencies this philosophy was manifested in policy setting on the premise that the technically trained agency people knew most about the complexities of the resources. Hence they were in the best position to formulate management policies and did so internally within the agencies (Nelson 1995a). Motivated by their agencies’ historic missions and their professional training in resource extraction, agency professionals prescribed policies advocating efficient, sciencebased harvesting of resources. With the nation largely rural and agricultural, these policies were consistent with the societal values of the time. As the nation changed to become largely urban during the twentieth century, societal values toward natural resources and the outdoors shifted away from those of the agencies. Concurrently, the science of ecology developed a deeper and more comprehensive understanding of natural ecosystems and the affects of commodity extraction on them. Consequently, the public became less accepting of the agencies’ stewardship of natural resources. The result was a shift in political pressures that precipitated legislative and administrative changes away from progressive-style, internal policy setting. The

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Forest and Rangeland Renewable Resources Planning Act of 1974 ordered the Forest Service to engage in long-term planning for each of its national forests. The National Forest Management Act of 1976 directed that agency to include representatives of all concerned public interests in the planning process. Thus national forest policies are now being set by combinations of interested parties. No such legislation has been passed for the BLM, but the agency has for some years followed planning procedures that bring together representatives of interest groups—ranchers, wildlife groups, environmental organizations, and wild-horse advocacy groups—who deliberate on how many animals of each class are to be carried on the public domain lands. The procedure is called coordinated resources management planning. Moreover, in 1996 Secretary of the Interior Bruce Babbitt further instituted public involvement in BLM policy setting by establishing Resource Advisory Councils (RACs) in each state to recommend policies for the agency. The RACs have balanced memberships of individuals representing all groups concerned with management of lands under BLM aegis. Thus, policy setting for the national forests and public domain lands has changed from almost complete discretion by Forest Service and BLM professionals during the progressive era to procedures that today are more responsive to the full range of societal values. In the political science lexicon, these procedures are called “interest-group pluralism.” I dwell on this matter because the literature on the role of research in policy making describes a number of ways in which the policy positions of agencies color research objectivity (Wagner 1989). These range from organizational loyalty on the part of researchers who find it uncomfortable to report research results that question their organization’s policies, to overt coercion of scientists by administrative higher-ups, as in the cases described by Wilkinson (1998). Freeing executive branch agencies from the authorship and prescription of policies through some form of interest-group pluralism reduces their sense of ownership of the policies, their tendency to defend them, and thus their resistance to contrary research results. A national forest supervisor recently remarked in my presence, “I no longer set policies. I just carry out what I am told by others.” Thus, the other agencies have followed two administrative strategies for promoting scientific freedom: administrative separation, as discussed in the last chapter, and policy setting by interest-group pluralism that frees them of ownership of the policies. The NPS has not had the same history of legislative changes as the other agencies. Unlike the Forest Service and the BLM, the NPS does not for the most part allow commodity extraction on the lands it administers. Consequently it has not been subjected to the same commodity-oriented political conflicts experienced by the other two agencies and hence the industrial and national pressures to alter procedures. Rather, pressures tend to focus locally on individual parks. Sensitive issues include economic concerns of gateway communities, activities of park concessionaires, and the concerns of local wildlife and environmental groups (Wagner


et al. 1995a). Thus the pressure groups tend to be locally based and unique to each park and less likely to form regional and national power blocks, like the timber industry in connection with Western national forests and the livestock and mining industries in connection with BLM lands. As a result, the NPS has not been the subject of national legislation mandating public involvement in policy making like the National Forest Management Act (NFMA). And although the NPS is required to follow the broader environmental legislation prescribing public participation in such matters as the National Environmental Policy Act and the Endangered Species Act, it has not had legislated or executively ordered formal protocols for public involvement like the NFMA and RACs. With this different history, policy setting for the parks within the NPS historically has been highly decentralized and substantially at the discretion of the superintendents, a legacy of the Progressive Era (Wagner 1999b, 2001). The NPS’s 1988 policy manual (Anonymous 1988) stated, “servicewide policy will be articulated by the Director of the National Park Service. . . . Park-specific instructions, procedures, directives . . . may be set by superintendents within formal delegations of authority.” At the same time, public involvement in decision making has been variable and discretionary. However, numerous authors both inside and outside the NPS have urged greater public involvement in NPS policy setting (Bishop et al. 1989; Risser and Lubchenco 1992; Tuler and Webler 2000). I have suggested that new legislation on the order of the NFMA might be needed to formalize public participation in park planning (Wagner 1999b, 2001). Perhaps in response to such recommendations, the NPS has in recent years strengthened its commitment to public involvement in policy setting. Its new policy manual (Anonymous 2000) prescribes: “Public participation in planning and decision making will ensure that the Park fully understands and considers the public’s interests in the parks. . . . Members of the public . . . will be encouraged to participate during the preparation of a . . . [park general management plan] and the associated environmental analysis.” Whether this will be routinely followed in Yellowstone and other parks without a formal statutory mandate remains to be seen. As this is written, Rocky Mountain National Park (RMNP) is using the environmental impact statement (EIS) process to gather public participation in developing an elk and vegetation management plan. RMNP has an elk situation similar to that of Yellowstone’s northern range: the elk population has more than tripled since 1969, and concentrations and migration patterns are also outside the range of variation under natural conditions. As a result, willow and aspen stands no longer regenerate effectively, depriving other wildlife of the food and habitat they need to survive. . . . Recovering, to the extent possible, the natural range of variability in the elk population and affected plant communities requires identifying and implementing measures to maintain, restore, and protect the inherent integrity of natural resources. (Anonymous 2004)

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The park superintendent invited the public (Baker 2002) in summer 2003 to participate in planning for management. It held a series of public workshops in summer and fall 2003 and solicited public comment. Along with this input and collaboration with local, state, and other federal agencies, it developed a list of five elk management alternatives that include no action, culling, fertility control, wolf reintroduction, and combinations thereof. These were placed before public meetings for reactions during summer 2004. A final plan will be developed by winter 2005–2006. As discussed in the last chapter, the administrative separation of research from the parks before its transfer to the NBS was variable but minimal. At most it was administered out of the regional offices. In those cases where it was administered in the parks, it was nonexistent. Thus the administrative strategies for ensuring scientific freedom in the parks have been weak. And although I do not suggest that research throughout the national park system has suffered objectivity problems—there clearly has been excellent research in many of the parks (see Halvorson and Davis 1996)—there do appear to have been problems in Yellowstone. Policies have been set internally: The decision to control the herd in the 1930s was made by park officials. While elk culling was stopped in 1969 by external political pressures, the naturalregulation policy was written and released by the park in 1967. The internal policy setting that generated the natural-regulation policy and associated pressures on its science have not served Yellowstone well. By setting a policy and posing a hypothesis that were contrary to available evidence and prevailing theory, and then doggedly defending them over time as more contrary evidence accrued, the park has damaged its credibility and professional image in professional circles, as commented in chapter 1. One reviewer of the manuscript for this book, anonymous to me, wrote Oxford University Press, “I think the situation in Yellowstone is a travesty. . . . the book will open the curtain for a debate and reexamination that is long long overdue and sorely needed.” Moreover, the policy and public relations path that Yellowstone has taken on the northern range could well increase the difficulties of changing directions in the future if change were ever to be contemplated. There are constituencies that favor the natural-regulation policy and accept the park’s pronouncements that it is well supported by the science. Their resolve in supporting the policy has undoubtedly been strengthened by park reassurances. If conditions on the northern range were to decline to some point where remedial action were deemed desirable, it could be more difficult to gain support of advocacy groups, which have been told for decades that all was well, than if they had been informed all along with valid scenarios of the true conditions. If political forces persisted in forcing management directions inimical to the resources, the park could shift the onus to those forces and disclaim culpability. I do not fault the park for yielding to irresistible political forces in adopting the no-control policy in 1967. But I do consider it unfortunate that it has closed its eyes to decades of prior evidence, and perpetuated an illusory paradigm to justify the policy in the face of contrary evidence.


The transfer of research to the NBS, and ultimately the Biological Resources Division, had promise for addressing the objectivity problem. But with Yellowstone rebuilding its own in-house research program and to the extent that it continues to set its own policies, administrative separation is not likely. The remaining hope is that it will adopt the strong public-involvement procedures prescribed in the new policy manual. If policies were set collaboratively by coalitions of interest groups in the Forest Service and BLM mold, it could reduce the park’s sense of authorship and ownership of the decisions, a felt need to defend them, and sensitivity to research results that point out problems with policy-directed management programs (Wagner 2001). Research could function with greater freedom.

EXPLICATION OF PROPOSED GOALS AND MANAGEMENT REGIMENS Need for Explicit Park Goals I have suggested (Wagner et al. 1995a; Wagner 1999b) that the goal component of the above policy model for the parks needs more explicit attention. It is not clear how effective management programs can be designed (and thus policies set) or their accomplishments judged without clearly articulated end points that they are prescribed to achieve. Numerous authors, both inside and outside the NPS, have made this same point (Bonnicksen and Stone 1982a, b; Foresta 1984; Johnson and Agee 1988; Bishop et al. 1989; Bonnicksen 1989). In an analogous example, Lackey (2004) stated that “many of restoration ecology’s socalled failures are due less to technical inadequacies than to a lack of straightforward and broadly accepted restoration goals.” Rogers (2003) commented, “An almost universal barrier to good management [in Kruger National Park] is poor translation of policy into achievable and scientifically defendable operational goals, supported by a process with which to audit their achievement.” Huff (1997a) took issue with the advocacy of explicit, park goal statements in Wagner et al. (1995a), but seemingly reversed himself in the close of his article by stating, “I suggest we start with some common-sense revisions to our Servicewide and park-specific policies, clearly iterate our purposes.” And the new NPS policy manual, Management Policies 2001 (Anonymous 2000), now prescribes: Managers will be held accountable for identifying and accomplishing measurable long-term goals. . . . Management prescriptions will . . . clearly define the desired natural . . . conditions . . . to be achieved and maintained over time. . . . Objective, measurable, long-term goals in the park strategic plan . . . will define the resource conditions . . . to be achieved in the near future. (emphasis added) Thus, national park policies for natural resources will now define measurable goals that state desired natural conditions. Such definitions clearly require ecologically explicit understanding of park ecosystems.

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A number of goals and management protocols have been proposed for the YNP northern range by individuals both inside and outside the NPS. Our understanding of the northern range ecosystem now makes it possible to examine the ecological implications of these proposals and perhaps stimulate a dialogue on what the desired natural condition of the northern range should be. In terms of the new policy directives, that dialogue should take place in a collaborative setting involving representatives of public-interest groups concerned about park policies, that is, interest-group pluralism. As stated at the beginning of this chapter, there is no intention here to advocate any of the proposed goals but only to explore their ecological implications. Attainment of some goals requires management procedures. To point these out does not imply recommending them.

Potential Goal Options for the Northern Range Preserving the “Natural” HISTORY OF THE NATURAL PRESERVATION GOAL. Undoubtedly the most frequently advocated goal for Yellowstone National Park, and by implication the northern range, has been preservation of the “natural” condition. It began with the very establishment of the park in the language of the enabling legislation quoted in the epigraph at the beginning of this chapter. It continued in the language of the National Park Organic Act of 1916: “to conserve the scenery and the natural and historic objects and the wildlife therein . . . by such means as will leave them unimpaired for the enjoyment of future generations.” This general intent, with slight modifications, has been reiterated periodically through ensuing decades. NPS historian Sellars (1997:29) comments that “the legislative history of the Organic Act provides no evidence that either Congress or those who lobbied for the act sought a mandate for an exacting preservation of natural conditions” (emphasis added). At that point in time, major concerns were for the preservation of scenery, economic benefits of tourism, and efficient management. But Pritchard (1999) traces a long history of the concept advocated by numerous individuals both inside and outside the NPS. A National Academy of Sciences review in 1963 (Ackerman et al. 1963) recommended that “The purpose of national parks should be the preservation of nature, the maintenance of natural conditions.” Probably best known is the recommendation by Secretary of the Interior Udall’s 1963 Advisory Committee, the Leopold Committee, that “A national park should represent a vignette of primitive America” (A. S.Leopold et al. 1963). That intent, with some nuanced variation, has continued to the present. The September 19, 1967, memorandum announcing management objectives for the YNP northern herd began by stating: “The primary purpose of Yellowstone National Park is to provide present and future visitors with the opportunity to see and appreciate the natural scenery and native plant and animal life as it occurred in primitive America” (Anonymous 1967c).


However, by this time focus began to shift to concern for both components and processes. Thus the 1967 memorandum stated that “Primitive scenes with a variety of wildlife species only result from allowing natural regulatory processes between plant and meat-eating animals to take place.” In the NPS’s 1978 policy manual (Anonymous 1978), the language stressed preservation of both natural objects and processes: “Management of park lands possessing significant natural features and values is concerned with ecological processes.” And the 1988 policy manual (Anonymous 1988) continued this theme and included concern for entire ecosystems: “Managers and resource specialists . . . will try to maintain all the components and processes of naturally evolving ecosystems.” By the latter 1980s, a number of observers shifted emphasis entirely to concern for preserving the processes. Because ecosystems change over time, they have pointed out that preservation of the state of a system at any point in time is contrary to preserving the “natural,” and that a focus on preserving the processes allows it to undergo natural change. Thus Yellowstone’s research chief stated (Varley 1988): “National parks were founded on the principle of natural processes. . . . This includes the perpetuation of natural processes essential to the existence of a healthy ecosystem.” And NPS biologists Huff and Varley (1999), writing on Yellowstone, stated that “Implementation of natural process management . . . remains the umbrella policy for most natural resource management issues.” They further consider their procedure synonymous with the naturalregulation policy, to be discussed shortly. More recently, emphasis seems to have moved back to a balanced concern for both components and processes, with frequent emphasis in the new policy manual (Anonymous 2000) for preserving the systems unimpaired. This is undoubtedly a reference to the wording in the Organic Act: The Service will try to maintain all the components and processes of naturally evolving park ecosystems, including the natural abundance, diversity, and genetic and ecological integrity of the plant and animal species native to those ecosystems. . . . The Service will re-establish natural functions and processes in human-disturbed components of natural systems in parks . . . [including] the natural abundances, diversities, dynamics, distributions, habitats, and behaviors of native plant and animal populations and the communities and ecosystems in which they occur. WHAT IS NATURAL? These terms and concepts (e.g., natural, pristine, natural process management, natural regulation) have generally acknowledged ambiguities. It is not clear how they can be translated into the objective and measurable goals for the northern range, which the policy manual calls for, without resolution of these uncertainties. Hence it seems useful to explore these questions in the context of what we know about the northern range to contribute to discussions held in the collaborative mode, prescribed by the policy manual, on setting goals and management policies for the northern range. Wagner et al.

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(1995a) explored these ambiguities in detail. But it is useful to touch on them here in light of newer documentation and research findings. Two major questions contribute to the ambiguity of a goal for preserving the “natural” on the northern range. One is the question of chronology: preservation of the natural at what point in time? As the NRC report (Klein et al. 2002) discussed, the area that we now call the northern range has been changing continuously through the Earth’s history, including tectonic, solar, and glacial changes. Though certainly valid, these are not meaningful time scales for deliberating on park goals relevant to lifetimes of the American public. A more pertinent time scale is U.S. history, and a key date is the timing of the Yellowstone Act. The congressional intent at the time it was passed in 1872 must surely have been preservation of the natural wonders existing at that point in time. This intent was reiterated more explicitly 91 years later by the Leopold Committee: “As a primary goal, we would recommend that the biotic associations within each park be maintained, or where necessary re-created, as nearly as possible in the condition that prevailed when the area was first visited by the white man.” As Wagner et al. (1995a) pointed out, a memorandum issued by Interior Secretary Udall to NPS Director Wirth on May 2, 1963, instructed the director to incorporate the philosophies of the Leopold report “into the administration of the National Park System.” The logic of this interpretation of the natural is that it fits into the overall raison d’être of the entire National Park System: to preserve vestiges of the nation’s prehistoric and historic settings out of which it has been formed. Thus, in passing the General Authorities Act of 1970, Congress declared, as quoted in the new policy manual, that the system should constitute “one national park system as cumulative expressions of a single national heritage . . . preserved and managed for the benefit and inspiration of all the people of the United States.” A conference of NPS officials in Vail, Colorado, in fall 1991 (Briggle et al. 1992: “the Vail Agenda”) concluded that the system should “preserve, protect, and convey the meaning of those natural, cultural, and historical resources that contribute significantly to the nation’s values, character, and experience.” The Vail Agenda called this overall purpose “nation building.” Boyce (1998) protests that the exact nature of the northern range at park establishment can never be known. This is, of course, precisely true. But with an immense photographic archive, the many accounts of individuals who saw the system in its early days, and what we know about it on the basis of nearly a century of research, that 1872 character could be approximated. Judging by the responses of the system over the past 132 years as the elk population has varied in numbers, it might well move back toward its early state if, as Boyce advocates, all of the processes present at park establishment were restored. The northern range today is substantially changed from its early condition, largely in response to human actions, and restoring it would likely require some human intervention as advocated in the Leopold report and the NPS policy manual. A more fundamental problem with a goal of maintaining a system in its early state, as pointed out by numerous observers, is that ecosystems change over time.


As commented in chapter 15, global warming is likely to alter the northern range profoundly. So any near-term (e.g., decades) effort to maintain an early state could only be a temporary accomplishment. A second ambiguity in the word natural is whether it implies the presence of humans and their effects. Views diverge on this question among different observers. Wagner et al. (1995a) reviewed these at length, but some more recent developments make it worth brief comment here. At one end of the continuum, natural is considered a descriptor for ecological systems totally lacking humans and human influences. As an example, Anderson (1991) stated: “The most common usage of ‘natural’ in the ecological literature is understood to mean a process, situation, or system free of human influence . . . a usage applied in the context of aboriginal as well as technological humans.” And Yellowstone’s research chief Varley (1988) stated, “Management’s primary purpose is to maintain the area’s pristine condition to the fullest extent possible. . . . This includes the perpetuation of natural processes in the absence of human interference.” But perhaps responding to the growing evidence that subsistence cultures significantly influenced North American ecosystems (see Kay 1998; Kay and Simmons 2002), including the Yellowstone area (chapter 3), NPS officials more recently acknowledge pre-Columbian aboriginal effects in their use of the term natural in NPS policy statements: the presumption that, at some time in the historic past, North American ecosystems existed in some natural state of grace which was immediately sullied when humans of European origin set foot on the continent. Resultingly, the environmental influences of European derived humans on America, and thus on the national parks, are unnatural while the influences of 10,000 years of habitation by Native Americans are natural. (Huff 1997b) And NPS has established informally, but not in its management policies, a more-or-less accepted policy inference that “technological humans”— generally those who developed the country after the discovery of the New World by Europeans—are to be considered not part of the nature and natural processes that NPS is to perpetuate. NPS provides no formal policy assessment of the role of pre-Columbian Native Americans in the evolution of the nature and natural processes of today’s parks, leaving an implication that, for the purposes of management today, the pre-Columbian Native American role may have been more within, than without, what was natural. (Dennis 1999) All of this may seem like needless quibbling. But if Yellowstone is to respond to the charge of the new NPS policy manual to establish a management goal, in this case for the northern range, and that goal is to preserve the natural, the goal is vacuous unless it is defined with this specificity and provides no clear guidance for management.

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Ecological/Natural-Process Management The concept of ecological- or natural-process management is both a goal and a management protocol. I will touch on the goal aspect here and discuss the management implications later. As commented, concern for park ecosystems, both by NPS officials and outside ecologists, shifted to preservation of natural processes in the 1980s. Accomplishing this purpose came to be known as “ecological or natural process management,” Boyce (1991, 1996, 1998) being one of the more frequent advocates for the procedure. NPS biologists Huff and Varley (1999), writing in the 1990s, commented that the procedure “as implemented in Yellowstone National Park and many other units of the national park system. . . . [is] the umbrella policy for most NPS natural resources management issues.” Essence of the procedure, as advocated by Boyce (1991, 1998), is preservation of biotic and abiotic processes and reliance on them to maintain ecosystem function. Huff and Varley (1999) specify that these processes must not be induced or significantly affected by modern humans. This leaves open the question of aboriginal effects, as already pointed out. The evidence indicates that aboriginal hunting was an influential process in the Yellowstone region in preColumbian times. But Boyce (1998) asserts that any hypothesis of significant aboriginal effects “is essentially untestable and highly unlikely.” Thus he advocates “maintenance of natural disturbance regimes and natural population fluctuations, unhampered by human control even when those fluctuations might lead to local extinctions.” However, “human intervention may occasionally be necessary to restore or protect the functioning of ecological processes, particularly where the system has been significantly disturbed” (Boyce 1991:190). In the course of its process management, Yellowstone has restored wolves to the northern range and chemically controls exotic plants. Some of the discussion on process management treats processes in the abstract as though they exist independent of system components and by implication the components do not need to be considered in management efforts. But as Wagner et al. (1995a) pointed out, processes are the functions of both biotic and abiotic components of an ecosystem. There is no way that the processes can be preserved without preserving the components. Hence, the goal of process management is tantamount to preserving entire ecosystems.

Natural Laboratories Boyce (1996) has recommended that national parks should serve as laboratories for studying natural ecosystems largely uninfluenced by human actions. Sinclair (1998) has similarly argued at length that parks and protected areas can serve as ecological baselines to compare with and provide understanding of human effects on unprotected landscapes that cover most of our globe. This is a utilization goal for national parks, not a goal that specifies some desired future condition. Sinclair reviews the changing nature of ecosystems, often over long time scales, and therefore argues that no particular state is desirable or should be preserved in national parks. In essence he advocates preservation of “natural” systems that are


free to undergo these long-term changes. Consequently it is not advisable to engage in “whimsical” management designed to preserve any given state. However, according to Sinclair, it is desirable to maintain some “ecological baseline” (criteria not defined). To this end, management should eliminate exotics, employ prescribed burning where natural fire regimes have been altered, and maintain habitats for rare species. “Whether prehistoric humans were a component of a system at some other time or not is beside the point with respect to modern impacts.” In fact, this and the goal of preserving the natural need not be mutually exclusive. National parks functioning to preserve some natural state, uninfluenced by technological humans, can serve as ecological baselines to be compared with areas subject to major human alteration. Although the goal of preserving some natural state may suggest a qualitative difference because of its implicit need for management, this may be only a matter of degree vis-à-vis preserving ecological baselines. However, in North America, and perhaps the entire Western Hemisphere, the presence of prehistoric humans is clearly not beside the point in preserving the natural.

Preserving Biodiversity and Heterogeneity Although Huff (1997a) states that “maximum biodiversity has never been a management goal for national park units,” the 1988 policy manual stated that managers and resource specialists “will try to maintain all of the components and processes of naturally evolving park ecosystems.” And Brussard (1991b) argued that public lands, including YNP, should play a major role in preserving biodiversity. More recent NPS policy explicitly prescribes this role as the quotes from the new policy manual indicate. And the deliberations of the NPS general conference in September 2000, Discovery 2000, concluded that “national parks will play an increasing role in biodiversity preservation” (Selleck 2001). This goal also is not necessarily antithetical to preservation of the natural and to the use of parks for outdoor laboratories. It may also imply advertent management as needed to preserve species and habitats, an intervention that does not differ in principle from Sinclair’s maintenance of habitat to preserve rare species and more generally preserve an ecological baseline. Preservation of biodiversity and savannah heterogeneity has been adopted as the goal for Kruger National Park (du Toit et al. 2003). Whyte et al. (2003) conclude that uncontrolled elephant populations in the park “can simplify habitats in ways that diminish biodiversity.” Their proposed solution for the preservation of biodiversity and ecosystem heterogeneity is a culling effort, temporally rotated around four regions of the park, to maintain elephant numbers “within some predetermined bounds.” Preserving Ecological Integrity/Health The Canadian Parliament amended the nation’s National Parks Act in 1988 to require the Canadian Parks Service (now Parks Canada) to include in its national

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park management plans the “maintenance of ecological integrity” (Woodley et al. 1993:157–58). This charge stimulated a great deal of intellectual exploration into the meaning of ecological integrity and culminated in the Woodley (1993: 157–58) definition for Canadian national parks: Ecological integrity is defined as a state of ecosystem development that is optimized for its geographic location, including energy input, available water, nutrients and colonization history. For national parks, this optimal state has been referred to by such terms as natural, naturally, evolving, pristine, and untouched. It implies that ecosystem structures and functions are unimpaired by human-caused stresses and that native species are present at viable population levels. Woodley goes on to discuss the complex monitoring procedures needed to observe ecological integrity across the Canadian national park system. Woodley’s preclusion of human impairment again raises the question of native versus technological human effects, an issue that apparently had not arisen at the time he published in 1993. But recent research in Canadian parks is producing evidence of the same aboriginal effects as have been shown for the western United States (White et al. 1998; Kay et al. 2000; Kay and White 2001), and this aspect is being reconsidered by Parks Canada officials.

Management Options for the Northern Range Adaptive Management A number of management options have been proposed for the northern range. These entail some of the same ambiguity problems as many of the proposed goals: 1. Some are proposed without specifying goals and therefore leave it unclear precisely what management is to be carried out, and what is to be achieved by the management. 2. In some cases, the proposed goal and management are one and the same. One example is adaptive management said by Bishop et al. (1997) to be the management regimen for the northern range, and recommended by Klein et al. (2002) to be the appropriate procedure. As described by Holling (1978) and Walters (1986), adaptive management is experimental management, employed in situations where a system is not so well understood that the managers can be certain (or confident) that the procedures will accomplish the desired outcome. The responses are carefully monitored over time to ascertain whether the system’s trajectory is moving toward the desired end point. If it is not, the protocol can be altered in the hope of adjusting the trajectory toward the desired future condition, the change often revised on the basis of what was learned during the original trajectory. Rogers (2003) calls this “learning by doing.” Borman et al. (1999) describe three forms of adaptive management. “Reactive” is modifying the management trajectory over time as political pressures of the moment dictate. “Passive” is deciding on a single management protocol and


carrying it through. “Active” adaptive management, also called “parallel,” employs more than one simultaneous management protocol on the rationale that it increases the probability of attaining the goal. All of these variants imply or require some goal, as Klein et al. (2002) state. It is not clear why or how the management should be carried out if it is not employed to achieve some end point. But neither Bishop et al. (1997) nor Klein et al. (2002) specify the goal to be achieved in recommending adaptive management for the northern range. The variants, as already stated, also require a monitoring program designed to trace a system’s response to the management efforts and determine whether it is on course toward the goal. Although Yellowstone censuses northern range ungulates annually and has followed wolf populations since their reintroduction, the other major components of the system have not been systematically followed over time. Thus, employing adaptive management on the northern range would require a more systematic monitoring effort. Kruger National Park provides an excellent model of the use of adaptive management in a national park. Using a collaborative policy-setting mechanism, Kruger has adopted the preservation of savannah heterogeneity and biodiversity as its goal. It has designed an extensive monitoring program to evaluate the effects of management efforts, and has set “thresholds of potential concern (TPCs)” as reference points beyond which the system should not be allowed to proceed (Mills et al. 2003) and management would intervene.

Natural-Process Management The management aspect of this procedure contains some of the same conceptual and definitional hiatuses as the goals discussed. Freemuth (1999) commented, “it becomes problematic what a concept like natural process management actually means.” By the above definitions, this regimen consists of restoring and maintaining natural processes. But given Boyce’s differing definitions, it is unclear whether management should intervene to prevent loss of processes (in fact components), or curtail processes risen to excessive levels. Huff and Varley (1999) ask, “Should hands-off natural process management be continued, with or without some precisely defined thresholds for intervention?” They answer their own question by stating, “The National Park Service should . . . reconsider the possibility of establishing a set of ecological thresholds indicating the need for intervention or additional research.” The greater conceptual blank in the natural-process issue is the question of what processes are to be considered and what is to be done with them? Each species on the northern range engages in scores, perhaps hundreds or even thousands, of processes—physiological, demographic, community, and ecosystem. If there are a thousand plant and animal species, and additional abiotic components on the northern range, there are orders of magnitude more processes than species and abiotic components. What of these are to be identified and considered? Measured? Monitored over time? Managed?

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I do not raise these questions facetiously or to denigrate. I raise them because I believe the policy needs to be clarified and defined ecologically and viewed realistically before it can be carried out in a management program. Conceivably it may be pragmatically unworkable when the actual meaning is explicated ecologically. Or it may prove to be nothing more than preserving the natural components with decisions needed on permissible levels of management intervention, if any.

Natural Regulation/No Management Like other management regimes advocated for the northern range, there is inconsistency in what is meant by the natural-regulation policy. As discussed in chapter 1, the policy was announced in three park releases or memoranda in late 1967 following the March Senate hearings and decision by NPS Director Hartzog to stop ungulate culling, at least for a period of time. The park had been culling elk, bison, and pronghorn, in some cases (elk) for more than 30 years. Natural regulation has been widely considered to be a no-management policy. The park has not resumed elk culling in the ensuing 37 years. Yellowstone biologist Meagher (1974) wrote that “bison management in Yellowstone National Park may be termed ‘no management.’” The Gordon Commission (Bishop et al. 1989) referred to NPS management as “laissez faire,” as did Keiter (1989). And Huff and Varley (1999) commented on “hands-off natural process management,” the latter considered by them to be synonymous with the natural-regulation policy. No management has its supporters in the environmental community (see Nash 1973; Rolston 1990; Pritchard 1999) who consider naturalness as the appropriate goal for national parks. Management intervention is looked on as an intrusion on that goal. However, although natural regulation has been widely construed to be a no-management policy, the 1967 documents included provisos that the park would cull elk in the event of undesired population increase and unacceptable impacts on the northern range system. Thus, “Resort to limited control of elk numbers only in park areas where natural regulatory mechanisms fail to hold numbers below levels where they cause lasting damage to vegetation or . . . other park wildlife” (Anonymous 1967c). And the park restored wolves in 1995, controls exotic plants, and cooperates in controlling bison that leave the park in winter in response to the politically sensitive issue of brucellosis infection of cattle (Cheville et al. 1998). Thus there are versions of natural regulation as with the other policies. Moreover, NPS policy documents have consistently contained provisos authorizing management corrections for systems that diverge from desirable states as a result of modern human activities, especially ungulate populations that have increased to levels where they significantly impact park ecosystems. Because natural regulation has been the general policy for the entire NPS system since the 1960s, these authorizations implicitly apply to the northern range. Thus, “Natural processes will be relied on to control populations of native species to


the greatest extent possible. Unnatural concentrations of native species caused by human activities may be controlled” (Anonymous 1988). And “Biological or physical processes altered in the past by human activities may need to be actively managed to restore them to a natural condition or to maintain the closest approximation of the natural condition in situations in which a truly natural system is no longer attainable” (Anonymous 2000).

Discussion The intent of the goals discussed here are in fact more similar than different. They all advocate preserving national park ecosystems largely unaltered by technological humans. There is yet no general agreement on how to consider effects of preindustrial humans. But there appears to be movement toward acknowledging the growing evidence that pre-Columbian cultures did exert significant effects and considering these natural in North American systems. Some decision needs to be made on this matter to arrive at fully explicit and generally recognized goals for the northern range. As sketched in the schematic, goals are set in public policy process to satisfy societal values, and the values to be attained by the goals for American parks are threefold. The first is preserving remnants of American heritage and the cultural enrichment this self-knowledge brings to its society. This is surely the major value intended in the Yellowstone Act and more definitively in the Organic Act. The second value is the aesthetic and intellectual enrichment to be gained by American citizens in contemplating and understanding the structure and function of natural ecosystems. The third is the scientific, conservation, and educational values of maintaining ecosystems largely unaltered by Euro-Americans to understand their structure and function so that the effects of contemporary human use can be predicted and sustainability of manipulated systems assured. The goal statements all contain provisos for some degree of management. What management is appropriate for the northern range is entirely contingent on the goal. As Wagner et al. (1995a) emphasized, management is a means, not an end. There is no point in setting a goal if the management means for attaining it are not acceptable and agreed on. There is irony in the northern range policy trajectory since 1872. Early EuroAmerican actions altered the natural state: Park protection, removal of Native Americans, predator control, and artificial winter feeding enabled the northern herd to burgeon and impact the northern range ecosystem. Culling between the 1930s and 1969—condemned at the time by hunters who wanted to do the shooting, and vilified decades later on ideological grounds by environmental and animal welfare groups—was in fact consistent with what would become the provisos for corrective action in the 1967 natural-regulation documents and later policy manuals. The subsequent laissez-faire implementation of natural regulation by the park has been contrary to those provisos and the goal of restoring and maintaining the natural state of the northern range.

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In my view, there is a need for adopting an explicit goal for the northern range in a collaborative setting involving representatives of all concerned interests. It should presumably be consistent with the generic goals articulated in the 2000 policy manual. It should also be clear and agreed on by the participants that the park would have the prerogative of employing whatever management efforts are needed to attain the goal if natural processes do not do so. It may be that wolf restoration will move the northern range to the agreed-on goal. But if it does not, other options should be open. Some authors protest that ecosystems are so complex that the results of management interventions cannot be predicted. They point to management failures in the past—for example, predator control, winter feeding—that have been counterproductive. But these judgments focus on missteps taken when our science was in its infancy, if in fact born. And they dwell on what we do not know while belittling the extensive understanding of ecosystems that ecology has produced in recent decades. Again, this is not an unqualified recommendation for management. There is no suggestion here that our knowledge of ecosystems is anywhere near complete or that management steps cannot go astray. But it is again argued from the view that it is pointless to set goals for national parks if we are not prepared to do what is necessary to achieve those goals. We surely need an enlightened, wellarticulated, and proactive approach to protecting and preserving these priceless assets so that they continue to provide their values to current and future generations. That approach needs to function in the bright light of objective scientific understanding.

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Index

Page numbers in italics are tables or figures, those in bold are photographs or maps. adaptive management, 330–331 “Administrative Policy for the Management of Ungulates,” 5 alder, 246–247 algae, 250, 277 Allen, H. F. A., 281, 293–295, 298 allogenic silica, 237–239, 253–254, 256 117 allysum, 116–117 Alpine fir, 131 amphipods, 250–251 Anderson, G. S., 150, 208 annual net above-ground primary production (ANPP), 261–263 Argentine Pampas, 269 arthropods, 89–90, 123 aspens. See also conifers; deciduous trees arthropod diversity, 89–90 avian diversity, 88–89, 211–212 browsing of and clone vigor, 75–76, 80 browsing rate effects on, 62, 125, 133, 135, 271 chronology of tree establishment, 68– 71 climate change effects, 77–78 clone in Eagle Creek, 60 death of, 62, 66 66, 80–81

decline of in YNP, 62, 64–68, 136, 208, 283–284, 291 elk and, 5, 63 63, 71–75, 177 and exotic grasses, 116 fire suppression effects, 78–79, 83–84 Foster Lake, 252 growth characteristics, 59–60, 81–82 heart rot, 69–70 highlines (1932), 65 in the Little Blacktail area (1893), 67 67, 71 palatability of, 74, 85, 107 pollen data on, 246–247 pre-YNP establishment, 84 ramets, 59, 62, 64 64, 79 regeneration of, 74 research on, 173 and retrogressive succession, 7 61 shrub, 60–61 61, 73 avifauna, 88–89, 123, 212, 214, 288– 289, 291, 297. See also individual species Babbitt, Bruce, 309, 320 Barlow, John W., 174 Barmore, William J., Jr., 6, 7, 38, 93–94, 145, 147, 177–178, 264, 286, 308, 312–313. See also aspens

359

360

Index

bears attacks by, 138 Craighead research on, 308–309, 313 food sources, 138–139, 290 riparian systems and, 214, 296 and whitebark pine, 301 in the Yellowstone northern range, 294–295 beaver aspen death and, 80, 84 and avifauna diversity, 212 Big Slide Lake, 249, 255 decline of in YNP, 208–209, 288–289, 291, 297 historical evidence of, 32, 207 willow habitat and decline of, 183 183, 194–197, 201, 246 Bebb willow, 195 beggar’s lice, 247 Belknap, Jane, 255 below-ground production (NBP), 267– 268 berries, 138–140, 290. See also shrubs big rabbitbrush, 109 big sagebrush. See sagebrush Big Slide Lake, 244–246, 249–250, 254– 255 Big Trumpeter Lake, 234 234, 236–237, 245–246, 251 bighorn sheep. See also ungulates archaeological data, 30 census data for northern range, 147 historical evidence of, 34 interspecific competition, 145–149, 168 population of in YNP, 210, 288 wintering distribution, 160 biodiversity, 140–142, 296–298, 329, 331. See also ecosystem bioenergetics, 259–260 biogenic silica, 239, 241–242, 253–254, 258 Bishop, N., 147, 154–155, 163, 168, 179– 180, 208–209 bison. See also ungulates archaeological data, 30 and brucellosis, 151 control of, 332 historical evidence of, 34 interspecific competition, 146

northern range, impact on, 160–161, 294 nutrition research, 144, 266 photographic record and conifers, 128 population behavior, 153–160 prehistoric population of, 161–162, 288 snowmobiles and, 157–158, 310–311 winter range of, 169 YNP census of, 154–155 bitterbrush, 138–139 BLA (Boundary Line Area), 4 , 91, 102– 103, 123, 164, 234 234. See also Yellowstone National Park (YNP) Black Butte Creek, 111 black cottonwood, 202 black twinberry, 138, 213 Blacktail Creek exclosure, 96–97, 114, 186, 194, 287 Blacktail plateau, 216 bluegrass, 86–87 bobcat, 171 Bottler, Frederic, 138 Boundary Line Area. See BLA (Boundary Line Area) Bouteloua gracilis, 37 Boyce, M.S., 302, 326, 328 Bozeman, MT, 36, 309 Brewer’s sparrow, 122 Bridger-Teton National Forest, 85–86 browsing. See also grazing; highlines aspen, effects on, 62, 75–76, 80 134 dendrochronological research, 133–134 and elk effects, 88, 92 lines, 62, 65 obligates, 151 pronghorn antelope, 165 sagebrush, effects on, 101–102, 107– 108 willow, impact on, 174, 180–181 brucellosis, 151, 332 Buck Lake, 234 234, 239, 244–246, 256–257 Buffalo Ford, 234 234, 246, 256 buffaloberry, 138–139 bulrush, 252 Bureau of Land Management (BLM), 319–321 butterflies, 212–213 Cahalane, V. H., 176–177 Calfee Creek, 225

Index

Camp Creek, 139 Canada thistle, 247 Canadian Park Service, 329–330 carnivores. See predation carrying capacity, 22, 29, 44–45 cattle, 92–93, 151, 246, 249, 256 Caughley, G., 45 census data. See also elk during 1923-1968, 24–26 discontinuity of, 54 Estimated Percentage of Fall Populations, 25 methodology of counts, 15–16 natural-regulation policy 1969-2003, 17–24 reliable estimates of 1914, 48 variations of, 16–17 Winter Censuses of Fall Populations, 19 Centennial Mountains, 197, 212 Chapman, W. S., 250, 252 Chara, 252 Chase, A, 8–9, 310 117 cheatgrass, 115–117 Chlamydia, 149–150 140 chokecherry, 139–140 140, 213 climate change, 77–78, 183 183, 188–194, 200–201, 247, 301, 327. See also winter weather Cole, Glen, 6–7, 312 Columbia River, 32 common juniper, 252 common snowberry, 138 conifers. See also highlines; individual tree types avian diversity, 89 and bison, 128 decline of in YNP, 135–136 and fires, 77, 126 global warming models, 301 invasion of in aspen territory, 79–80 palatability of, 125 as percentage of YNP, 124, 259–260, 285, 291 types of, 124–125 as winter elk feed, 144 Cooke City, MT, 36, 250 Cook-Folsom-Peterson expedition, 138 Cooperative Fisheries and Wildlife Research Unit, 74, 308–309

cottonwood trees. See also deciduous trees 203 architecture of (Keigley), 202–203 avian diversity, 88, 211–212, 214 browsing history data, 204–206 decline of in YNP, 202, 286–287 highlines, 131 increase in sapling release in 2002, 11 palatability of, 107 pollen data on, 246–247 wolves, effect on, 287, 300 cougars, 294–295 Coughenour, M.B., 19–23, 85, 112–115, 269 coyotes, 166, 170–171, 252, 289–290, 294–295, 300. See also predation Crabtree, R. L., 171 Craighead, John and Frank, 308–309, 313 Crow Indians, 35 culling. See also hunting bison, 151, 156, 159 condemnation of, 333 direct reduction, 5 elephants, 329 elk, 322, 332 natural-regulation policy, 17, 153, 322 pronghorn antelope, 166, 170 in RMNP, 322 Debinski, D., 212–213, 289, 297 debris flow. See erosion deciduous trees, 138–140, 271, 286, 297. See also individual types; shrubs 103 Deckard Flat, 102–103 deer, 30 defoliation, 269–270 DeLacy journal, 33 DelGiudice, G. D., 143–144 dendrochronology, 202–206 denitrification, 276–277, 279 density-dependence forage supply and population, 119– 120, 144, 266 in northern elk herd, 21–22, 46–47, 50, 143–144, 283, 298 and regulation, 26 winter mortality and bison, 155

361

362

Index

Desert News, 197 diatoms, 243–244, 257 Dome Mountain, 92, 169–170 Douglas fir. See also conifers browsing history data, 135 elk browsing effect on, 177 Foster Lake, 252 highlines, 131 on Middle Rainbow Lake, 248 palatability of, 107 percentage of in YNP, 124 photographic record and conifers, 129 drought. See also climate change and aspen decline, 77 elk population decline, 23, 50, 300 Floating Island Lake, 251 and herd size variations, 11 lake sedimentation and, 236 and soil erosion, 216 vegetation, impact on, 111 and willow decline, 189–190 duff, 279 dwarf shrubs. See sagebrush-steppe system Eagle Creek, 63 63, 74–75, 80, 87, 116 East Elk Refuge, 139–140 Ecological Applications, 269 ecosystem. See also Greater Yellowstone Ecosystem (GYE) bioenergetics and, 259–260 defined, 292–293 ecological/natural process management, 328–330 policy setting and, 317–318, 325, 334 processes of, 290 Serengeti, 271–272 steady state, defined, 298 elephants, 329 elk. See also aspens; census data; naturalregulation hypothesis ANPP rates, 270–271 archaeological data on, 30–31, 38 avifauna, effects on, 123 beaver decline and, 197, 209–210 biogeochemical functions, 272–273 butterflies and, 212 carrying capacity in pre-park era, 29 census data for northern range, 147

conifer browsing, 125, 131 cottonwood decline, 202, 206 diet of, 162, 264–266 drought impact, 11 erosion research, 216–220, 223–224, 232 herbaceous vegetation and, 110–111, 115, 121, 285 herd estimates, 40–44, 151, 282 historical data on, 31–35 interspecific competition, 143–149, 161 lake sedimentation and, 254, 257–258 management phases in YNP, 3–6 migration paths, 35–36 population fluctuations, 26–27, 48–52, 282–283 4 , 10– range and feeding patterns, 3–4 11, 20–22, 92, 169–170, 278–279, 294 riparian systems decline, 214 RMNP environmental impact statement, 321 sagebrush, browsing effects on, 95 95, 100–108, 119 stream bank sloughing, 225 Superintendents reports on, 41 understory vegetation, impact on, 86– 87 ungulates per 100 elk, 152 white-tailed deer decline, 211 willow decline, 183 183–187, 190–191, 200 YNP increase in, 167–168, 281, 333 Elk Creek, 174 Endangered Species Act, 213, 321 Engelmann spruce, 125, 126 126, 131, 133– 134 134, 137 137, 252. See also conifers Engstrom, D. R., 232–243, 246–249, 252–254, 257 environmental impact statement (EIS), 309, 321 equilibrium state, defined, 22 erosion. See also soils elk contribution to, 257–258 grazing effects, 276 gully erosion, 221–223 historical observations, 215 Kittams photos, 232

Index

in Lamar River valley, 110–111 landslides in YNP, 234 runoff rates, 220 sediment yields, 221 surface erosion in YNP, 277–278, 292 topsoil of northern range, 217 in YNP, 221–223, 229, 277–278 eruption (population), 26–27 eutrophy. See lakes evapotranspiration, 198 Everts, Truman, 139 exclosures. See also individual areas ANPP research, 263 aspen growth in, 67, 72 berry production within, 139–140 drought and aspens, 77 erosion research, 217–221, 224 established, 76, 96 exotics vegetation measurements, 116– 117 fires and aspen growth, 79 herbaceous vegetation in, 111–115 as measurement tools, 55 natural-regulation hypothesis, 108 overcompensation research, 268–269 Range Plots, 72–73, 87 sagebrush measurements, 97–100, 106–107 shrub cover measurements, 98 understory vegetation, 85–86 ungulate grazing research, 292 186 willow decline evidence, 184–186 exotics (non-native) vegetation, 115– 117, 285, 328–329, 332. See also vegetation false mountain willow, 195 federal land-management agencies, 319 Finley, R. B., Jr., 313 Firehole River valley, 153 fires. See also lightening controlled burns, 329 effects on conifers, 126 1988 event, 8, 10, 61, 96, 136, 199, 291 riparian systems, effects on, 199 sagebrush, effects on, 100 suppression and aspen decline, 78–79, 83, 84

suppression and willow decline, 183 183, 197–199, 201 fishing, 30 Floating Island Lake, 234 234, 236–237, 246, 250–251, 255, 256 floodplain, 227–228, 230 forb cover, 86–88, 112, 115. See also vegetation Forest and Rangeland Renewable Resources Planning Act of 1974, 320 Fort Collins, CO, 309 Foster Lake data analysis, 257 eutrophication of, 244–245 location in YNP, 234 pollen decline, 246, 256 sedimentation in, 236–237 setting and history, 251–252 water chemistry of, 253 fox, 171 Frank, D. A., 260–264, 266–267, 271– 272 Freedom of Information Act, 313 frost heave, 217 217. See also erosion Gallatin National Forest, 85–86, 223–224 Gallatin River, 178, 190 Gardiner, MT, 43, 91, 202, 216, 249 Gardiner exclosure, 97–98, 100 100, 109, 117 117, 121, 284 Gardner River, 173–174, 176, 210, 221– 222 General Authorities Act of 1970, 326 Geyer willow, 195, 196 Glacier National Park, 212, 301 Gogan, Peter, 153, 158 Gordon Commission, 332 graminoids, 87, 266 Grand Loop Road, 250 Grand Teton National Park, 30, 139, 289, 313 grasses. See also individual species; vegetation ANPP research, 250, 262 avian diversity, 88, 212 and elk browsing, 88, 144, 218, 283– 284 and exclosures, 112, 115

363

364

Index

grasses (continued) overcompensation research, 267 percentage of in northern range, 86–87 and sagebrush, 106, 118–119 and sheetflood sediment, 222–223 grasslands, 259–260, 272 gray partridge, 123 grazed net above-ground production (NAP), 267–268 grazing. See also browsing; individual animals cattle and big sagebrush, 92–93 effects of, 5, 92, 275–276 erosion and, 218 overcompensation, 266–271 overgrazing, 8–9, 92, 309 Great Plains, 37 Greater Yellowstone Ecosystem (GYE). See also Yellowstone National Park (YNP) archaeological data, 30, 299 berry production within, 139 global warming models, 301 habitat reduction of, 288–289 moose remains, 188 northern range ecosystem hierarchy, 293–295 northern range steady state debate, 298–299 research in, 280–282 and the Serengeti, 271–272 understory vegetation, 85–86 winter range expansion in, 299 green rabbitbrush, 109 Green River, 30 Grimm, R. L., 72 grizzly bears. See bears GYE. See Greater Yellowstone Ecosystem (GYE) habitat complementation, 296 Haines, A. L., 35 Harris, M., 35–36, 41 Hartzog, G.B., Jr., 6, 308 Hayden expedition, 31 Hayden Valley, 128, 153 Haynes, F. Jay, 67 Heap, David P., 174

herbaceous vegetation, 110–111, 113– 114, 119, 120–122, 284–285, 290– 291. See also vegetation 134 highlines, 65 65, 125–134 134, 285. See also conifers Hoback River, 138–139 horses, 128 houndstongue, 248, 250 Houston, D.B., 7, 15, 18, 26–28, 32–33, 40–44, 96, 145, 148, 162, 215–216, 313 Huff, Dan, 318, 323, 325, 328, 329, 331 hunter-gathers, 30 hunting. See also culling and bison, 156 and census data, 16–17, 24–25, 43 and elk, 36, 41–42, 92, 282–283 and the natural-regulation policy, 6 offtake and herd size, 21–22 state laws and deer, 45 YNP management phase two, 5 hydrogen sulfide, 126 Iddings, J. P., 127, 177 insects, 183 183, 188, 200, 297 Interagency Grizzly Bear Study Team, 313. See also bears interest-group pluralism, 320–321, 324 Intermountain West, 30–31, 37 interspecific competition, 141–145, 170– 171 intraspecific competition, 143–144, 156, 160 Jackson, W. H., 129–131 Jackson Hole, WY, 30, 138–139 Jefferson River, 32 Journal of Paleolimnology, 233 Junction Butte exclosure, 76, 87, 96–97, 186 178, 185 185–186 186, 191–194 61 Kay, Charles E., 60 60–61 61, 64 64, 71, 75, 85– 87, 130 130, 168, 173–174, 177, 185, 187 187–188 Keating, K. A., 147–149 Keigley, Richard B., 133–136, 177–179, 202–206, 227 227–228, 246, 286–287, 313–314

Index

Kentucky bluegrass, 116 kettle lakes, 236, 239–240, 245, 253. See also lakes Kirschner Marsh, 243 Kittams, Walter H., 173, 175 175–177, 181, 190, 202, 231, 233, 286 Kruger National Park, 329, 331 lakes catchments and exotics, 247 Chara, 252 diatom stratigraphy of, 243–245 234 Engstrom research, 232–234 234, 254–256 eutrophication of, 239, 244–246, 250, 254, 258 focusing, 255 sedimentation rate changes, 236–243 settings and water chemistry synthesis, 252–254 water-milfoil, 252 Lamar East/West exclosures berry production within, 139 established, 76 sagebrush measurements, 96, 121 understory vegetation, 87 willow growth in, 185 185, 191–194 and the Yellowstone checkerspot butterfly, 213 Lamar Ranger Station, 153 226 Lamar River, 225–226 226, 292 Lamar River valley aspen trees in, 73–74 bison population in, 153, 160 cottonwood trees in, 204–206, 287 elk herds in, 36, 41–42, 48, 216 erosion and, 223 exotics invasion, 116 herbaceous vegetation in, 110–111 sage grouse, decline of, 123 sagebrush in, 105 stream flow in, 198 175 willow growth in, 174–175 175, 179 179, 181, 187 187, 191, 287 Landslide Creek, 248 Lane, J. R., 217–220 least chipmunks, 123 Lemke, Thomas, 11, 50, 164 Leopold, A. Starker, 6, 27, 44–45, 308

Leopold Committee, 324, 326 Lewis and Clark expedition, 31–32 lichens, 228–229, 248–252, 277. See also vegetation lightening, 79, 100–101, 279. See also fires limber pine, 125, 130 130–131. See also conifers limitation, defined, 26 Limnological Research Center, University of Minnesota, 232–233 Little Blacktail, 67 67, 71 Little Ice Age, 77, 298 Livingston, MT, 35 lodgepole pine, 88, 124, 131, 212. See also conifers Lost Creek, 190 Lower Rainbow Lake, 249 lungworm infections, 149 Madison Junctions, 127 magpies, 214 Mammoth Cave National Park, 315 Mammoth exclosure, 76, 87, 96–97, 118, 139, 185 185, 191–194 Mammoth Hot Springs, 36, 73–74, 77, 217 131, 153, 202, 216–217 Mammoth Terrace, 129–131 management. See also culling; elk; natural-regulation policy adaptive, 330–331 BLM policy procedures, 320 ecological/natural process, 328–330 “Management Objectives for Northern Yellowstone Elk,” 5 Management Policies 2001, 323 military of YNP, 4 natural process, 331–332 phases of in YNP, 3–7, 10–11, 51, 314–316 policies in RMNP, 321–322 Resource Advisory Councils (RACs), 320–321 market hunting. See hunting Mary Mountain, 153 Mattson, David, 313 McGee, Gale W., 6 microphytic crusts, 277, 292

365

366

Index

Middle Rainbow Lake, 234 234, 244–246, 248–249, 254–256 Midway Geyser Basin, 128 military, 4, 15 Miners Creek, 190 minimum number of individuals/animals (MNI), 30 Minnesota National Guard, 67 montane voles, 123 moose archaeological data, 31, 151 decline of in YNP, 288 diet of, 162, 211 willow decline and, 183 183, 188, 200 Mount Everts, 103, 145, 147–149, 166, 221–223 mountain alder, 138 mountain big sagebrush. See sagebrush mountain gooseberry, 138 mountain lions, 35 mountain willow, 195 mule deer. See also ungulates archaeological data, 30 coyote predation and, 170 diet of, 162 grazing effects, 119 historical evidence of, 34 interspecific competition with elk, 168 northern range census, 163 nutrition of, 164–165 population trends, 162–165, 210, 288 sagebrush, effects on, 93, 101–102, 123 Murie, A., 145, 164 narrowleaf cottonwood, 202 National Academy of Sciences, 324 National Biological Survey (NBS), 309, 323 National Bison Range, 170–171 National Elk Refuge, 30, 139 National Environmental Policy Act, 321 National Forest Management Act of 1976, 320–321 National Park Service (NPS). See also individual parks biodiversity, management of, 329 census data from 1969-2003, 17 elk census (1920s) in YNP, 41

formation of, 15 goals for YNP, 323–326 Gordon Commission report, 332 history, 319 Management Policies 2001, 323 manuals and natural preservation policies, 325 NRC report on, 315 policy making pressures, 320–321 pre-Columbian aboriginal effects in YNP, 327 research and, 308 scientific freedom and, 322 National Parks Act (Canada), 329–330 National Parks Advisory Board, 6 National Parks Organic Act of 1916, 15, 324–325, 333 National Research Council (NRC), 105, 157–158, 201, 297, 302, 315, 326 Native Americans. See also individual tribes archaeological data, 30–31 diet of, 38, 138 and elk ecology, 5 evicted from Yellowstone, 4, 333 fire use and aspen growth, 79 hunting and ungulate population before YNP, 50, 72, 282, 294–295 “Natural Control of Elk,” 5 Natural Science Studies office, 312 natural-regulation hypothesis. See also elk; natural-regulation policy beaver decline and, 206–207 competition as component, 143 contradictions in, 81–85 development of, 6, 308 and exclosures, 108 as experiment, 51–52 Houston predictions, 7, 44–45, 303 public information portrayal, 314 scientific research problems, 310–312 skepticism about, 7–8, 45–46 testing of, 8–10, 26–28, 302–303 willow decline and, 201–202 natural-regulation policy. See also management; natural-regulation hypothesis; science and bison, 151 census data from 1969-2003, 17–24

Index

Chase critique of, 9, 313 ecological effects of, 317–318 impact of, 298 instituted, 5, 231, 322 natural, defined, 325–328 and politics, 6 synthesis, 10 NBS. See National Biological Survey (NBS) nematodes, 276 New Zealand, 45 nitrogen, 274–278, 292 Norris, Philetus, 4, 36, 41, 138, 150, 174 number of individual specimens (NISP), 30 nutrition, 23, 50, 143–144, 157, 266, 282 obligate browsers, 165. See also browsing Odell Creek, 197, 209 O’Gara, Bart, 74, 178, 286 Old Faithful, 128 oligotrophy, 244 O’Neill, R. V., 59, 281, 293–295, 298 organic sedimentation, 238–239, 240– 241 overcompensation, 266–271, 291 overgrazing. See grazing packrats, 30 Paradise Valley, 186 Parker, Samuel, 138 Parker transects, 113–115, 120, 269– 270, 284–285 Pebble Creek, 174 Pelican Valley, 153 phosphorus, 239–244, 253–254 photography, 54, 67 67, 71–73, 125–126, 284 Pierre’s Hole, 30 Pinedale, WY, 85–86 pinkeye, 149–150 planeleaf willow, 195 policy, 5–6, 9, 307, 318–323. See also natural-regulation policy; science pollen, 246–247, 255–256 population, 26. See also census data; individual animals pre-Columbian America, 31, 37–38, 50, 327, 333

predation. See also individual animals among carnivores, 171 and apparent competition, 170–171 on elk calves, 278–279 and the natural-regulation policy, 5 in the northern range, 294–295 in pre-Columbian era, 31, 38 ungulate population before YNP, 50, 282 and vegetation, 11 Progressive Era, 321 pronghorn antelope. See also ungulates coyote predation, 170–171, 289, 300 diet of, 93, 123, 162, 165 historical evidence of, 34 interspecific competition with elk, 168 population trends, 166–167, 288 research on, 74 sagebrush, effects on, 101–102 winter range of, 169 public relations, 309. See also Yellowstone National Park (YNP) pygmy rabbits, 123 rabbitbrush, 107, 249 rainbow trout, 250, 252 ramets. See aspens raptors, 214 ravens, 214 Raymond, W.R., 139 red cedar, 107, 131. See also conifers Red Desert, 30 red fox, 290 Red Rock Lakes National Wildlife Refuge, 212 red-osier dogwood, 138 regulation (population), 26 Rens, R.J., 99, 104, 115 Resource Advisory Councils (RACs), 320–321 retrogressive succession, 7–8 riparian systems. See also elk; individual components avifauna of, 211–212 and beaver, 196–197, 206–210 browsing effects, 271 cottonwood trees, 202 decline of in YNP, 291, 296 fire effects on, 199

367

368

Index

riparian systems (continued) importance of, 213–214 invertebrate fauna, 212–213 percentage of in YNP, 172, 180–181 and stream bank sloughing, 226 ungulates, 210–211 willow photographic record, 173 river birch, 138 rivers, 234 234. See also individual rivers RMNP. See Rocky Mountain National Park Rocky Mountain juniper, 125, 130 130, 131 131– 132, 177, 248 Rocky Mountain maple, 138 Rocky Mountain National Park (RMNP), 88–89, 194–196, 321–322 Romme, W. H., 68–72, 81–84 Roosevelt, Theodore, 319 Rush, W. M., 32, 139, 145, 164, 215– 216 Russell, Osborne, 150, 207 sage, 88 sage grouse, 123 sage sparrow, 122 sage thrasher, 122–123 sagebrush. See also sagebrush-steppe system avian diversity, 212 big sagebrush, 92–93 browsing, effects on, 107–108, 164, 271 competition with shrubs, 109 in Deckard Flats area, 103 elk, effects on, 92 exclosures, measurements in, 93 93, 97– 100 fires, effects on, 100–101 herbaceous vegetation and, 113, 291 Lamar River valley, 105 mean percent cover at higher elevations, 99 mountain big sagebrush, 92 northern range abundance, 96 palatability of, 107 photographic record of decline, 93 93– 95 95, 100 pronghorn antelope, effects on, 165 as ungulate food source, 91, 123 as winter cover, 198

sagebrush voles, 123 sagebrush-steppe system. See also vegetation ANPP research, 263 biology of, 92–93, 109 biomass of, 265 browsing, effects on, 106, 284 composition of, 91 dwarf shrubs, percentages of, 115 and elk browsing, 106 exotics, invasion of, 115–117 fauna diversity decline, 122–123 research on, 92 vegetation and ungulate use, 117–122 Schmitt, J. G., 221–223 science, 5, 307–312, 318–319, 322, 334. See also natural-regulation hypothesis; policy sedges, 144, 252, 262, 267. See also vegetation self-pruning, 126 Serengeti, 260–261, 269, 271–272 serviceberry, 138–140 sheep, 119 sheet erosion, 215–216, 222–223. See also erosion Sherwood, H. W., 98–99, 104 Shoshoni Indians, 32 shrubs, 86–87, 98, 136–140, 222–223. See also vegetation Sinclair, A. R. E., 328–329 Singer, F. J., 97, 99, 101–102, 104–105, 145–146, 163, 179–180, 195–196 Skinner, M. P., 5, 31–32, 62–63, 66, 72, 107, 202, 308 slide lakes, 236, 239, 245, 253. See also lakes 226 Slough Creek, 225–226 Slough Creek Lake, 234 234, 244–246, 253, 256 Slough Creek valley, 174, 198 Smith, C. L., 42, 46, 48–49, 71–72, 91, 107, 139, 174, 184 smooth brome grass, 116 smooth sumac, 138 snails, 90, 289 Snake River, 139 snowmobiles, 157–158, 310–311 Soda Butte Creek, 225, 227 227–228, 230 230, 292

Index

Soda Butte Creek valley beaver ponds in, 208 cottonwood trees in, 11, 202, 204– 206 erosion and, 223 willow growth in, 173–174, 256, 287 soils. See also erosion cryptogam, 37, 255 drought, 216 duff and elk grazing, 279 Engstrom research design, 233–235 frost heave, 217 herbivore improvement of, 276 microphytic crusts, 277, 292 Middle Rainbow Lake sedimentation rate, 249 pH levels, 277 sedimentation research, 232–233, 236–243, 247–248, 255 Tertiary Lost Creek Tuff, 249 spiders, 89–90 spiny horsebush, 109 sprouting shrubs. See sagebrush-steppe system stationarity, defined, 22 Stermitz, Raymond, 248–249 Stevens, D. R., 196 stream bank sloughing, 225–230, 292 stream turbidity, 223, 225–228 subshrubs. See sagebrush-steppe system summer mortality, 50 synthesis, 10, 51–55, 167–168, 280. See also science tannin, 194–195 Tertiary Lost Creek Tuff, 249 thistle, 252 Thompson, Ben, 131 timothy grass, 86–87, 116, 247, 252, 267. See also vegetation toadflax, 247, 250 187 Tom Miner basin, 186–187 Tom Miner Creek, 138 Tower Creek, 250 Tower Junction, 127, 137 137, 140 140, 208 Tower Junction exclosures, 180–181, 286 trapping, 207–208. See also hunting Typha, 249

Udall, Stewart, 324 Uhl Hill, 139 understory vegetation, 85–87, 89, 125, 137 137, 265, 283–284. See also vegetation ungulates. See also browsing; individual animals archaeological data, 30–31, 151, 299 biomass of, 168–170, 291 census data, 331 consumption rates of herbaceous vegetation, 263–266 erosion research, 224 grazing and exotic grasses, 116–117 herbivory in YNP, 226, 259–260, 294 historical evidence of, 34, 272 interspecific competition research, 145–146 and Native American diet, 38 phases of in YNP, 169 population behavior theory, 45 population of in YNP, 168, 288 in riparian systems, 210–211 sagebrush reduction and winter nutrition, 106 sagebrush-steppe as winter forage, 118–119 species in northern range, 162 winter food sources, 138 upland steppe. See sagebrush-steppe system Upper Geyser Basin, 128 Upper Rainbow Lake, 248 U.S. Bureau of Biological Survey, 5, 41– 43 U.S. Department of the Interior, 315– 316 U.S. Fish and Wildlife Service, 250, 308– 309, 319 U.S. Forest Service, 41–43, 320 U.S. Geological Survey, Biological Research Division, 136, 309, 314– 316, 323 U.S. Sheep Experiment Station, 197 USDA Dubois, Idaho Research Station, 119 USDA Forest Service, 5, 103 103, 308, 312, 319 Utah, 89

369

370

Index

Vail Agenda, 326 Varley, John, 323, 325, 328, 331 vegetation. See also exotics (non-native) vegetation; individual types; understory vegetation abundance of and herd size data, 27 ANPP procedure, 261–263 consumption rates of, 263–266 drought impact, 111 and elk, 11 enclosures built for research, 5–6 evolution of and herbivores, 37, 38 global warming models, 301 grazing effects by 1914, 5 lichens, 228–229 natural-regulation policy report, 6–7 nitrogen budget schematic, 275 overcompensation research, 266–271 and predation, 11 vesper sparrows, 123 volatilization, 276–277, 279 Wambolt, Carl.L., 98–99, 102–104, 109 Washburn, Langford, and Doane expedition, 31, 138–139 water birch, 246–247 water-milfoil, 252 weather. See drought; winter weather wet meadows. See sagebrush-steppe system whitebark pine, 125, 301. See also conifers white-tailed deer, 162, 210–211, 288, 297. See also ungulates whole-plant production (NPP), 267–268 wild rose, 249, 252 willows. See also deciduous trees avian diversity, 88, 211–212 browsing impact on, 133, 174, 286 canopy cover trends in exclosures, 185 186 185–186 climate change, impact on, 188–194 decline of in YNP, 136, 176–177, 181, 183 183, 200–202 in early YNP, 172–173 elk and, 5, 178–180, 184–187 fire, effects on, 199 Foster Lake, 252, 256 height increase in 2000, 11

highlines and Engelmann spruce, 133– 134 insect herbivory and decline, 188 Kittams photos, 231 Lamar River valley, 175 175–177, 179 179, 181 moose herbivory and decline, 188 pollen data on, 246–247, 256, 258 and retrogressive succession, 7 simulation model of decline, 181–184 stream bank sloughing and, 225, 230 tannin content and palatability, 194– 195, 311 wolves, effect on, 287, 300 Wind River, 30 winter weather, 17, 21–22, 27–28, 41, 49–50, 92, 144. See also climate change winterfat, 109 Wirth, Conrad L., 326 wolves. See also predation aspen tree establishment and, 72 census data, 331 as competition pressure on bighorn sheep, 150 control of initiated, 36 and coyote predation, 170–171 and elk population, 23, 26, 51, 294– 295 environmental impact statement, 309 and Foster Lake, 252 historical evidence of, 35 introduction of and effects, 11, 171, 281, 287, 296 and the natural-regulation hypothesis, 303 and the natural-regulation policy, 332, 334 as process management tool, 328 reintroduction research, 8 slaughter of in YNP, 41, 296 willow regrowth effect, 186 Woodley, S., 330 Wraith Falls, 174 Wright, G. M., 95 95, 131 Yancey’s Hole, 61 61, 73–74, 173, 190 Yankee Jim Canyon, 186 Yellowstone Act, 326, 333

Index

Yellowstone checkerspot butterfly, 213, 289, 297 Yellowstone Ecological Research Center, 309 Yellowstone National Park (YNP). See also BLA (Boundary Line Area); exclosures; Greater Yellowstone Ecosystem (GYE); management; natural-regulation policy archaeological data, 29 avifauna of, 89 beaver decline in, 195 bighorn sheep, decline of, 150–151 biodiversity decline, 296–298 bison population, prehistoric, 161–162 conifers and, 124, 133–136 deciduous trees/shrubs, 138–140 elk and pronghorn antelope, 102 erosion in, 221–223, 229 exotics invasion, 115–117 goals for northern range, 324–327 interspecific competition in, 149–150 lightening strikes, 100–101

mule deer, decline of, 165 northern range map, 4 and pre-Columbian aborigines, 327 pronghorn antelope, decline of, 167 public information program, 314–315 public relations information, 280, 322 research and, 308–310, 312–316, 322–323 sage grouse, decline of, 123 95 sagebrush in, 94–95 sagebrush-steppe system, 91 Serengeti comparison, 271–272 sheet erosion in, 215–224 and synthesis design, 51–55 ungulates per 100 elk, 152 vegetation of, 259–260, 287–288, 290–292 Yellowstone River, 172, 223 4 , 35, 48, Yellowstone River valley, 3–4 186 Yellowstone Science, 314 YNP. See Yellowstone National Park (YNP)

371

Yellowstone's Destabilized Ecosystem: Elk Effects, Science, and Policy Conflict

Urban and Regional Policy and Its Effects

A Century of Ecosystem Science

Introduction To Forest Ecosystem Science And Management

Elk Tactics

The New ICT Ecosystem: Implications for Policy and Regulation

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