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

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.

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

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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

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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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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

History of the Northern Range Dispute

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-

History of the Northern Range Dispute

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

History of the Northern Range Dispute

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

History of the Northern Range Dispute

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

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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

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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

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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 Census Period: 1923–2003

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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ELK POPULATION CHANGES AND THE SYNTHESIS DESIGN

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.

The Census Period: 1923–2003

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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ELK POPULATION CHANGES AND THE SYNTHESIS DESIGN

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

The Census Period: 1923–2003

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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ELK POPULATION CHANGES AND THE SYNTHESIS DESIGN

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.

The Census Period: 1923–2003

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 Census Period: 1923–2003

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-

Prehistory to the 1880s

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:

Prehistory to the 1880s

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.

Prehistory to the 1880s

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

Prehistory to the 1880s

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-

Prehistory to the 1880s

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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The First “Experiment”: 1878–1923

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

The First “Experiment”: 1878–1923

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

The First “Experiment”: 1878–1923

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-

The First “Experiment”: 1878–1923

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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The 132-Year Population Trajectory and Associated Synthesis Design

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

The 132-Year Population Trajectory and Associated Synthesis Design

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

The 132-Year Population Trajectory and Associated Synthesis Design

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.

The 132-Year Population Trajectory and Associated Synthesis Design

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

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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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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.

Influences on Upland System Structure I: Aspen Woodland

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-

Influences on Upland System Structure I: Aspen Woodland

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.

Influences on Upland System Structure I: Aspen Woodland

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

Influences on Upland System Structure I: Aspen Woodland

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.

Influences on Upland System Structure I: Aspen Woodland

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

Influences on Upland System Structure I: Aspen Woodland

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).

Influences on Upland System Structure I: Aspen Woodland

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.

Influences on Upland System Structure I: Aspen Woodland

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.

Influences on Upland System Structure I: Aspen Woodland

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.

Influences on Upland System Structure I: Aspen Woodland

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

Influences on Upland System Structure I: Aspen Woodland

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

Influences on Upland System Structure I: Aspen Woodland

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

Influences on Upland System Structure I: Aspen Woodland

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-

Influences on Upland System Structure I: Aspen Woodland

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.

Influences on Upland System Structure I: Aspen Woodland

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

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

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.

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

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.

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

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.

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

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 (<22 cm) sage

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Figure 7.5 East Gardiner exclosure and surrounding terrain. Note virtual absence of sagebrush outside exclosure. Taken July 17, 1997.

bushes at each of 33 sites across the northern range (total of 1,320 plants) and back-dated to their years of establishment. They found a concentration of stand establishment dates during the 1950s to early 1970s among the large plants and a second establishment period in the latter 1980s among the small plants. INTERPRETATION. Investigations of this plant species on the northern range, though not consistent on every point, allows a number of reasonably probable inferences of the effects of elk on sagebrush vegetation. I will pose these chronologically. The first is the question of sagebrush abundance prior to and in the early years following park establishment. Fire is an important consideration. Sagebrush is highly sensitive to fire and is readily eliminated by a burn. Houston (1973) found evidence of a 20- to 25-year fire interval on the northern range prior to park establishment. But Singer and Renkin (1995) comment that there were no general fires across the northern range grasslands between 1870 and 1988. Sagebrush is slow to reestablish following fire. Wambolt et al. (1999) found sagebrush recovery 19 years after one burn in the northern range to be “minimal.” At seven postburn sites, sagebrush “had not recovered” after 10–19 years. Houston (1982:107) speculated that reestablishment “might take 10–30 years.” Moreover, there are questions about ignition sources, timing, and frequencies. Kay (1990) tabulated lightning-strike dates for the Yellowstone area and found these largely occurring between June 1 and September 15 during the time vegetation was active and nonflammable. Despite 4 strikes/km2/year, ignitions

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

were rare. Kay, although embracing the idea of fire frequency, hypothesized aboriginal burning as the major causation. Thus, fire may have played some role in sagebrush distribution and abundance in the years prior to and following park establishment. But the fact that most of the early sagebrush photographs prior to 1920 (figure 7.3) show moderate to abundant amounts of the species suggests that it had not played a large role in the 2 or more decades prior to the photography dates. The two photographs among the 22 early ones that show clear evidence of recent fire and probable elimination of sage are Meagher and Houston’s plates 52 and 53 showing 1885 scenes. The evidence indicates that sagebrush was a major component of the vegetation at park establishment and into the first two decades of the twentieth century. This was the case both in the higher- and lower-elevation ranges. There may have been some degree of patchiness as a result of small local fires. Its density may have been lower in the soils of the Gardiner area (figure 7.5) than at the higher elevations. Houston (1982:94) cites early travelers who commented on the austerity of the area. But elsewhere in the lower range, early photos show significant amounts of the species in the vegetation (compare figure 7.2, photos cited in figure 7.3). The early part of this period up to the middle 1880s coincides with the low elk population size and winter movement outside the park (chapter 3). That the apparent sagebrush abundance continued for 3 to 4 decades into the first elk population high (chapter 5) needs some rationale and will be explored later. By at least the 1920s, the species was in sharp decline in both the higher and lower elevations of the winter range, reaching very low levels in both by or before 1958. With no research in the park prior to the 1920s, it is not possible to suggest how far back the condition occurred. But by the 1920s, the northern range had been subjected to 3 or 4 decades of use by a large elk herd. One might be inclined to suspect that the decline was more extreme in the lower range, given the austere appearance of the Gardiner area today (figure 7.5) and the greater contemporary abundance of sagebrush at the higher elevations. But the 3 sets of park measurements in 1958 and 1962 all indicate similar low abundances at both higher and lower elevations (figure 7.4). Second, the park cover measurements show that sagebrush increased between the initial 1958–62 measurements and 1990, both inside and outside the higher exclosures but only inside the lower-elevation ones. Abundance in the lower areas remained low during the period of measurement and up to the present (figure 7.4). Wambolt and Sherwood (1999) attribute the lower amount of sagebrush outside the exclosure in the Gardiner area to lower site potential in terms of lower precipitation and less favorable soils and to the fact that the lower snow accumulations leave the plants more exposed to browsing. In addition the area is used by three species of browsing ungulates: pronghorn and mule deer as well as elk. Singer and Renkin (1995) attributed most of the browsing pressure in the area to pronghorns because sagebrush constitutes a much larger proportion

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(48.7%) of their diet than of elk diet (3.8%). But proportion of diet alone is not a valid criterion of utilization (proportion of vegetation browsed), as Bishop et al. (1997) recognize. Key variables in the equation are the numbers of ungulates of each species and the amount of vegetation consumed per animal. Singer and Renkin’s (1995) surveys of ungulate numbers in the vicinity of the exclosures at the time of their studies showed elk outnumbering pronghorn by about a factor of 2. This must be conservative for the typical ratio of wintering elk and pronghorn in the area in winter. From 1971 to 1996, spring pronghorn counts on the YNP winter range varied between 102 and 588 (Bishop et al. 1997:126) and averaged 268. They have not exceeded 900 since 1908, and mostly have fallen well below. In the same 1971–96 period, the “estimated population” (Bishop et al. 1997:112) of the northern elk herd averaged approximately 15,000, a conservative number compared with the estimates discussed in chapter 2. I do not suggest that all of the northern range elk winter in the area, but it does become a concentration region. Singer and Renkin themselves cite Houston (1982) who “concluded densities of ungulates in the BLA [Gardiner vicinity] were unnatural and artificially high due to animal avoidance of hunting outside the park.” This comment must apply primarily to elk, given the differences in population sizes. Vore (1990) also concluded that “movement and distribution data from 1984–87 show that hunting concentrated elk at the hunting/ no-hunting boundary.” Even if only a tenth of the northern herd wintered in the area, it must have exceeded pronghorn numbers by at least an order of magnitude. Moreover, the 4.8× larger body size of elk (Houston 1982:157) implies an equivalent, larger food intake. Despite the lower percentage of sagebrush in the elk winter diet, total consumption by elk in the Gardiner area over the years must have far exceeded, perhaps by two orders of magnitude, that of pronghorn when numbers and per animal consumption are factored into the calculations. Mule deer, on the other hand, have largely moved out of the park in winter (Singer and Renkin 1995) and probably have not had a major role in the sagebrush dynamics (see figure 9.5). In discussing the status of sagebrush in the lower part of the range, Bishop et al. (1997:45) focus on the Gardiner area, suggest that its condition may result in part from the effects of earlier private ownership, and repeat the Houston suggestion that the decline of sagebrush may be a return to “pristine” conditions. In either case, the effect is to discount possible elk influence. But the exclosure evidence clearly shows that the area has the potential to develop robust sagebrush vegetation if protected from browsing (figures 7.4, 7.5). And the elimination of sagebrush extends more broadly over the lower range. Although hunting outside the park boundary is a deterrent to free elk movement northward, some movement does occur as mentioned, and with it pressure on the vegetation in that portion of the northern range outside the park boundary. Carl L. Wambolt (personal communication, August 3, 1999) reported seeing a USDA Forest Service photograph taken in the early 1900s that showed sagebrush cover over Deckard Flat. When Wambolt first saw the area in 1980,

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

the sagebrush had largely disappeared (figure 7.6). Wright and Thompson (1935) commented on its decline on Mount Everts. The failure of the sagebrush to increase in the Gardiner area during the herd reductions may have been due to the herd’s tendency to concentrate at the west end of the range, even with reduced populations. Houston (1982:398) commented that “Big sagebrush remained heavily browsed [in the Gardiner area] from 1964–69 (88%) and to 1974 . . . even at lower ungulate densities.” Park publicity (Bishop et al. 1997; Finley 1997) also implies that these effects are unique to the BLA (Gardiner area). But the photographic evidence for areas of the lower winter range other than the BLA (figure 7.2), and the anecdotal observations for Deckard Flat (figure 7.6) and the west slopes of Mount Everts, indicate that the sharp reduction is a more general phenomenon of the western portion of the lower northern range than just the BLA. By 1958–62, sagebrush abundance on the northern range had been subjected to browsing by a greatly increased elk herd for more than 70 years and had been driven to extremely low levels (figure 7.4). The subsequent sagebrush increase inside all of the exclosures was almost certainly facilitated by protection from browsing. The increase outside the higher elevation exclosures (figure 7.4) began during the herd reductions when the censuses were producing num-

Figure 7.6 An early USDA Forest Service photograph showed Deckard Flat, in the lower winter range just outside the park boundary, covered with sagebrush growth (Carl L. Wambolt, personal communication, August 15, 1999). This had largely disappeared when Wambolt arrived in the area in 1980. Remnants remain in small pockets in the lower left, lower center, and far right of the photo. Photo taken August 15, 1999.

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bers well below 10,000 (Bishop et al. 1997). Barmore (1980) commented on the higher-elevations exclosures: “An increase of shrub aerial cover, primarily between 1962 and 1967, on both protected and unprotected plots may have been proportionately greater in the absence of ungulates” (p. 299). In referring to the reduced browse utilization rates at the higher elevations during the herd reduction period (p. 302), he observed: “This strengthens the conclusions that increases in aerial cover, particularly after 1962, were due to reduced browsing.” Wambolt and Hoffman (2001) attributed the establishment spurt during the 1950s–70s, shown by their dendrochronological studies already cited, to the population reductions of that period. The herd remained below a censused 10,000 level until 1974 (see table 2.1), halfway through the 1958–90 span of park measurements at the exclosures. Sagebrush increase continued outside the upper exclosures (figure 7.4; Norland and Reardon 1996) during the second half of this span, but to a lesser degree than the inside recovery. Thus the browsing pressure was not so intense as to prevent this outside increase, even with the large herd sizes of the latter 1970s and 1980s, but not to the degree inside the exclosures. By 1994, the expansive Wambolt and Sherwood studies showed a 2-fold difference in canopy cover, density, and production per plant and a 3-fold difference in production per unit area between the insides and outsides of the upper fenced areas. Because of the large differences in sampling design, it is not clear whether this large inside-outside disparity reflects the downturn after the 1958– 90 increase shown by the park measurements. These authors attributed the difference between theirs and the Singer and Renkin results to the more extensive sampling (24 times as many transects and 30 times as much area sampled) and to their stratified design, which partitioned out variance associated with the considerable topographic and soil variations at the sites. But Wambolt (personal communication, November 20, 1998), who had been studying sagebrush on the northern range since 1980, opined in 1998 that the species outside the upper exclosures was declining (see figure 7.7). Thus a diversity of time-based evidence sources—the early and Barmore’s anecdotal reports, the photographic evidence (figure 7.3), periodic inside-outside exclosure measurements (figure 7.4), and the Wambolt and Hoffman dendrochronological measurements—point to variations in sagebrush abundance paralleling changes in the northern herd. None of the sources spans the entire 132-year period, and each has its weaknesses. But they are internally consistent, and there is no significant body of contrary evidence. The Rens study indicates continued browsing impact through 1998. Thus the effect persisted 10 years into the recent period, in which one-third of the herd has moved out of the park in winter, on average; 1–2 years into the period in which the wolves have been dispersing the herd; and throughout the period in which the herd has tended to drift toward the northwest end of the northern range. The inside-outside measurements on cover (Singer and Renkin 1995; figure 7.4; table 7.1; Rens 2001) provide a second dimension of evidence on im-

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

Figure 7.7 Appearance of sagebrush in much of the higher-elevation range. Photo is a southfacing slope in the Lamar River valley. Note the low density and short, hedged stature of the shrubs, which may not protrude above snow depth in winter. Compare with plants in exclosures (figures 7.1, 7.9). Photo taken July 17, 1997.

pact. The density measurements are less consistent, Singer and Renkin’s table 6 showing essentially similar inside-outside densities at the higher elevations, and higher seedling numbers on the outside. But the more extensive and recent Wambolt and Sherwood results show 1.6× higher inside densities (table 7.1). The higher outside seedling numbers, as Norland and Reardon (1996) discuss, may reflect reduced intraspecific competition associated with lower densities that enable initial seedling survival. But ultimate recruitment of mature plants into the population, as the Wambolt and Sherwood data show, is higher in the unbrowsed interiors. Singer and Renkin (1995) commented that sagebrush in the high-elevation winter range is “stable or increasing,” a view repeated by Bishop et al. (1997) (“holding its own or increasing”) and by Singer et al. (2003) (“healthy, vigorous, and even increasing”). The National Research Council review (Klein et al. 2002) followed suit with “these areas do not show sagebrush decline and may even show increases” following the curious comment that “These areas [94% of the park portion of the northern winter range] are important for ungulates during the non-winter portion of the year.” None of these sources cites Wambolt and Sherwood (1998), and these inexplicit descriptions fail to characterize the actual ecological status of sagebrush on the northern range. All of the evidence is circumstantial, but there is little

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doubt that elk browsing significantly affects the structure and function of the sagebrush vegetation, and consequently the entire shrub-steppe subsystem, as will be discussed. Browsing pressure has essentially eliminated sagebrush at the lower elevations (figures 7.4, 7.5). Norland and Reardon (1996) comment that it could become locally extinct in that portion of the northern range. Effects at the higher elevations are clearly shown by Wambolt and co-workers, including a two-thirds reduction in sagebrush forage production. Norland and Reardon (1996) comment: Sagebrush is an important forage item for wintering ungulates because it is a good source of nitrogen and energy during a time of year when the other forages are low in nitrogen and energy. . . . The reduction in forage availability of sagebrush means that the opportunity for ungulates to forage on a high protein, high energy source is reduced. After ~70 years of browsing by a large herd, sagebrush had virtually been reduced to vestigial status (figure 7.4). By the end of the 1900s, the herd had essentially recovered to high numbers for ~20 years. Whether these effects change during the new wolf era remains to be seen. But that uncertainty emphasizes the importance of continued monitoring on a scale comparable with the Wambolt et al. studies. In closing this discussion on Artemisia tridentata in the northern range, four questions need to be addressed to place the species in its proper ecological perspective. First, Barmore’s (1980) and Houston’s (1982) comments on sagebrush imply a view of the species that is essentially the range management “increaser” paradigm (compare Allen-Diaz and Bartolome 1998, for an excellent example): “These data seemed consistent with the hypothesis that the decrease in big sagebrush over most of the [Gardiner] area represented a return to more pristine conditions following a grazing disclimax established and maintained by livestock” (Houston 1982:128 and repeated by Bishop et al. 1997). The reference to livestock in Houston’s comment refers to the fact that a 1,600-ha area of private land in Montana (the BLA), north of what was then the park boundary and now in the lower area containing the Gardiner exclosures, was acquired by the park in 1932. Prior to that time this area had been privately owned and used for livestock grazing. The exclosure evidence now clearly shows that sagebrush is an effective competitor in the northern range area and, in the absence of browsing becomes the dominant species. It needs no grazing of perennial grasses to facilitate its increase, although such grazing may enhance its expression further. It has persisted without grazing, and without being succeeded by grasses, for nearly half a century. Houston’s reference to the pristine condition in the quote—presumably that prevailing at the time of park establishment—appears to imply a condition of no livestock and absence of sagebrush. But I have presented evidence that the pristine condition was a vegetation with a significant amount of sagebrush and probably dominated by it on favorable sites more like the conditions now pre-

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

vailing inside the exclosures. The decline of sagebrush in the Gardiner area to which Houston refers, rather than being a return to pristine conditions, is almost certainly due to heavy browsing by a greatly increased ungulate population, especially elk. A second question arises as to why the impacts on the sagebrush were not observed, or at least reported, until the 1920s and 1930s when the herd had existed at high levels for 40–50 years. There may be two reasons. As discussed in chapter 6, the first biologist, M. P. Skinner, arrived in the park in the 1920s. Except for the Smith et al. (1915) one-year excursion and the report by Graves and Nelson (1919), there had been no prior investigation. Hence there were no professional persons to make the observations earlier. As mentioned previously, sagebrush is less likely to attract attention, aesthetically, than aspen and riparian vegetation. A second reason could involve the chronology of vegetation change, a point mentioned in chapter 5. One cannot know how much time elapsed before the effects of the herd build-up at the end of the nineteenth century started to become evident on sagebrush. The herd build-up was gradual, and some emigration out of the park probably continued for a time after park establishment. An extensive sagebrush steppe, covering half of the northern range, was probably able to absorb gradually increasing use for some time before an extreme effect became evident. As discussed in the last chapter, the first observations of significant effects on aspen and willow were reported by Smith et al. (1915), perhaps 25 years after major increase in elk numbers, although the aspen-aging evidence indicates marked effects by or before the turn of the century. The delay has also been attributed to a shift down the palatability scale. Skinner (1928) commented, “But aspen . . . cottonwood . . . and coniferous browse, especially the needles of Douglas fir . . . were better liked [than sagebrush]. . . . Elk even stood on their hind legs to reach and pull down the lowest branches of limber pines . . . red cedar . . . and cottonwoods.” This widespread stereotype of low sagebrush palatability is now being challenged by recent research (Welch and McArthur 1979). A third question is the seeming paradox between the reportedly low sagebrush utilization rates and the evident impacts on the species at the higher elevations. Singer and Renkin (1995) reported 11.6% of current annual growth browsed between 1985–88 at the higher elevations compared with 86.8% at the lower elevations, although C. L. Wambolt (personal communication) believes the utilization at the higher elevations may be as high as 40%. The resolution would seem to lie in the morphology and growth pattern of the species. Sagebrush is widely considered among range ecologists to be browsing intolerant (Billbrough and Richards 1993). In simulated-browsing research designed to explain this sensitivity, Billbrough and Richards (1993) found that “after even a moderate level of [shoot removal] . . . node replacement was generally lower than controls, resulting in fewer potential growing points the following year.” Because the plants did not allocate additional resources to the new growth points, the growing shoots “are unable to produce sufficient biomass to

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maintain whole-plant production at higher intensities of. . . . [shoot] removal. . . . Thus, Artemisia was more susceptible to damage caused by even moderate browsing levels.” Fourth, a comment on exclosures that could have been made appropriately in the last chapter is also relevant here as well as in later chapters. Since proposal of the natural-regulation hypothesis, park attitudes and publications have repeatedly criticized their relevance. Some have been dismantled, and elimination was proposed for others until protests by outside investigators deterred the action. The structures have been criticized on the grounds that the conditions inside are unnatural, containing vegetation that does not coexist with and sustain the effects of ungulate use. Conditions in the park outside the exclosures are considered by park personnel to be the more natural. There is the further value judgment that these conditions are the ones appropriate for national parks. This entire rationale is based on the unstated premise, fundamental to the natural-regulation paradigm, that elk were present in large numbers in prehistory in the Yellowstone area, as were the current conditions of the vegetation. But as discussed in chapter 3, the weight of the evidence now points to low elk numbers in prehistory and migration out of what is now the park in winter. Moreover, the historic and photographic evidence shows a considerably different vegetation at park establishment than what prevails today. Rather than being irrelevant or unnatural, the conditions inside the exclosures in all probability more nearly approximate those of prehistory, as Kay (1990) suggested. The exclosures thus resemble past conditions that endure nowhere else in the park portion of the northern range and indeed if at all on the northern range, even outside the park. Their removal would eliminate any vestiges and physical evidence of what the pristine conditions once were as long as the elk herd remains at high levels. Whether the conditions inside or outside the exclosures are the appropriate ones for the park is a value and policy question not addressed here. In sum, (1) the abundance of sagebrush early in the park’s history when elk numbers were low and/or migrating out of the northern range in winter; (2) its sharp decline to low levels following decades of use by a greatly increased herd, mostly inhibited from the winter exodus; (3) its virtual elimination from the lower elevation concentration area; (4) its increase during the herd reduction period, more markedly inside the exclosures; (5) the Wambolt and Hoffman (2001) evidence of recruitment pulses when the herd is reduced; (6) the more pronounced postfire recovery inside the Blacktail exclosures; and (7) the current inside-outside exclosure differences all leave little doubt that elk browsing has significantly affected sagebrush abundance on the northern range. The extent of effect has depended on the number of animals and their tenure of use. What the trend is at present is not known, but it certainly should be monitored with an extensive sampling scheme.

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

Other Shrubby Species In addition to big sagebrush, there are several species of small shrubs in the sagebrush-grass steppe that have been variously called sprouting shrubs (Wambolt and Sherwood 1999), dwarf shrubs (Coughenour et al. 1996), and subshrubs (Singer 1995): winterfat (Krascheninnikovia lanata), green rabbitbrush (Chrysothamnus viscidiflorus), big rabbitbrush (C. nauseosus), spiny horsebush (Tetradymia canescens), and others. Collectively, these occur at roughly half the densities of big sagebrush in the vicinity of the exclosures, but at an order of magnitude lower cover because of their small stature. Nonetheless, they contribute some measure of structural and species diversity to the sagebrush-steppe vegetation. Wambolt and Sherwood (1999) concluded that “response of the sprouting shrubs was similar to that of big sagebrush.” Both mean canopy cover and mean density were higher inside the exclosures than outside at the 19 sampling sites in their 1994 studies, although only cover values were significantly different. However, the data in their tables 5 and 6 show that the consistent differences were at the Gardiner exclosures: canopy cover was higher inside the exclosures at all of the 8 sampling sites, density was higher inside at 7 of the 8. The two parameters were about evenly divided in opposite directions at higher elevation exclosures, even though on average they were higher on the inside. These differing responses can be interpreted in terms of two major environmental variables operating on these species: negative effect from big sagebrush competition; and elk herbivory operating both negatively through browsing and positively by browsing sagebrush and moderating its competitive pressure. At the Gardiner exclosures (figure 7.5), browsing pressure on the outside is so intense that the small shrubs have been driven to extremely low densities (1/15 m2), and cannot benefit from the near elimination of big sagebrush (table 7.1) and its competition. Inside the exclosures, where they are protected from browsing, they can coexist sufficiently with sagebrush to achieve a density of 1/9 m2, roughly half the sagebrush density. At the higher elevation exclosures, browsing pressure is lighter. It is also lighter on sagebrush, but still sufficient to reduce density of the latter to about half that inside the exclosures. Consequently the small shrubs on the outside achieve ~84% of the inside density (1/2.5 m2 versus 1/3 m2), where they must be coexisting with significant sagebrush competition. Wambolt and Sherwood (1999) comment in a similar vain: “On the NYWR the dominant big sagebrush has been reduced from intense browsing, thereby allowing the subdominant sprouting shrubs to increase.” In sum, at this point in the history of the northern range, elk browsing appears to have nearly eliminated the small shrubs from the lower winter range and at the higher elevations has reduced them to a density about 16% below that of unbrowsed vegetation. The best hypothesis might be that the 1872 condition is approximated by their current status in the exclosures.

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Herbaceous Vegetation Context Assessing whether there are changes in herbaceous vegetation structure associated with ungulate use is more difficult than for woody vegetation because ideally cover, density, and production or biomass of the entire herbaceous vegetation, and at least the major species, need to be measured over time. The time and effort cost would be considerable. The difficulty is compounded by the greater year-to-year variability of herbaceous cover and production or biomass and, in the case of annuals, their densities and actual species present in any given year. Houston (1982), Merrill and Boyce (1991), Coughenour et al. (1996), and others have concluded appropriately that vegetative production varies markedly from year to year with annual variations in precipitation. Such correlations have been shown repeatedly for arid and semi-arid vegetations worldwide (compare Le Houérou and Hoste 1977; Wagner 1980; MacMahon and Wagner 1985). Thus, inferences about trends in vegetation based on point-in-time measurements must be tempered with the reality of this stochastic variation. However, herbaceous vegetation competes with sagebrush. And it is affected by elk grazing, both from direct herbivory effects, and indirectly from variations in sagebrush competition associated with varying browsing impacts on the latter. Thus, a number of factors complicate assessment of the effects of ungulates on the northern range herbaceous vegetation. The considerable amount of research on this component of the system has examined both structure and function. I will address the structure aspects here and examine function in later chapters.

Relationships between Elk Numbers and Herbaceous Vegetation Historical Evidence There are parallels between the chronology of evidence on sagebrush and that on the herbaceous vegetation. No research was started on either until the 1920s. Until such studies were begun, the only available evidence was photographic and anecdotal. Given the difficulty of discerning much about herbaceous vegetation from photos, the total amount of early evidence is considerably less than that for sagebrush. Whether the fact that accounts of ungulate effects on the herbaceous vegetation appeared 1 or 2 decades later than reports on the woody vegetation is attributable to the lesser visual appeal of the former or in fact the effects occurred later, is an open question. The same question was raised for sagebrush. The first lengthy published report of which I am aware was Rush (1932) who, commenting on the vegetation condition in the Lamar Valley in 1928–29, considered it “badly overgrazed . . . [with] much rich topsoil washed away.” He

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

estimated that winter range had deteriorated 50% since 1914 and that sheet erosion had taken 1–2 inches of soil over more than half of the range. At this point, the northern herd had been at high levels over approximately 40 years. Wright and Thompson (1935) commented on the severity of the range problem, showing two September 17, 1933, photographs in which “there is not enough grass to cover the stones. This is typical of a large part of the elk winter range in northern Yellowstone.” Houston (1982) pointed out that many of these observations were made during the drought of the 1930s. Showing repeat photos of sites during dry and wet years that show stark differences in vegetative standing crop (e.g., his figure VI.7), he generally concluded that the conditions reported by the early observers were attributable to drought rather than ungulate impacts. Undoubtedly the drought did play a role in the vegetation conditions of those years. But Grimm (1939) reported 1931–34 to be the drought years and commented that park officials, recognizing the “bad conditions,” began the first studies in 1928, possibly referring to the Rush study. He commented on the serious denudation and soil erosion. Wright and Thompson (1935) remarked: The range condition portrayed in these pictures is not a merely transient one nor is it of recent occurrence. The pictures have been secured . . . over a period of 4 years. The range was in deplorable condition when we first saw it, and its deterioration has been progressing steadily since then. It is noticeably worse now than it was in 1929. Moreover, Kittams (1952) commented on the serious condition of the northern range 18 years after the end of the drought, showing a May 1951 photograph captioned “Spring runoff carried soil down this draw north of Black Butte Creek and left mounds around sage bushes and almost covered some bushes with soil.” Black Butte Creek is not in the northern range, but it is just outside the northwest corner of the park where the smaller Gallatin herd winters. Thus reports of significant vegetation impacts both preceded and followed the drought years. The 1930s drought may well have exacerbated the effect in those years, but discounting an ungulate involvement ignores evidence to the contrary, inadequate as that evidence may be. Furthermore, the reports of significant vegetation effects in the early 1950s are consistent with the first measurements following 1957 and 1962 construction of the large exclosures, which produced the lowest values of the 32 years of measurement, as will be discussed shortly. There is thus reason to infer significant alteration of the herbaceous vegetation at least by the 1920s, with progressive decline, probably aided by the 1930s drought, to markedly altered conditions until the 1960s herd reduction. As pointed out for sagebrush and will be discussed for the herbaceous vegetation, the measured values for the higher elevations were as low as those taken around the Gardiner exclosures. EXCLOSURE EVIDENCE ON THE NATIVE VEGETATION. Four sets of year-end herbaceous biomass measurements, relative to the exclosures, have been reported

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for the northern range. These are based on varying methodologies, varying series of years, and varying locations and combinations of exclosures. They thus challenge generalization, especially because there are both consistencies and inconsistencies between them. Coughenour and Singer (1996) review production measurements, based on year-end clipping of “volume plots” by park personnel in 1935–41, 1947, 1949, and 1950, all years prior to construction of the large exclosures. And the authors included their own clipping studies inside and outside four upper range and one lower exclosure in 1986–88. It is not stated whether the early studies were conducted at the same general locations of the exclosures. The major conclusions of these studies—reported in an excellent paper primarily addressing elk population processes—were that year-to-year variations in herbaceous production were correlated with variations in several precipitation parameters. The further inferences were drawn that these correlations represent the primary determinants of variations in vegetative production, and elk grazing does not significantly affect that production. The same conclusions were repeated in Bishop et al. (1997). I have three observations in response. First, the authors did not state whether the early and later studies were carried out in the same areas and thus can be appropriately compared. Second, the mean standing crops for the early measurements were reported at 66.6 g/m2 for the higher-elevation ranges and 52.8 for the lower. But no values were cited for the later studies to illustrate the comment on the similarity between the two periods. However, Coughenour (1991) reported other insideoutside exclosure measurements for September 1987 and July 1988 that range from 15.5 to 29.4 g/m2 for the outsides for the upper exclosures, 16.3–18.4 for the lower ones. All are lower than the early means by one-half to two-thirds. Third, the authors’ correlations between precipitation and primary production are certainly in keeping with numerous other studies in arid and semi-arid areas. But the correlations do not imply that production cannot be influenced by other variables, especially ungulate grazing effects. In an analogy, annual elkpopulation rates of change have been shown in several studies to vary with growing-season precipitation and severity of winter weather. But those rates are also correlated with elk population size (figure 2.1). The weather effects produced scatter around the long-term population trajectory. By the same token, the precipitation effects do not necessarily alter a long-term vegetation trajectory that I will suggest shortly. In a second set of studies, Coughenour (1991) reported year-end aboveground biomass measurements in 1987 and 1988 inside and outside 3 higherelevation exclosures and one lower. His results for September 1987 and July 1988 showed significantly higher grass biomass inside one of the three upper 1987 areas and in all three in 1988. Inside-outside differences were not significant in 2 of the 3 in 1987. There were no significant differences in any of the forb measurements or the grass measurements at the lower elevations. Thus the insideoutside indications are inconclusive.

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

Two sets of measurements at all 8 exclosures provide the longest-term data sets: Parker transect counts and the chart quadrats used to measure canopy cover described in the sagebrush discussion (figure 7.4). The early phases of these studies are described at length in Barmore (1980) and Houston (1982). These measurements, inside and outside the exclosures, have been repeated periodically over the period 1958–90, and the most recent reviews of the series are in Coughenour et al. (1996) for the Parker transects and Reardon (1996) for the chart quadrats. Any inferences of changes in the northern range herbaceous vegetation over time can only be drawn from these two data sets, which span a 32-year period during which the elk herd changed from the moderately high levels of the 1950s to the low numbers of the intensive 1960s herd reduction, to the high numbers achieved by population recovery in the latter 1980s. In all probability, the changes in sagebrush abundance have also constituted a second, varying factor in the herbaceous vegetation environment. Sagebrush competitive effects have likely covaried inversely with elk numbers as the latter have changed over time and between the interiors and exteriors of the exclosures. Coughenour et al. (1996) begin their analysis with a lengthy discussion of the statistical inadequacies of the Parker method, and it has largely been abandoned for vegetation management. But the trends shown for the combined herbaceous vegetation and “dwarf shrubs” are the same inside and outside the exclosures and between the upper and Gardiner exclosures although at different levels of abundance (figure 7.8). The major problem with the Parker technique is that it does not provide unbiased comparative measures of the different plant species because it is biased toward those with larger plants. But this bias would not occur in measurements of a given vegetation class if the plants were of a similar size (e.g., grasses or forbs of similar size), or in total-vegetation comparisons over space or time if the species compositions were similar. Hence the close similarities of the four trends in the Coughenour et al. measurements would appear to be valid, if approximate, indices of the 32-year changes. Although the Parker measures were

Figure 7.8 Trends in herbaceous vegetation inside and outside exclosures. Parker transect data from Coughenour et al. (1996); chartquadrat data from Reardon (1996).

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taken in more years than the chart quadrats (Reardon 1996), the trends over time in these two sets of observation are similar (figure 7.8) as follows. As with the sagebrush (figure 7.4), both methods produced low measurements in 1958 and 1962, with the vegetation functioning at lower abundance in the Gardiner area. These presumably reflected the condition of the vegetation in the northern range at the end of the first elk population high when the large exclosures were constructed. The 1967 measurements were similar to the earlier ones, both inside and outside the exclosures. The Parker measurements, both inside and outside the exclosures, increased in 1974 and 1981, then declined each in 1986 and 1989. Chart-quadrat measurements were apparently not taken in 1974 and 1981, but the 1986 measurements both inside and outside the exclosures were 2 to 3 times the early measurements, presumably reflecting the same increase shown by the Parker measurements sometime during the late 1960s and 1970s. The quadrat measurements declined between 1986 and 1990, possibly reflecting the same decline portrayed by the 1986–89 Parker measurements. In sum, both sets of measurements imply herbaceous vegetation abundance, both inside and outside the exclosures, rising from extremely low levels in the late 1950s and early 1960s to some high point between 1974 and 1986. The chart quadrats measured the highest value in 1986. But the 1967–86 measurement hiatus prevents any suggestion that the 1986 value represented a high point. Both methods recorded declines between 1986 and the following measurements: 1989 in the Parker transects and 1990 in the chart quadrats. The quadrat measurements at the Blacktail exclosures were an exception to the others, increasing from the 1986 values to higher ones for 1990, both inside and outside (figure 7.8). I will hypothesize the cause of this exception shortly. Coughenour et al. (1996) concluded from their Parker-transect study that elk grazing had no significant effect on the composition and abundance of the herbaceous vegetation during 1958–89. This was based on the lack of any significant differences between the measurements inside and outside the exclosures. Several authors have inferred from the evidence (see Merrill et al. 1994; Coughenour and Singer 1996; Singer et al. 1998a) that elk use on the northern range does not significantly affect the herbaceous vegetation. This is generally attributed to the fact that the plants are dormant and snow-covered during the elk winter use. Coughenour and Singer (1996) also hypothesized that the 32-year trends were driven by year-to-year weather variation. This was based on the comparable trends inside and outside the exclosures and on 19 tests showing correlations between various weather parameters and vegetation groups or individual species. Such variations do appear in the Coughenour et al. (1996) analyses of individual species (their figures 2, 3). But their trends for the combined grasses, which dominate the combined herbaceous vegetation, are less variable. Moreover, 11 of their 19 correlations with weather variables (their table 3) are with the proportions of each vegetation class or species. Proportions of a species can vary without any change in its absolute amount if the amounts of other species (and

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

their proportions) vary and will tend to do so inversely with the changing species. The result can produce correlational artifacts. Thus one is inclined to be suspicious when grass and forb percentages in the higher-elevation exclosures have almost identical correlation coefficients with previous-year winter precipitation but of opposite signs (their table 3). Similarly, forb and dwarf-shrub percentages are both correlated with spring precipitation, but the former positively and the latter negatively. I will propose an alternative set of inferences below on the effects of elk on herbaceous vegetation, but I will cite some additional evidence here before doing so. In his 1998 studies, Rens (2001) measured higher abundance of perennial grasses outside the Blacktail exclosures than inside. Although the differences are slight and probably not statistically significant, the Parker transects and chart quadrants consistently measured greater herbaceous vegetation outside the higher elevation exclosures from 1974–90 (figure 7.8). The same comparisons are mixed between the 2 data sets for the Gardiner exclosures. Evidence on whether or not there have been significant variations in the relative abundance and diversity of native herbaceous species that could be associated with elk grazing is ambiguous. There have been changes over time in the absolute abundance of the herbaceous vegetation (figure 7.8), so that abundance of the constituent species must have varied in time and space. And I have suggested that these changes could have been induced by a combination of varying intensities of herbivory and sagebrush competition. Coughenour et al. (1996) recorded marked variations between years in the proportions of several species inside and outside the exclosures with the Parker transects, which they attributed to variations in weather conditions. The measures were highly variable, and the authors prefaced their analysis with cautions to the reader about the statistical shortcomings of the methodology. It is not clear whether the method can adequately detect changes in gross vegetation classes, such as total grasses or forbs, but has difficulties at the level of individual species. Reardon (1996), measuring vegetative cover with chart quadrats, found no change in species composition and “No significant correlations . . . between the weather parameters and plant community characteristics.” In a two-year study, Singer (1995) sampled biomass, number of species, ground cover, and morphological and nutritional characteristics of herbaceous vegetation and subshrubs inside and outside the higher-elevation exclosures and percent cover at the Gardiner exclosures. The number of grass (x\ = 4.5), forb (17.5), and shrub (3.5) species “did not differ between grazed and ungrazed plots.” Thus if there have been changes in the composition of the native herbaceous vegetation attributable to elk grazing, they have not been detectable with the existing intensity, frequency, and methods of measurement. Reardon speaks of the low statistical power of the methods. Alternatively, there may in fact have been no significant alteration by elk grazing. INVASION OF EXOTICS. Although Despain et al. (1986) commented that “nonnative grasses, such as cheatgrass, [Bromus tectorum] have not invaded ranges

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managers used to call overgrazed” (p. 79), park investigators have been reporting such invasions since the early 1930s (Rush 1932; Wright and Thompson 1935; Grimm 1939; Kittams 1948). More recent quantitative investigations have reinforced these reports. Houston (1973), in studying the relation of fire frequency to aspen distribution in the northern range, observed: “It seems improbable that other remnant [aspen] stands with understories now dominated by the exotic grass Phleum pretense [sic] would respond to burning. This grass has completely altered composition of the understory vegetation.” Other investigators have reported similar observations. Wallace and Macko (1993) commented that timothy (Phleum pratense) “is a highly invasive exotic species that has occupied large areas of Yellowstone’s northern range.” In their view, the second most invasive exotic species is Kentucky bluegrass (Poa pratensis). The evidence that ungulate grazing stimulates expansion of exotics is again provided by inside-outside exclosure comparisons and inside-outside park comparisons. I reviewed Kay’s (1990) results in the last chapter in which he observed diverse vegetation understories in aspen stands inside exclosures with only a minor grass component. But outside the exclosures the understories were largely graminoid and dominated by timothy and Kentucky bluegrass. Singer (1995) measured plant cover and standing green biomass inside and outside the Gardiner exclosures in 1986 and 1987. Although he measured no significant inside-outside differences on ridges and slopes, grass cover at swale sites outside the exclosures was more than twice that inside the exclosures (his table 6), with the exotic Kentucky bluegrass and cheatgrass making up 73% of the total. He commented: “Apparently, grazing by native ungulates, especially by elk, contributed to the persistence of exotic grasses, while complete protection from ungulates since 1958 contributed to their decline.” Frank and McNaughton (1992) studied plant production at 4 higher-elevation winter-range and 3 transition-range sites. Presumably focusing their investigations on the dominant graminoid species at the sites, they studied 3 species of exotic grasses at 2 winter and 2 transition sites: timothy, Kentucky bluegrass, and smooth brome (Bromus inermis). While measuring no significant insideoutside differences at the higher-elevation exclosures, Singer (1995) commented on the Frank and McNaughton (1993) study: “The dominance of the Lamar Valley study sites of Frank and McNaughton (1993) by exotic grasses also suggests heavy grazing by native ungulates, both elk and bison, contributes to the persistence of the Eurasian grasses.” As discussed in the last chapter, Kay (1990) measured understory vegetation of aspen stands in the northern range inside the park and in the Eagle Creek area outside the north boundary. Mean cover of grasses was 58.8% inside the park, 18.1% in Eagle Creek. Timothy and Kentucky bluegrass were again the dominant grasses, their combined canopy covers averaging 37.9% and 8.2%, respectively. On July 16, 1997, I photographed an almost pure stand of the annuals cheatgrass and allysum (Allysum allysoides) outside the east fence of the west

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

Gardiner exclosure (figure 7.9). This constitutes the almost complete reduction of diversity by maximum disturbance in the Connell (1978) intermediate disturbance model. In sum, exotic plant species are a conspicuous component of the northern range herbaceous vegetation. Their greater abundance outside the exclosures than inside, and inside the park than outside, is circumstantial evidence that ungulate grazing inhibits the competitive ability of the native species sufficiently to allow persistence of the nonnatives. That the latter are relatively inconspicuous inside the exclosures indicates that they cannot compete effectively in lightly grazed or ungrazed native vegetation.

THE SAGEBRUSH-STEPPE ECOLOGICAL SYSTEM The Vegetation and Its Ungulate Use At 43,900 ha, the sagebrush steppe occupies slightly more than half of the northern range area inside the park and about 44% of the 100,000 ha Houston (1982) considered to be the northern range. With the elk increasingly moving out of the park in winter, the newer values for the northern range area rise to 152,663 ha (Lemke et al. 1998), but even in this terrain, sagebrush steppe still consti-

Figure 7.9 East side of the west Gardiner exclosure. Outside vegetation has been reduced to almost complete dominance by two exotic annuals, Bromus tectorum and Allysum allysoides. Photographed July 16, 1997.

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tutes a major fraction. In total, it is the most extensive vegetation type of the northern range. As such, it plays a major role in the structure and function of the northern range ecosystem. It provides the majority of the winter forage for 5 of the 7-species ungulate guild and year-round food for an herbivorous small mammal and arthropod fauna largely unstudied in the park. It plays a large role in the northern range energy flow and nutrient cycling. And it functions in northern range hydrologic process, stabilizing soils and surface hydrology, as will be discussed in later chapters. Most of the research on the northern range shrub-steppe vegetation has focused on the major vegetative components individually. But those components function in a matrix of biotic interactions and collectively form a community or species assemblage. Any factor significantly affecting one component would be expected to have second-order effects on the others. Park research shows evidence suggesting intraspecific competition in sagebrush. Norland and Reardon (1996) showed sharp decline in sagebrush density between 1962 and 1990 in the Mammoth exclosure, the one with the most luxuriant sagebrush growth. They suggested this was due to self-thinning. Singer et al. (1998) reported 6,700 sagebrush seedlings per ha in grazed sagebrush, 300/ha in ungrazed stands. This difference could indicate that the lower densities of mature plants in the grazed stands permit coexistence of the seedlings while the protected communities are closed. Interspecific competition between sagebrush and perennial grasses has been well chronicled (see Laycock 1967; Tueller and Tower 1979; Bork et al. 1998). Elk, which feed on both, can potentially impact both directly by tissue removal and indirectly by modifying their competitive relationships. There is also evidence, discussed in a later chapter, that ungulate feeding accelerates part of nutrient cycles by grazing, digesting, and excreting nutrients, thereby fertilizing the vegetation. The species that occur in a given community are obviously, but not necessarily perfectly, adapted to the physical conditions of an area. In the absence of herbivory or other significant disturbance, as in exclosures, the species that prevail in the competition do so because they are best adapted to the physical conditions or can compete more effectively by virtue of their biological characteristics, whether demographic, morphological, or physiological. The endpoint of competition in semi-arid areas may result in dominance by a low diversity and small number (e.g., 2 to 5) of competitively superior species (Holmgren and Hutchings 1972; Wagner 1987). The dominance of sagebrush in the northern range exclosures is a case in point. Any disturbance, including herbivory, that affects the competitive ability of the species can limit their expression in the community. If the effect disproportionately impinges on the competitive dominants or is more evenly applied to the entire community but includes effects on the dominants, it can facilitate coexistence of otherwise competitively inferior species. The result is enhancement of diversity. This scenario has a long history in ecological theory, applied to a wide range of both plant and animal communities: Physical dis-

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

turbance or predation/herbivory are commonly invoked as factors which promote diversity. How ungulates affect the competitive balance between plant species in a community depends on (1) their feeding specialties (e.g., grazers or browsers); (2) the season of feeding (vegetation is often more susceptible to damage during the growing season than during its season of dormancy); (3) accessibility of the vegetation to the animals; (4) palatability of the species, including presence or absence of herbivory defenses such as thorns and secondary compounds; and (5) how hard-pressed the animals are for food. The season-of-use effect has been reported by Laycock (1967) for domestic sheep grazing on the USDA Dubois, Idaho, Research Station about 100 km west of Yellowstone. When grazed in summer, sheep feed primarily on herbaceous vegetation and shift community structure to a dominance of sagebrush. When grazing is deferred until fall, and herbaceous vegetation becomes dormant and of low quality, the sheep browse sagebrush more intensively, reduce its presence in the vegetation, and shift the structure toward grassland. An analogous case is the effect of browsing mule deer on vegetation structure of foothill winter ranges over much of the Intermountain West. Increasing numbers of deer during the twentieth century reduced shrubby vegetation including sagebrush and favored perennial grasses (Smith 1949; McArthur et al. 1988). Using both sagebrush and herbaceous vegetation largely in winter, elk affect northern range vegetation structure in ways that can be interpreted in these contexts. With live tissue projecting through the snow in winter, and a morphological sensitivity to browsing, sagebrush abundance has waxed and waned with the number of animals and their tenure of use. At the extreme it was driven to very low levels over the entire northern range after 7 decades of use by a large herd (figures 7.3, 7.4). It clearly increased inside the eight exclosures following their construction in 1957 and 1962. It recovered abundance over the higher elevation ranges as a whole between the early 1960s and 1990 (figures 7.3, 7.4), probably due in part to the lower herd levels of the 1960s and 1970s, in part to the inclination of the herd to concentrate in the northwestern portion of the range and in part to the increasing tendency for major portions of the herd to migrate out of the park in winter. But even with this recovery, the species abundance and per unit area production over 95% of the winter range in 1995 were only about one-half and one-third, respectively, of those parameters in unbrowsed sagebrush. The present trend is not known. The herbaceous vegetation, primarily the perennial grasses, is subject both to the direct pressures of grazing and indirect influence of sagebrush competition. Although it is dormant and covered with snow through most of its period of use, there are indications of significant impacts on its abundance. With woody vegetation making up less than 10% of the elk diet (Singer and Renkin 1995), most of the forage consumed in winter is herbaceous vegetation. Coughenour and Singer (1996) comment that “forage supply does limit the population . . . [and] there is density-dependent competition for food.” The

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ecological implication of competition is reduction of the resource to where it is in short supply for the competitors. I will review evidence in chapter 9 indicating that the elk experience nutritional deprivation in winter. Thus, there is evidence of significant utilization of herbaceous vegetation. The historical accounts and low level to which it had been reduced by 1958–62 (figure 7.8) indicate that if grazing is sufficiently intensive and protracted (in this case, seven decades of high elk population), it can be markedly suppressed. The exclosure measurements were all low and of a similar order of magnitude in 1958 (one year after construction of the first exclosures), in 1962 (when the second group was constructed), and in 1967 (figure 7.8). This was also the low period of sagebrush abundance (figures 7.3, 7.4). There was some limited fluctuation in the herbaceous vegetation between the 1958 and 1967 measurements (figure 7.8) that could have resulted from weather correlations like those calculated by Coughenour et al. (1996), but these remained largely within a limited range compared with that of the entire time series. Both the Parker transects and the chart quadrats show herbaceous vegetation increases inside and outside the exclosures after the 1967 measurements, continuing through 1974 and 1981 in the Parker measurements, and rising in some measure to high values in 1986 with the quadrats (figure 7.8). The Parker increase was roughly 2 times, and the quadrat increase was of a comparable ~2.5 times. This was also the period of some sagebrush recovery (figure 7.4). The initial herbaceous increase inside the exclosures can reasonably be attributed to protection from elk use. The first half of the 1967–81 period outside the exclosures coincided with the low point of the elk reduction and the population recovery that did not reach a census of 10,000 until 1974. Although the herd censuses returned to levels above 10,000 after 1974, the Parker measurements outside the exclosures showed and the quadrat measurements may imply some continued increase in herbaceous vegetation into the early 1980s. Thus, if the 1967–81 herbaceous trend is associated with variations in elk pressure, there is some lag in the process at both ends of the period. Significant herd reduction began in 1962, yet the Parker and quadrat measurements did not show response by 1967, either inside or outside the exclosures. Reardon (1996) recognized this lag and commented, “This lack of an increase in total cover during a 9-year period of low elk numbers is not unexpected. A slow response to exclusion from grazing on sagebrush-grassland sites has been reported in other studies. . . . The differences in the rate of response are influenced by the type and intensity of disturbance.” The fact that the vegetation increase outside the exclosures continued for a few years into the period of resumed high elk numbers also suggests a lag in response. What the mechanisms might be, and whether this lag is similar to the one discussed previously—the apparent lag (or at least observations) in vegetation response to the burgeoning population in the early part of the twentieth century—is not known at this time. The Parker measurements imply herbaceous decline both inside and outside the exclosures after 1981, as do the quadrat measurements after 1986 (figure 7.8). I hypothesize the following mechanisms for these parallel changes.

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

Reardon, as did Norland and Reardon (1996), showed shrub cover, primarily sagebrush, inside the exclosures rising to its highest level by 1990 (figure 7.4) in the high-elevation and Gardiner exclosures. At some point that increase could be expected to exert significant competitive pressure on the herbaceous vegetation, and could explain decline of the latter inside the exclosures from 1981–89 in the Parker transects and from 1986–90 in the Gardiner and Lamar exclosures shown by the chart quadrats. That the herbaceous vegetation increased inside the Blacktail exclosures from 1986–90 (figure 7.8) could be the exception that proves the rule: The Blacktail exclosures were swept by the 1988 fires, which, as shown by Reardon’s measurements, sharply reduced the sagebrush. The 1986–90 increase in herbaceous vegetation inside the Blacktail exclosures could be the result of lower sagebrush competition from reduced sagebrush density. By 1998 at the time of Rens’s (2001) studies, the community might have been approaching similarity with the other exclosures. His studies showed that perennial grasses “covered slightly more . . . of the area outside exclosures than where protected from grazing.” The 1981–89 decline of herbaceous vegetation measurements outside the higher elevation exclosures shown by the Parker transects and the 1986–90 outside decline shown by the chart quadrats (figure 7.8) could be the result of sagebrush regrowth (figure 7.4) and increased competition, increased grazing pressure imposed by a recovered elk population, or both. That the herbaceous vegetation shown in figure 7.8 declined sharply outside the Gardiner exclosures from 1981–89 (Parker transects) and from 1986–90 (quadrats), despite the fact that sagebrush had been almost totally eliminated in the area, indicates to me a largely elk causation there. The fact that the herbaceous vegetation declined more sharply from 1986–90 outside the Lamar exclosures than inside (quadrats), despite higher sagebrush density inside, also suggests elk involvement. Thus both forces appear to have been involved, with varying strengths of each between the areas. In sum, the 1958–90 parallel trends in the herbaceous vegetation inside and outside the northern range exclosures could have been produced by changing combinations of elk numbers and associated grazing pressures and changing competition with an increasing sagebrush vegetation, both varying in space and time to produce the parallel trends. In the absence of well-designed experiments to test the mechanisms, this remains a hypothesis. But I prefer it to the scenario of weather causation because the 32-year trend is smooth and secular (figure 7.8), paralleling changes in elk numbers and without the abrupt, random yearto-year variations characteristic of weather-driven trajectories. Whether the herbaceous vegetation outside the upper exclosures will decline further cannot, of course, be predicted. But figure 7.8 suggests that it lost one-third to one-half of the recovery experienced before the downward trend. The northern range has now experienced only about 30 years of large, recovered elk herd compared with the 70–75 years preceding the reduction. And the pressure on the vegetation may not have been as stringent since herd recovery as it was before the reduction. This is suggested by the appearance of continued

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sagebrush increase (figure 7.8), although Wambolt believes the species is declining. It is suggested by the appearance at casual summer observation that there is a profuse herbaceous vegetation at the higher elevations with only slight differences between the interiors and exteriors of the exclosures, as Coughenour et al. (1996) point out. Yet there are symptoms of disturbance. Sagebrush emerges as the dominant competitor. That nonnative species are common and apparently increasing outside the exclosures but not significantly inside suggests one or both of two alternatives. First, a robust, relatively undisturbed stand of sagebrush may be required to “close” the community and largely thwart invasion by exotics. This presumes that there are no significant differences between the inside and outside herbaceous vegetation, and the latter, though relatively intact, plays little if any role in competitively excluding the nonnatives. Alternatively, the disturbance of the native herbaceous vegetation outside the exclosures, though too subtle and variable to detect with the measurements made to date, may open the community to invasion. The undisturbed herbaceous vegetation may play a role in resisting exotic invasion inside the exclosures. Chapters 11 and 12 will present evidence that surface erosion has accelerated since park establishment, probably an indication of herbaceous vegetation disturbance.

Other Fauna Unlike the elk effect on aspen woodland, which is the complete elimination of one subsystem and conversion to another (shrub steppe), the sagebrush steppe is being changed structurally by the reduction of sagebrush. The result is to reduce habitat structural diversity. As Chapin et al. (1997) comment, “by definition a change in species composition or abundance must affect ecosystem functioning.” Park publications have largely disclaimed any elk effects on sagebrush abundance over most of the northern range (Singer and Renkin 1995; Bishop et al. 1997; Finley 1997; Singer et al. 2003), and little or no park research has been directed at studying the ecological effects of sagebrush reduction. But there is an extensive literature on sagebrush-steppe ecology, and brief reference to this provides a basis for what the likely effects are on the northern range. Rotenberry and Wiens (1980) and Wiens and Rotenberry (1981) have pointed out that the avifauna of the sagebrush steppe is a combination of species occurring largely in the sagebrush type, and more cosmopolitan steppe or grassland species with ranges extending out into the central North American grasslands. They found three species in particular—Brewer’s sparrow (Spizella breweri), sage sparrow (Amphispiza belli), sage thrasher (Oreoscoptes montanus)— that Welch (2004) calls “near obligate” sagebrush species, whose densities were a function of both sagebrush patchiness and vertical habitat structure. The two sparrows nest in the shrubs while the thrasher nests in or under them. Welch (2004) cites several studies observing reduction or complete elimination of these species following herbicidal removal of sagebrush.

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

The extreme elk effect on the northern range avifauna was evident on July 13, 2001, with casual observation outside the west Gardiner exclosure. Inside there was considerable activity by sage thrashers, vesper sparrows (Pooecetes gramineus), and a female gray partridge (Perdix perdix) with brood of young. Outside there was no avian presence (figure 7.5). Welch also lists sage grouse (Centrocercus urophasianus) as an obligate sagebrush inhabitant. Sagebrush foliage is a major component of sage grouse diet year-round. He generalized that optimum canopy cover for nesting and brood rearing is 20–50%, with areas having less than 15% cover rarely occupied. Wambolt and Sherwood (1999) measured 24% cover inside the higher elevation exclosures, 11% outside (table 7.1). Bailey (1930) commented, “A few sage grouse . . . have been reported in the lower part of the Yellowstone and Lamar River valleys.” But 10 years later, Murie (1940) remarked that “sage hens, at one time present in limited numbers, are now gone.” A similar range of sagebrush relationships appears to prevail with the mammalian fauna. Welch (2004) characterizes the interaction of pygmy rabbits (Brachylagus idahoensis) and sagebrush voles (Lagurus curtatus) as “obligate-like,” both species using a large amount of sagebrush foliage in their diet and requiring canopy cover of 16–33% and 51–55%, respectively. Yet the total number of small mammalian species in a limited sagebrush steppe area may total 8–10 characteristic of the type but not requiring sagebrush. Welch calls these “facultative” species of the type. Parmenter and MacMahon (1983) removed the sage shrubs from a 1.25-ha area in southwestern Wyoming and observed no changes in the populations of four rodent species characteristic of the type. But least chipmunks (Eutamias minimus) declined, and montane voles (Microtus montanus) increased. As discussed at the beginning of this chapter, sagebrush is an important source of food for upland browsing ungulates wherever both occur in the West. In Yellowstone this includes, besides elk, pronghorn and mule deer. Wambolt and Sherwood’s (1999) data, which show a reduction in sagebrush production per unit area over the northern range somewhere between two-thirds at the higher-elevation exclosures and total elimination in the BLA, clearly imply reduction of much of the sagebrush forage for these sympatric species. I defer analysis of the effects on these species until chapter 9. The picture on arthropods is less clear. West (1999) cites on environmental impact statement that reports over 1,000 species of insects in a sagebrushsteppe area in southern Idaho. Wiens et al. (1991) recorded 168 species of arthropods on sagebrush alone. In total, the alteration of the northern range sagebrush steppe, resulting from significant reduction in the dominant plant species, is almost certainly having far-reaching ecological effects on the sagebrush-steppe subsystem and in turn the entire northern range ecosystem. Later chapters will consider likely effects on the hydrology.

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Influences on Upland System Structure III: Conifers and Deciduous Shrubs

8 Objects . . . may be concealed from our view because there is no intention to perceive them. —Henry David Thoreau

CONIFER TYPE Introduction Houston (1982:85) reported the proportion of Yellowstone National Park in coniferous vegetation at 79%. Most of this is on the high-elevation (e.g., >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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• 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.

Influences on Upland System Structure III: Conifers and Deciduous Shrubs

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.

Influences on Upland System Structure III: Conifers and Deciduous Shrubs

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

Influences on Upland System Structure III: Conifers and Deciduous Shrubs

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

Influences on Upland System Structure III: Conifers and Deciduous Shrubs

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.

Influences on Upland System Structure III: Conifers and Deciduous Shrubs

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

Influences on Upland System Structure III: Conifers and Deciduous Shrubs

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-

Influences on Upland System Structure III: Conifers and Deciduous Shrubs

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 <10%, indicating that they were

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“severely undernourished, and that undernutrition was a major contributing factor in their deaths.” The authors also analyzed, and found “high” urinary urea N:creatinine (UN:C) ratios, which indicated “accelerated net protein catabolism . . . a useful indicator of severe or prolonged undernutrition.” In the previous winter, DelGiudice et al. (1991b) had collected 688 urine deposits on snow in the vicinity of 25 predominantly cow-calf herds using the northern range over the period January 13–March 29, 1988. They analyzed the urinary UN:C and urinary potassium:creatinine (K:C) ratios of the samples. They observed a decline in K:C from early January to late March and an increase in UN:C from February to late March. The authors concluded that these trends indicated “progressive nutritional deprivation . . . progressive energy restriction and accelerated net catabolism of protein” through the winter. The effects were more pronounced in the middle and lower northern range, toward which much of the population had moved during the winter, than in the upper (eastern) portion of the range. The winter was considered mild. The overwinter decline in calves/100 cows ratio “suggested that calves were more vulnerable to nutritional deprivation than cows.” In a lengthy synthesis paper, DelGiudice et al. (2001) reported results of snow-urine measurements and dietary analyses of both elk and bison, taken periodically through the winters of 1987–88, 1988–89, and 1989–90. These were, respectively, mild, severe, and moderate winters. They sampled across the northern range, and at two higher-elevation areas in the park. Their results showed decline in nutritional plane through each winter. The degree of decline was a function of winter severity and elevation occupied by the animals. In both cases, snow depth was the key variable. Majority of the diet was grasses and sedges, but conifer foliage was as much as 40% of the elk diet on the northern range in the severe 1988–89 winter. These results imply progressive diminution of the animals’ forage in winter in the portions of the northern range that they occupy. The density-dependent demographic responses—for example, overwinter calf mortality and calf recruitment rates (Coughenour and Singer 1996)—indicate that the amount of forage per animal is greater at low densities than at high. But the relationships are linear, indicating that some nutritional deprivation occurs at all but perhaps the lowest densities. Coughenour and Singer’s (1996:figure 5) analysis implied some calf mortality at all censused population levels above perhaps 6,000 and suppression of calf recruitment at numbers down to about 5,000. Thus there is compelling evidence of intraspecific competition among elk for the northern range winter forage occasioned by usurpation of the resource to the point of shortage for the animals. The result is varying levels of winter mortality, the magnitude depending both on population size and winter severity. The question then follows as to whether depletion of the resource by elk creates shortages for other ungulates using it. That question in turn depends on the degree of spatial and forage overlap of the other species and the timing of their use. And it is competition if there is an effect on the other species’ nutrition and demographic performance.

Influences on Upland System Structure IV: The Ungulate Guild

INTERSPECIFIC COMPETITION Sympatric Grazing Species Bighorn Sheep RESOURCE EVIDENCE. Park investigators have expressed concerns for elk impacts on bighorn sheep since the 1920s. In discussing what he considered to be overgrazing of the northern range by elk in the 1920s, Rush (1932:79) commented that bighorn sheep “Numbers are small and decreasing each year.” On the basis of his observations in the park during the 1930s, Murie (1940:115) commented:

In winter, the bighorn are in direct competition with the elk for practically all food plants. . . . The competition for food each winter is severe. On Mount Everts the bighorn during the last half of the winter subsist on a range so heavily utilized that elk for the most part avoid it, after taking the “cream” of the forage. Several investigators have analyzed forage and habitat overlap between elk and bighorns in efforts to address the competition question. Barmore (1980) based his conclusions on his own research conducted in the 1960s that preceded and continued into the herd-reduction period. He presented evidence on close similarity in dietary preference. Although elk ranged widely over more varied terrain and habitat than sheep, the elk range encompassed most of the sheep range. He concluded (p. 420) that these similarities, plus sheep population response, to be discussed next: all indicate that competition for food occurred between the two species under certain conditions even when the elk population was low, but differences in habitat selection reduced the intensity of competition and permitted coexistence regardless of elk population size. Houston (1982), whose work in the 1970s followed Barmore’s, addressed some of the same questions. Houston examined dietary overlap, facial morphology, chest height, foot load, and habitat preference of the northern range ungulates, all for the purpose of assessing the potential of significant overlap in plant species use, ability to forage in snow, and habitat use. He generally discounted the likelihood of interspecific competition between elk and bighorn on the grounds of habitat partitioning, while also discounting the population evidence to be discussed. He thus differed with his predecessor. Singer and Norland (1994) subsequently studied resource and niche overlap. These authors analyzed dietary composition and observed terrain and vegetation use by five ungulates in the northern range during 1986–88, by which time the elk herd had returned to high numbers following the 1960s reductions. Their purpose was to compare their measurements with similar ones reported by Barmore (1980) for 1968–70 when the elk herd was at its low point. Citing

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competition theory posed by several authors, they hypothesized four criteria of intra- or interspecific competition. The authors calculated several indices of preference and of niche and dietary breadths for the five species. They also posed population hypotheses that I will discuss later in this chapter. Both elk and bighorn sheep increased the areas of winter range occupied between the early and later periods. The evidence showed increases in both vegetation and diet niche overlaps between bighorn and elk, and the authors concluded: “interspecific competition (probably with elk) more likely explains their [bighorn] increased niche occupation. We observed high diet overlap . . . and high habitat use . . . also indicated overlaps between bighorn sheep and elk, suggesting potential competition.” Bison, with some dietary overlap with bighorns, also extended their range between the two periods and thus were potential competitors of sheep. But Singer and Norland generally discounted the possibility because “different mechanisms are inferred from their life histories.” Moreover, bison numbers ranged between 600 and 700 (their figure 2) during the 1986–88 period compared to 16,000– 19,000 elk. The authors concluded that the strongest probability of competition on sheep was with elk. POPULATION EVIDENCE. In discussing the difficulty of demonstrating the reality of interspecific competition, Barmore (1980:603–6) commented that one persuasive form of evidence is a population response in one putative competitor following a population change in another. The elk population reduction of the 1960s provided an opportunity to determine whether the northern range bighorn population would respond. I fully agree, and there is a 74-year data set on populations of the two species on the northern range. But analysis must be somewhat piecemeal because the bighorn census methods changed twice over the period, different investigators disagreed over interpretation of the data, in one case 2 sources published different numbers for the same years, and the population was not censused every year. Thus, bighorns were censused with ground counts from 1924 to 1950 (figure 9.1). Following 5 years without census, sheep counts were resumed in 1955 by air and continued through 1978, then returned to ground counts through 1996. Aerial counts were resumed in 1992 and 1994–96 in addition to the ground counts of those years (figure 9.1). Consequently, it is not possible to reconstruct a clean and precise time series for northern range bighorns over the 1924–96 period, and some interpolations between census methods and gaps are necessary. In addition, bighorn numbers between 1968–69 and 1985–86 in Singer and Norland (1994:figure 2) do not agree with those of Barmore (1980) and Bishop et al. (1997) for the same years. Because the numbers of the latter two sources agree for all years of overlap except 1970, I have used them. Given these limitations, I review the inferences drawn by a sequence of investigators and close with some of my own for the entire 74-year series.

Influences on Upland System Structure IV: The Ungulate Guild

Figure 9.1 Northern range bighorn sheep and elk censuses, 1922–23 to 1995–96. Elk numbers are from tables 2.1 and 2.2. Bighorn sheep numbers for 1923–24 to 1969–70 are from Barmore (1980:612–19), for 1970–71 to 1995–96 are from Bishop et al. (1997:130–31). Three years with faulty elk censuses (table 2.1) are not included in the elk trajectory.

Barmore (1980) observed fall, winter, and spring lamb:ewe ratios for the northern range from 1966–70 when elk numbers were low and compared these with ratios reported by earlier park investigators from 1937–50. Winter-spring lamb:ewe ratios in 1967–70 for the entire winter range were higher (though not significantly) and less variable than those of 1938–50. Spring-fall ratios in 1967– 70, for the Mount Everts subpopulation, which overlapped more with elk during winter use, were significantly higher than those of 1937–45. He surmised that this probably reflected “higher lamb production and/or survival to early winter rather than overwinter survival” (pp. 103–4). He concluded (p. 108) that rising censuses and increasing lamb production and/or survival to spring in the Mount Everts band “all suggest consistent population increase between 1967 and 1970 that may have begun earlier in the 1960s.” He also cited Oldemeyer et al. (1971) who reported mean instantaneous rates of population increase (r) of 0.016 and 0.055 for 1955–65 and 1965–68, respectively, and his own calculations of 0.193 and 0.201 for 1967–70 “based on maximum winter-long and late-winter counts, respectively.” Barmore commented (p. 109) that “This apparent population increase was associated with major reduction of the northern Yellowstone elk herd. Consideration of the niches of the two species and interspecies interactions suggests a cause and effect relationship would be expected and did occur.” Keating (1982:25–28) plotted bighorn censuses on the Mount Everts range for the period 1963–81, regressed an exponential curve to the points, and

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obtained an R = 0.96. Though the exponential fit was strong, he inferred the “Numbers appear to have stabilized at approximately 200 since 1974.” Keating (pp. 35–36) also calculated annual rates of change for the interval t–1 to t of the Mount Everts sheep population and regressed these on northern range elk censuses for t–4. He obtained a strong (R = –0.81) negative exponential correlation and concluded: “ [northern range] elk numbers above 3,000 will reduce the rate of population increase in bighorns on the EWR [Mount Everts range] and . . . EWR bighorns will stabilize or decline in numbers when elk numbers on the . . . [northern range] exceed 9,500.” Houston (1982) disagreed with the inferences Barmore (1980), Oldemeyer et al. (1971), and Keating (1982) drew from the evidence. He analyzed the possible effects of ground temperatures on bighorn census results, concluding that temperatures are a serious variable on sheep visibility and census results. He suggested that increasing census proficiency might explain the apparent population increase from the 1960s to 1978. On the basis of these two points, he disagreed that the evidence showed population increase, and concluded “The status of bighorn sheep on the northern range since the 1920s has probably not been as precarious as sometimes suggested.” Keating (1982, 1985) examined the temperature effect at greater depth, noting that Houston assumed a linear relationship between temperature and sheep visibility. His analysis showed some effect at temperatures below 0°C, but none above. Because the censuses are typically conducted in spring at temperatures above 0°C, he concluded that the temperatures had not significantly affected the census results and that they correctly reflected the population trend. More broadly, he concluded (p. 28): “The apparent trends observed here generally agree with the conclusions of Barmore (1980) but directly contradict those of Houston (1980) [sic].” Houston (1982:181) also commented on Keating’s regression of sheep rt–1 to t on elk Nt–4. “The biological relationships behind the four-year lag are, however, unclear to me.” Because it is usually appropriate in such r/N tests to regress the r-value on the state of the independent variable at the beginning of the time interval over which the population changes (in this case t–1 to t), the usual test would be Nt–1. Hence the time lag implied in Keating’s test is actually only 3 years suggesting that the vegetation (which must be the direct factor to which the sheep respond) requires up to 3 years to respond to changes in elk numbers, either positively or negatively. We have seen such northern range response lags previously in chapters 6 and 7. Keating (1982) also observed a number of characteristics of population quality during the 1980–81 period of his study. He observed mean suckling times of lambs, and the 26.9– and 25.1–second periods he recorded compared favorably with similar observations of other investigators studying healthy sheep populations. Lamb suckling times are a measure of the energy reserves of the ewes. In addition, Keating observed age composition of rams to ascertain maturation rates and concluded that the northern range rams during this period were reaching sexual maturity 1 year earlier than less healthy populations reported

Influences on Upland System Structure IV: The Ungulate Guild

by other authors. He also measured mean levels of lungworm infections by counting the number of lungworm larvae per gram of sheep fecal material. A log-normal mean of 16 larvae per gram was far below the incidence even in populations with light infection rates elsewhere in North America. In total, all of these criteria pointed to a healthy population during these years. With 14 more years of bighorn and elk censuses, it is now possible to evaluate the population evidence for interspecific-competition over 74 years of park history (figure 9.1). The 21 bighorn ground counts between 1924 and 1950 varied largely between 100 and 200, and averaged 151. Although there is some suggestion of a decline during the drought of the 1930s, there does not appear to be a net trend over the 26-year period, suggesting rough equilibrium during a period when the elk population averaged 9,834 censused animals. It is perhaps no coincidence that Keating concluded nearly a half century later that the sheep population equilibrates at a censused elk population of ~ 9,500. With the first four sheep aerial censuses (1955–62) in the same range of numbers as the preceding 21 ground counts (x\ = 144), it might be tempting to conclude that the accuracy is comparable for the two methods. But the four ground counts conducted in the same years as aerial counts in the 1990s (figure 9.1) averaged only 0.63 of the latter. Thus comparison of the 1965–78 aerial counts with the preceding ground counts must assume that the former are somewhere between equivalent to and 1.6 times the latter. Whatever the differential, I infer that the bighorn population increase implied by the 1965–78 aerial counts (figure 9.1) was real, and that it began during the northern herd elk reduction. The mean elk census for 1965–78 was 7,824, and for 1962–75 it was 7,344. Five of the 12 effective censuses during 1965–78 were below 6,000. The last five aerial sheep censuses in the 1970s averaged 435. If these are taken as comparable with 151, the mean of the 2 ground censuses of the 1920s– 50s, they imply an almost threefold increase in the population. If the ground counts are considered to average 0.63 of the aerial counts, the increase was approximately 1.8 times. The actual increase was probably somewhere between these two factors. The evidence supporting reality of the increase is Keating’s (1985) reanalysis of the temperature effect, Barmore’s (1980) observations of higher lamb:ewe ratios during the increase, the higher mean r-value during the increase calculated by Oldemeyer et al. (1971), and Keating’s (1982) inverse correlation between sheep r-values and elk numbers. The Mount Everts sheep population dropped precipitously between 1980 and 1981 due to a disease outbreak, with Chlamydia identified as the causative organism. The disease (pinkeye) blinded the affected animals. At the same time, sheep censuses reverted to ground counts, which showed the entire northern range population reduced to well below 100 (figure 9.1) even though the Mount Everts animals were the only subpopulation affected (Meagher 1982). Keating (1982:54–55) suggested that the bighorn population reduction could serve as a second quasi-experiment (after the elk reduction) to test the reality of elk competitive suppression on sheep. If, following recovery from the epidemic, the sheep population returned to the levels of the 1970s, it would

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suggest that interspecific competition had not been a significant constraint on sheep numbers during the 1920s–50s period. This would be especially true with the elk population now risen to a range of 13,000–19,000 censused animals (figure 9.1). If the sheep population did not return to preoutbreak levels, it would be circumstantial evidence of elk competition. The sheep population does appear to have increased somewhat between 1983–90 (figure 9.1). But it then appears to have dropped back to an average ground-censused level of 105 in the 1990s. Thus the population now appears to be averaging about two-thirds the level of the 1920s–50s, and a wide range of evidence points to elk competition. The question then arises as to the continued viability of the population. Another Chlamydia outbreak could eliminate it. The decline from the last epidemic killed substantially more animals—whether from a 1970s average population of 434 or ground-censused average of 274—than the current, entire population with a censused 105 average. Several authors have analyzed survival probability of bighorn populations driven to low levels. Berger (1990) concluded that populations reduced below 50 animals “are unable to resist rapid extinction.” Singer et al. (2001b) concluded that the major variable affecting survival is habitat patch size but also concluded that a population reduced to 50 or fewer animals had only a 5% probability of surviving an epizootic. The addition of wolves to the ecosystem poses a new, apparent competition pressure on the northern range population. I will return to this subject later. HOW MUCH DECLINE SINCE PARK ESTABLISHMENT? Barmore (1980), Houston (1982), Kay (1990), and Schullery and Whittlesey (1992) all concur that bighorns were more numerous and widely distributed in the area that is now YNP than was the case through the 1900s. Kay (1990:253–55) quotes the journals of trapper Osborne Russell, who traveled in the Yellowstone area in the 1830s. At one point Russell reported seeing “thousands” of sheep in the locale, and his journals reported seeing sheep as often as he saw elk. Other historical sources consulted by Kay suggest that sheep were more than a third as numerous as elk, and archaeological evidence suggests sheep in greater numbers than elk. Five years after park establishment, Superintendent Norris (1877) reported that “over 2,000 hides of the huge Rocky Mountain elk, nearly as many each of the big horn deer and antelope . . . were taken out of the Park in spring of 1875.” Barmore (1980) cites later superintendents’ reports as saying “sheep were abundant” (1890) and “Found in all of the mountain ranges” (1887). But by 1897, Superintendent Anderson (1897) estimated park sheep numbers at 200. Keating (1982) emphasizes the role of poaching and introduction of livestock associated with settlement of the Yellowstone area and reduction of the sheep population. By 1897 the elk herd reached the high numbers of the turn of the century. Some measure of the extent of decline between park establishment and the present can be gained by calculating the prehistoric relative proportions of the

Influences on Upland System Structure IV: The Ungulate Guild

ungulate species indicated by archaeological and historical data and applying these to the Keigley and Wagner (1998) estimate of elk numbers at park establishment (chapter 3). These proportions, expressed as ungulates/100 elk ratios, are shown in table 9.1. As Kay (1990) has pointed out, all of these sources indicate that elk were a substantially lower and the other ungulate species substantially higher proportion of the ungulate guild than is the case today. In fact the archaeological evidence suggests that elk in the Yellowstone area were the least numerous of the ungulates except for moose (table 9.1). That same evidence suggests that bighorn sheep were the most numerous. The historical evidence, while also indicating that elk were a smaller fraction of the ungulate guild than today’s numbers, suggest that they were the most numerous species (table 9.1). I will use these as the more conservative values. Kay’s data indicate the number of sheep per 100 elk to have been on the order of 38. Schullery and Whittlesey place this number at 19. If we apply these to the Keigley and Wagner (see chapter 3) approximation of 5,000 elk in the northern herd at park establishment, the respective estimates of northern range sheep numbers are 1,900 and 950 at the same date. If we set the mean, northern range population in the 1990s at 154, as discussed, the indicated decline since park establishment is somewhere between 92% and 84%, respectively. Obviously there has been extreme reduction of the population. If Houston’s (1982) surmise that the prehistoric northern herd elk population was 12,000–15,000 were correct, the estimated sheep population would have been 2 to 3 times the above numbers, and the percentage decline to the present would be more drastic than 84–92%.

Bison As obligate grazers, bison are the third species of the northern range ungulate guild that uses a finite herbaceous vegetation and thus are potential competitors with the other 2 grazing species. Bison have had less research and public visibility until the 1990s; the recent attention has been directed more to their potential for moving out of the park and transmitting the disease brucellosis to domestic cattle (Cheville et al. 1998) than to their ecological role in the northern range ecosystem. The lack of attention may in part be due to their small numbers in the park during much of this century as a result of poaching and herd reductions. From 1967 to the present, the natural-regulation policy has provided bison the same protection from culling inside the park that it has provided elk. The last park bison cull occurred in 1965. Starting from lower numbers in the 1960s than did the elk, and with a lower rate of population increase, it has taken bison a longer period to reach sufficient numbers to become a significant presence. From 1985 to the present, there have been periodic herdreduction efforts plus hunting kills outside the park. Recent research attention has focused largely on movements and populations. There are now two models of bison population behavior in the park. From my analyses, the evidence does not support one of the models and supports the

INTRODUCTION.

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152

Number per 100 Elk (No. Observations) Date Source

Elk

Deera

Pronghorn

Bighorn

Bison

Moose

Reference

Archaeologicalb Historicalc Historicald

100 (17) 100 (42) 100 (131)

541 (92) 71 (30) 54 (71)

65 (11) 71 (30) 50 (66)

853 (145) 38 (16) 19 (25)

282 (48) 7 (3) 24 (31)

0 (0) 0 (0) 5 (6)

Kay (1990:297) Kay (1990:273) Schullery and Whittlesey (1992)

aBoth

mule deer and white-tails. are the archaeologists’ minimum number of individuals in four sites excavated in the Greater Yellowstone Ecosystem. cThe number of occasions on which each species was reported in the journals of the 20 parties traveling through the Yellowstone area, 1835–76. See Kay (1990) for details. dThe number of occasions on which each species was reported by 168 parties traveling through the Yellowstone area, 1806–81. See Schullery and Whittlesey (1992) for details. bNumbers

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152 Table 9.1 Numbers of Ungulates per 100 Elk, and Number of Observations in the Yellowstone Area in Archaeological Sites and Historical Accounts

Influences on Upland System Structure IV: The Ungulate Guild

other only partially. Hence I wish to provide my own population inferences and then discuss the species’ ecological role in the northern range ecosystem. POPULATION BEHAVIOR. The YNP bison population functions much like a meta-population. There are three subpopulations: the Lamar Valley or northern range group, the Pelican Valley animals of the eastern portion of the park, and the Mary Mountain population, including animals both of the Hayden Valley and Firehole River valley in the central and western region of the park. Because the northern range herd only constituted 18% of the total park population from 1990–93, and there is considerable exchange between the subpopulations (Meagher 1993), it is useful to analyze the behavior of the entire park population rather than the northern herd independently. I focus more on the northern herd shortly. Bison were present in prehistory in the Yellowstone area (Kay 1990; Schullery and Whittlesey 1992). After park establishment, poaching drove them almost to extinction by 1902 when the park population was reduced to an estimated 40–50 animals (Meagher 1973). Additional animals were brought into the park to bolster the population and kept in enclosures first at Mammoth and later at what is now the Lamar Ranger Station (Meagher 1973). The captive animals increased, gradually escaped to mix with the surviving natives, and together grew to reestablish the park population. By the 1930s, YNP numbers had grown to more than 1,000, and park officials decided that the herd needed to be reduced to protect range conditions. Periodic removals continued through 1965, and the 1966 census counted 226 animals (Meagher 1973). Subsequent population changes began from that point. I have plotted the parkwide early winter censuses for 1966–97, and the northern range counts for 1972–93, all in figure 9.2. To gain some insight into how the annual rates of change have varied with time, I have plotted the parkwide censuses semi-logarithmically in figure 9.3. To determine how annual rates of change have varied with population size, I regressed annual r-values (lnNt+1– lnNt), calculated from the censuses, on the censuses of year t (figure 9.4). I interpret these results as follows. The population grew steadily from its culling-induced low point of 1966, as has the northern-range herd (figure 9.2). Periodic removals with culls and hunting kills were resumed in 1984–85, with significant numbers as follows (Meagher 1993; Peter Gogan, personal communication, May 19, 1999):

1984–85 1985–86 1987–88 1988–89 1991–92 1992–93 1994–95 1995–96 1996–97

88 57 35 569 271 75 424 426 1,084

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Figure 9.2 Total park bison censuses (solid circles) for the winters of 1965–66 through 1996– 97, and northern range censuses (half-open circles) for the winters of 1971–72 through 1992–93. Regression line is fit to the 1971–72 through 1996–97 park censuses. Numbers below regression line are major removals. Open circles represent hypothetical population trend without removals (see text for rationale). Census data for 1965–66 through 1967–68 from Meagher (1973), remaining years from Bishop et al. (1997).

Although the removals of 1988–89 and 1991–92 reduced the populations of the following years, the lack of or small removals of the intervening years allowed continued increase up to a high point of 3,956 in 1994–95. Only the large removals of that and the two following winters were sufficient to stop further increase, at least up to 1996–97. The curvilinear, semi-logarithmic plot of the censuses (figure 9.3) implies that the relative growth rates (r) declined steadily through the 1972–97 period, on average. The inverse correlation between the annual r-values as a function of censused population size (figure 9.4) implies linear density-dependent pressure on the reproductive and/or mortality rates as the population increases. Intercept of the regression line with the r = 0 line at 3,300 censused bison suggests from this test that the population would, on average, equilibrate at this size. I will suggest with further analyses that this estimate is conservative and the actual equilibrium value would probably be substantially greater were it not for the animals’ movement patterns.

Influences on Upland System Structure IV: The Ungulate Guild

Figure 9.3 Semi-logarithmic plot of total park censuses for the winters 1971–72 through 1996–97. Curved line implies declining annual relative growth rates (r). Census data from Bishop et al. (1997).

Meagher (1993) presented both early and midwinter censuses for 1972– 93. I calculated the percentage change from the first to the second in each winter and regressed the percentages on their respective early-winter values. The inverse relationship was highly significant at p < 0.01 with an R2 = 0.53, suggesting that density-dependent winter mortality is a partial cause of the densitydependent decline in r as the population increases (figure 9.4). The relationship in figure 9.4 is noisy, probably for several reasons: There undoubtedly is some error in the censuses. Severity of winter weather induces some random variation in winter mortality (Meagher 1971). And the periodic large removals add further variation to the time series. The removals also cause underestimation of r-values if they are based only on successive pairs of annual censuses, as in figure 9.4. If an early winter census in year t is followed by a significant removal, the postremoval population size is smaller and a better index of the breeding population size at the start of year t + 1 than the preremoval census. To use the latter in calculating r underestimates the increase rate by the following winter. To avoid this problem and calculate more realistic r-values, I derived the Nt values in removal years by subtracting the removals from the early winter censuses of the same winters. The r-values were then obtained by calculating the difference between these Nt values and the censuses of the following winter.

Figure 9.4 Regression of instantaneous rates of change (r) between successive pairs of total park early winter bison censuses in t + 1 and t on census of year t for the period 1971–72 through 1996–97. Census data from Bishop et al. (1997).

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Regressing these on the censuses at time t produced an inverse relationship similar to that of figure 9.4 with an R2 = 0.11, p ~0.10, and an implied equilibrium population (r = 0) of ~6,000. From these analyses, I infer the following pattern of population behavior by the park bison. When culling stopped in the latter 1960s, there were no population constraints sufficient to keep it from increasing through the next three decades (figure 9.2). As it increased, density-dependent constraints operating in conjunction with randomly varying winter loss progressively reduced annual rates of change (figures 9.3, 9.4), either through reduced natality and/or increased mortality. Without human intrusion, the population would equilibrate at ~ 6,000 except that, as Cheville et al. (1998) point out, when the population exceeds 3,000, the animals begin moving out of the park in winter. As part of the park meta-population, the northern range bison subpopulation has behaved essentially the same as the total. The northern herd began its increase in the early 1970s at approximately 100 animals and rose to a mean census of 549 in the early 1990s (Meagher 1993). The total park population increased from an early winter census of 565 in 1971–72 to a mean of 3,129 in the early 1990s, a 5.5(increase identical to that of the northern range subpopulation. Although the park herd’s equilibrium level without the winter exodus appears to be ~6,000, it has never reached it. As commented, Cheville et al. (1998) conclude that the herd begins moving out of the park when it reaches around 3,000 animals. They are then subject to sport hunting and artificial reduction. Thus they are never allowed to reach their equilibrium level set by their own density-dependent constraints. Except for the winter exodus and intercession of control efforts that prevent attainment of an equilibrium point, the bison pattern is similar in principle to the elk herd. With cessation of park control in the 1960s, the populations increased steadily, albeit with stochastic variation associated with varying winter severity. On average, both populations increased with declining rates of change until the elk herd approximately equilibrated (see figure 2.1), but the bison herd has been prevented from doing so by its movements and periodic removals. The density-dependent suppression of the elk r-values has clearly resulted from intraspecific competition for food, which probably has also occurred in bison, as will be discussed. This model of the park bison population departs almost completely from the prevailing view of park biologists. That view has held that the herd was “regulated” at roughly 1,400–2,000 (Meagher 1993:30) or at an “upper equilibrium” of approximately 2,000 (Meagher 1985; Schullery 1986) until the latter 1970s. In fact, the population reached about 2,000 animals in 1980–81 (Meagher 1993:42), but it was still clearly increasing (see figures 9.2, 9.3). At 2,000 animals, the mean r-value was ~0.08, or an annual percentage increase of 8%, as indicated by the census-based regression (figure 9.4). The regression based on removal-corrected r-values, already discussed, places r for 2,000 animals at ~0.11, or 12% per year increase. The population was clearly not close to equilibrium.

Influences on Upland System Structure IV: The Ungulate Guild

The park model further opines that beginning in the latter 1970s, efforts were expanded to snowplow and groom park roads to facilitate winter snowmobile travel by tourists. This provided travel routes for bison seeking more forage. And the ease of travel, by comparison with movement through deep snow, purportedly enabled the animals to expend less energy, which was then converted to higher survival rates, particularly among calves (Schullery 1989; Meagher 1993). The result has been variously termed “escalating . . . population increase” and “classical eruption” resulting in the 3,000–4,000 populations of the 1990s. This interpretation was quickly picked up by the media (Satchell 1996; Anonymous 1997a, b; McMillion 1997; Peacock 1997) and gained widespread acceptance. I not only agree with the conclusions of the recent National Research Council (NRC) study (Cheville et al. 1998) that there is no evidence supporting this roadgrooming paradigm; I conclude that the available evidence indicates there has not been such an effect. No bioenergetic studies have been conducted to support the park’s claim that the animals’ road use expends significantly less energy than movement across the snow-covered landscape. All movement requires energy, and bioenergetic studies have long shown that the majority of an endotherm’s energy use is in metabolism. There were no counts of the numbers of animals using the roads to determine that a significant fraction of the population availed itself of this amenity until the studies of Bjornlie and Garrott (2001). Through two winters of intensive investigation, these authors observed that the animals used the roads to only a minor degree, largely traveling along their own trails, streambeds, and thermal areas. Park publications have commented a number of times that the road use reduced winter mortality, but no population data have been shown to support this contention. As pointed out, the decline from early winter to midwinter censuses actually suggests progressively increasing mortality rates as the population increased through the 1980s and 1990s. The progressive decline in r-values during the same period (figure 9.4) implies increasing mortality and/or decreasing natality. Finally, there is no evidence of population equilibration prior to road grooming and then an upsurge in population growth. Rather, the population was growing at its highest rates in the 1970s (figure 9.3), and those rates declined steadily during the 1980s and 1990s. The mean census-based r-value from 1971–72 to 1982–83 was 0.11, or 12% increase per year. The mean for 1983–84 to 1996– 97 was 0.03, or 3% increase per year. The population has grown more slowly during the road-grooming period than before. In drawing these conclusions, I do not imply any advocacy for or against road grooming or bison population control. I have devoted this amount of space to the issue only to open for general scrutiny what I believe the evidence shows to be the truth of the situation. It is a sensitive issue, and disputed policy matters can be addressed effectively only when fully illuminated by the light of scientific truth. Finally, in closing this population section, I wish to comment on the NRC panel’s interpretation of the bison population evidence (Cheville et al. 1998). I

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agree with their general conclusion that there is no evidence that road grooming contributed to recent population increase. But I differ slightly with their interpretation of the population evidence, and I believe this difference merits clarification. The NRC authors fit a straight line to a plot of the 1971––72 through 1996– 97 censuses (their figure II-6) as I have done in figure 9.2. The fit was tight, and the slope, or absolute increase in bison numbers per year, was 145. In my own calculations, this proved to be 131, a difference that I will explain shortly, but one of no great significance here. The important point is the linearity of the relationship, and what that does or does not imply about the biology of the situation. The NRC authors’ interpretation of the relationship can appear to be somewhat ambivalent. Thus (p. 64) “there is little evidence of natural diminution of bison population growth except that induced by the severe winter of 1996–97.” Yet in the following sentence, “Although the absolute annual increase is essentially constant at 145 bison, the per capita rate is declining.” The second sentence clarifies the authors’ view, but the inexplicit first sentence can confuse the reader, coming as it does before the second. The authors go on to speculate on the biological reason for why the annual increase should be constant at 145. They suggest that somehow the herd controls this number through some unspecified regulatory mechanism, probably in terms of the number of calves produced because mortality is inferred to be low and roughly constant except for extreme winters. Given the linearity of the relationship and constancy of the slope, the annual increment of 145 must have been permitted each year over the full 1972–95 range of population sizes from 565 to 3,956. Although I don’t suggest that the authors intended this explicitly, it is not too much of a stretch to suggest that this interpretation implies that the herd can somehow count and control each year’s increment at this number. As a further implication, the authors point out that a population that continues to increase at a constant absolute rate of 145 per year without any evidence of density-dependent constraint on that number will continue to increase for some undetermined period into the future before it stops growth. There is no way of predicting what that number will be. For the present, no equilibrium point is in sight. I suggest an alternative, more parsimonious, and less speculative interpretation. As discussed, the park population was subjected to periodic removals in the 1980s and 1990s (Meagher 1993:table 6; Peter Gogan, personal communication, May 10, 1999). I have shown the largest of these in figure 9.2. In each year following these major reductions, the census was lower than the previous year. Thus, the linearity of this relationship and the NRC figure II-6, rather than being an expression of some intrinsic biological behavior of the population, is an artifact of the periodic reductions. And similarly the slope and its coefficient, whether 145 or 131, rather than being some measure of the herd’s biological behavior are simply statistical artifacts resulting from the regression fit to an array of points made linear by the removals.

Influences on Upland System Structure IV: The Ungulate Guild

Cheville et al. (1998) attempted to compensate in part for the removals by adding each to the following year’s census (figure II-6) on the reasonable assumption that had they not been removed, most would be present in the following year’s population and census. This tightened their regression fit and explains the difference between their slope coefficient of 145 and mine of 131 in figure 9.2, which is based only on the uncorrected censuses. However, these one-time additions do not fully compensate for the removals and allow one to approximate the population’s behavior without withdrawals. The 569 animals withdrawn in winter of 1988–89 would be largely present not only in the 1989–90 population but to a substantial degree in subsequent years’ populations, depending on the rate of mortality shrinkage. Moreover, each years’ survivors would be expected to add animals through reproduction, and they in turn to reproduce in subsequent years. In short, the animals removed would constitute a subpopulation that could reasonably be expected to increase over time with the remainder of the population, and at comparable rates. I have partially approximated this in figure 9.2 with the open circles, which are cumulative additions of the removals. This overestimates their addition to the population over time because it makes no provision for their shrinkage due to mortality. But that excess is compensated by there being no provision for what would be their reproductive additions over time. In sum, I contend that the closed circles through 1984–85 and the open circles in subsequent years more accurately (if crudely) approximate the natural trajectory the herd would have taken in the absence of removals than does the artifactual straight-line fit to the entire series of closed circles. It is not too much of a stretch to suggest that the trajectory might be sigmoid, even perhaps approaching logistic. The linearity does appear to represent a trajectory without removals from perhaps 1972–73 to 1985–86. But I have included the three postculling censuses of the 1960s (Meagher 1973:appendix IV), and at this point the trajectory clearly departs from linearity. These points may well reflect the lower tail of a sigmoid curve. Their departure from linearity reflects the obvious biological reality that a bison herd ranging in size from 226 to 418 is biologically incapable of increasing its numbers by 145 or 131 animals per year, given the species’ inherent reproductive rate. If the straight line reasonably fits the trajectory from 1972–73 to 1985–86, it no longer does so with the trend represented by the open circles. The entire pattern from 1965–66 to perhaps 1991–92 might reflect the rising lower half of a sigmoid. There is a suggestion of an inflection at about 1991–92, perhaps reflecting incipient asymptotic behavior of the upper part of a sigmoid. Such behavior is also suggested by the semi-logarithmic natural-log plot of the NRC figure II-5 and my own semi-log plot in figure 9.3. Finally, the linear r/N relationships in figure 9.4 and in my removal-corrected r/N regression are both characteristic of logistic population increase. I conclude from the present evidence that the population trajectory of the YNP bison population, in the absence of removals, might follow a sigmoid and possibly logistic pattern with K at ~ 6,000 animals,

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again hypothetically if the animals did not move out of the park at numbers above 3,000. All of this may seem like quibbling over esoteric details. But the goal of science is surely to describe reality as precisely and correctly as possible. BISON ROLE IN THE NORTHERN RANGE. Until about the mid-1970s, the northern bison herd wintered in a small area of the eastern Lamar River valley, the same area in which it summered. In 1976, the herd began expanding its winter range westward in the northern range (Meagher 1989, 1993). Today it winters over a major portion of the northern range, having increased its distribution threefold over its earlier Lamar occupancy (Singer and Norland 1994) and under some winter conditions moving outside the park boundary (Cheville et al. 1998). Bison and elk wintering distribution on the northern range now overlap substantially, and so do bison and bighorn to a small degree (Singer and Norland 1994). There is some partitioning of terrain, with bison preferring lower, flatter topography than either elk or sheep. Barmore (1980) comments that terrain overlap is greatest in spring. There is also some forage partitioning with bison consuming sedges and course lowland grasses more readily than elk or sheep. But all are grazers and use some graminoid species in common. Park publicity has maintained that the bison are in excellent condition and show no signs of nutritional deprivation (Meagher 1985; Schullery 1986; Bishop et al. 1997). But the evidence for bison is similar to that for elk. DelGiudice et al. (1994) collected 468 samples of bison urine voided on snow in the three wintering areas from January 13 to April 4, 1988. Declines in urinary K:C ratios and increases in UN:C ratios indicated nutritional deprivation and protein catabolism in all three sampling areas. The pattern was more pronounced in the middle upper (eastern half) than the lower (western half) of the northern range. The same pattern prevailed after two more years of study (DelGiudice et al. 2001). Moreover, the decline in r-values (figure 9.4) as the population increased, indicating either declining natality and/or increasing mortality, plus the evidence of rising overwinter mortality inferred from the censuses are consistent with progressive shortage of forage. Although the increasing winter mobility and range extension in the north have been attributed to the road grooming, there is no evidence (except coincidence) of the latter. The probability is greater that the movement has been induced by declining per capita amounts of forage and the need for the animals to seek alternative sources. Thus there is evidence at least of intraspecific competition for forage in the park population, and the DelGiudice et al. (1994, 2001) nutritional evidence suggests this is occurring in the northern range subpopulation. But there is very little additional evidence to allow exploration into the question of bison interspecific competition with the other two grazers. Despite all the years of park ungulate research, there has been no work on bison impacts on vegetation. It is possible to make crude relative comparisons of northern range forage consumption of elk and bison on the basis of numbers and weights. Northern

Influences on Upland System Structure IV: The Ungulate Guild

range elk numbers from 1990–95 averaged 15,094 animals. As already commented, northern range bison numbers in the early 1990s averaged 549. Because the bison both winter and summer on the northern range, it is reasonable to double their numbers for comparing forage consumption. And because their weight is near twice that of elk (Houston 1992:157), the numbers can reasonably be doubled again. Thus bison northern range forage consumption, vis-àvis that of elk, can be considered roughly equivalent to 4 × 549 = 2,396 elk, or for convenience 2,400. This makes no allowance for bison moving from the other two subpopulations to the northern range in winter. Some animals move from the Pelican Creek subpopulation to the Lamar Valley in winter, but there has not yet been any major movement of Mary Mountain animals to the northern range (Meagher 1993; Peter Gogan, personal communication, May 10, 1999). Thus the 2,400–animal estimate may be slightly conservative, but is perhaps a reasonable approximation for comparison with elk forage consumption. Converted to a percentage, the bison consumption may have approximated 2,400/15,094 elk = 16% that of the elk consumption during the early 1990s. Thus the bison consumption further depletes an herbaceous vegetation that is already reduced to the point of significant shortages for both elk and bighorn sheep. Whether or not the separation on the basis of terrain and forage preferences removes bison consumption from competitive pressure on the other two species is not known. Houston (1982:181) plotted bison fraction of young in the herd as a function of the two independent variables elk numbers and bison numbers. Although there appeared to be some influence of both, the samples were small, and he was not inclined to attach much significance to the results because bison and elk numbers were both covarying. It was possible above to gain some correlative evidence of elk competitive pressure on bighorns by observing sheep demographic responses to variations in elk numbers. Because bison and elk have covaried, it is not possible to separate out any possible effects of bison. And because elk numbers and putative forage consumption far exceed those of bison, elk effects can reasonably be assumed to be the major ones affecting sheep. But bison numbers and forage consumption far exceed those of sheep, and some bison effect on sheep is possible, even if small. For the present, the most that can be said is that bison consumption of the northern range herbaceous vegetation adds to the removal by elk and bighorns. Collectively, that removal creates shortages, nutritional deficiencies, and population constraints on all three species. The evidence indicates intraspecific competition in elk and bison, and elk interspecific pressures on bighorns. Although there is no evidence at this time, the possibility remains of some bison interspecific pressure on sheep and mutual interspecific competition between elk and bison. DelGiudice et al. (2001) concluded that the latter occurs. WHAT CHANGES SINCE PARK ESTABLISHMENT? It is possible to approximate, speculatively, the order of magnitude of bison numbers on the northern range

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at the time of park establishment in the same manner as calculated for bighorn sheep. The two historical ratios of bison per 100 elk in table 9.1 are 7 and 24. With an 1872 northern range elk population at ~ 5,000, the implied northern range bison numbers were 350 and 1,200, respectively. Meagher (1973) surmises that the total park prehistoric population was ~ 1,000. If the bison distribution in 1872 in the park areas was roughly similar to the present distribution—and Norris’s (1880) description suggests that this may have been the case—and the northern range subpopulation was approximately the same percentage (18%) of the park numbers as the contemporary one—of course an uncertain assumption—the implied northern range subpopulation was on the order of 0.18 × 1,000 = 180 animals according to Meagher’s information. Thus 1872 northern range numbers may have been somewhere between 180 and 1,200. Consequently the recent, northern range subpopulation—mean of 549 in the early 1990s, and reaching 838 in 1988–89 (Meagher 1993)—may have been on the same order of magnitude as the prehistoric numbers. The current level is significantly maintained by periodic removals, much as the prehistoric numbers may have been maintained by aboriginal hunting (Kay 1990, 1994a).

Sympatric Browsing Species There are four species in the northern range ungulate guild—mule deer, pronghorn, white-tailed deer, and moose—for which woody vegetation constitutes the major fraction of their winter diets. The latter two riparian species are discussed in the next chapter, and hence I will dwell only on mule deer and pronghorn in this chapter on upland types. Woody vegetation constitutes 52–65% of mule deer and 82–83% of pronghorn winter diets on the northern range (Singer and Norland 1994). Although browse constitutes only 11–15% of winter elk diets in the area, the fact that they outnumber mule deer and pronghorn on the northern range by factors of 7 and 100, respectively, and outweigh individuals of the two species by factors of 4–5 (Houston 1982:157) indicates that elk browse consumption far exceeds that of the other two species combined. Hence there is substantial reason to suspect elk competitive pressure on mule deer and pronghorn. This probability is strengthened by the marked reduction of nearly all categories of woody vegetation by elk discussed in the chapters in this part of the book.

Mule Deer POPULATION TRENDS.

Houston (1982) commented that mule deer were the second most numerous ungulates on the northern range. The historical evidence suggests that this was the case at about the time of park establishment (table 9.1), and archaeological evidence suggests that deer may have outnumbered elk and were second only to bighorn sheep in total numbers in the region (table 9.1).

Influences on Upland System Structure IV: The Ungulate Guild

Park censuses began in 1904 (Barmore 1980). But reconstructing the time series from that date to the present is problematic because of periodic long stretches without censuses, change from ground to aerial censuses in 1961, series of years in which both the animals inside and outside the park were counted, some years with counts only of the inside animals, and some with only the totals given. But subject to these difficulties, the data are plotted in figure 9.5, and I offer the following as potential explanations. The early censuses began with 120 animals in 1904 and rose to 1,000 by 1911. Because this increase is roughly twice the reproductive capability of the species, the rise must at least in part be the result of increasing facility with censusing. There is then a 19-year period without censuses, and nothing can be gleaned about the trend over that period. But with resumption of the counts in 1930, and seven conducted in the 1930s and 1940s, the total northern range wintering population varied between counts of 516–1,090 and averaged 738. Subject to the risk of generalizing from such a short series, the population may have been at rough equilibrium during this short period. However, the data suggest that starting in the 1940s the number of animals wintering inside the park began a decline that continued at least into the 1980s. Unfortunately, the available data show only one full northern range census (1968) between 1948–78, and hence it is not possible to say whether

Figure 9.5 Mule deer censused on the northern range. Closed circles represent the total counts, both inside and outside the park. Open circles are the numbers counted inside the park. No inside-park counts since 1983 are available. Data from Bishop et al. (1997), except for 1983 inside-park count from Singer and Renkin (1995).

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the total population declined at the same time the wintering animals in the park were declining. But most of what is now considered the northern range wintering herd (Bishop et al. 1997) winters outside the park at present. The proportion began increasing in the 1930s, rising from 11% in 1938 to 29% and 31% in 1939 and 1940, and reaching 65% in 1940. The percentage dropped to 10% in a single census in 1968, but Houston (1982) commented on the basis of his work in the 1970s that 65% of the herd wintered outside the park. By the latter 1980s, only 5% (on the order of 100 animals in a northern range total of ~2,000) wintered inside the park. By the 1980s the number of deer wintering in the lower elevations of the northern range inside the park had declined by 66% over the previous two decades (Singer and Renkin 1995). The total population censuses varied between 1,616 and 2,544 between 1986 and 1996 (Bishop et al. 1997) and averaged 2,043 during the 1990s. The total reached numbers substantially above those of most of the 1900s (figure 9.5). However it is not clear how comparable these values are. Recent telemetry studies have shown that some of the wintering animals summer in areas outside the park (Thomas Lemke, personal communication, May 20, 1999). The one unequivocal change is the sharp reduction of animals wintering inside the park from ~ 700 to perhaps fewer than 100. One interpretation of these trends is as follows. Mule deer were apparently more abundant at the time of park establishment than during most of the twentieth century. Barmore (1980) comments to this effect, and Schullery and Whittlesey (1992) observe that deer were seen by early travelers more widely over the park than is now the case. The relatively low numbers of the 1930s to 1940s (figure 9.5) occurred at a time when elk numbers had been high for several decades, and both species wintered largely in the park. Russell (1932) remarked “only in years of excessive snowfall . . . do the deer make a considerable effort to leave the protecting limits of the Park.” Deer numbers may have declined further during the 1950s and 1960s. The low deer numbers and decline may have been the result of declining forage conditions largely resulting from elk use. Rush (1932:65) stated: “All browse species are heavily overgrazed and will eventually disappear from the range unless improvement is shown in the next few years.” Murie (1940), referring to his observations in the 1930s, commented that there would probably be more deer if there were fewer elk. Range decline apparently continued for several decades. Singer and Renkin (1995) speculated that the two-decade (1960s to 1980s) decline in deer using the BLA might have resulted from sagebrush overuse, in part by pronghorn. However, I have commented previously on the tendency for park biologists to discount the impact of elk on browse species, including sagebrush, on the basis of dietary composition. Barmore (1980:74) cited demographic evidence that may reflect deer nutritional stress. Generalizing from Murie’s (1940) observations of the 1930s and his own of the 1960s, he commented that “Fall fawn:doe ratios were considerably below pregnancy rates normally found in mule deer . . . suggesting that preg-

Influences on Upland System Structure IV: The Ungulate Guild

nancy rates in Park mule deer were very low and/or considerable fawn mortality occurred between birth and fall, probably the latter.” The evidence is fragmentary and in some cases surmise. But collectively it suggests elk suppression of deer numbers in the park during the first two-thirds of the 1900s. The censuses were too few in the 1960s and 1970s to determine whether there was a deer population response to the elk reduction of the 1960s. Houston (1982) remarked that there was no correlation between the population trends of the two species. The deer population evidently increased substantially by the latter 1980s, by which time the elk had returned to high numbers. But this reasoning refers to a correlation in temporal trends. Spatially, deer occurred in low numbers when they were sympatric with elk on the northern range. Once they began wintering outside the park, away from most of the elk population, where vegetation had not experienced the near century of heavy use by a large elk herd that park vegetation experienced, the number of deer wintering on the northern range rose to nearly 3 times the numbers of the 1930s and 1940s. Moreover, inside the park where large numbers of elk winter, the number of deer wintering has dwindled to perhaps ~100. Thus there has been an inverse spatial correlation between numbers of the 2 species. Singer et al. (1998a) acknowledge that this may be evidence of competitive exclusion from the park. WHAT CHANGES SINCE PARK ESTABLISHMENT? The number of deer present at the time of park establishment can be approximated, as with bighorn and bison from the data in table 9.1. The ratios from the two historical sources (71 and 54 deer per 100 elk) produce estimates of 3,550 and 2,700 deer, respectively, based on an 1872 elk population of 5,000. These imply declines of 79% and 73%, respectively, to the mean censuses of 738 in the 1930s and 1940s (figure 9.5). And the contemporary number of animals wintering inside the park has shrunken to ~100, for all intents and purposes a decline to no population. With the deer now wintering almost totally outside the park, and censuses averaging 2,043 during the 1990s, the wintering population is now 42% and 24% below the numbers at park establishment. These percentages are conservative depending on the number of animals wintering on the northern range that do not summer in the park.

Pronghorn Antelope ECOLOGICAL POSITION IN THE UNGULATE GUILD. Among the three upland browsing species on the northern range in winter, pronghorn are closest to being obligate browsers. Singer and Norland (1994) report 82–83% woody vegetation in the diet compared with 52–65% for mule deer and 10–15% for elk. Sagebrush makes up the majority of the pronghorn diet. The animals are the smallest of the northern range ungulates and least capable of coping with snow (Houston 1982). Hence they have largely occupied the lower-elevation, western end of the winter range since park establishment.

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Barmore (1980) comments that this area historically was the upper end of pronghorn winter range extending down the Yellowstone River valley. Poaching and Euro-American settlement in the valley gradually eliminated most of the animals except those inside the park at the upper end. With their heavy dependency on sagebrush and occupying a limited portion of the northern range where the other two browsing species have also tended to concentrate, pronghorns would presumably experience heavy constraint on their numbers through competition for forage. I discussed the near elimination of sagebrush in this region in chapter 7. However, relating the pronghorn’s population trends to those of the other browsers is complicated by periodic population reductions through culling during the 1940s through the 1960s. Barmore (1980) summarized early reports of pronghorn numbers and winter censuses through 1970. Bishop et al. (1997) tabulated these and subsequent census data through 1996. The following generalizations about population trends are drawn from these sources. The early (1800s) reports speak of “thousands of antelope,” 500 wintering on Mount Everts, 800 wintering “on flat near Gardiner.” Estimates of numbers in the early 1900s placed the wintering population at 1,000–2,000. Clearly there were substantial numbers of animals in this western end of the northern range in the first few decades of park history. The first three censuses were taken from 1911–18, then carried out annually from 1922 to the present with an occasional missed year. From 1925–46 the censuses varied between 417 and 811, possibly fluctuating in a rough equilibrium during this period. Culling began in 1947, with periodic removals ranging from 6 to 294 through 1967. By 1968, the census had fallen to 149. But by March of the same year, the census counted only 85, possibly reflecting either heavy mortality from the severe 1967–68 winter, the heavy exodus of animals from the park, or both (Barmore 1980). Over the next 15 years, the population remained at roughly the postculling level, the censuses varying from 102 to 165. Both Barmore (1980) and Houston (1982) suggested that coyote predation could have been a significant constraint on population recovery during this period. But from 1983–91, the population increased to a high point of 588 in the latter year, then declined to a 1998–99 census of 204 animals, all remaining in the park (Thomas Lemke, personal communication, May 20, 1999). Any inference about elk competition from these results is equivocal. From about 1910 through 1946, a period of high elk numbers, pronghorns fluctuated at levels almost certainly well below their numbers at park establishment and the next few decades. Theirs and elk herd reductions roughly coincided, so that both species reached low points in 1968. The pronghorns did not increase during the ensuing low elk populations as did bighorn sheep, but they recovered somewhat during the 1980s when elk numbers had increased back to high numbers. Unlike mule deer, pronghorns have been more inclined to winter inside the park, except in severe winters.

POPULATION TRENDS.

Influences on Upland System Structure IV: The Ungulate Guild

But pronghorn numbers have again declined since 1991 to a low point of 204 in 1998–99 during continued high elk numbers. Bishop et al. (1997) cite an unpublished report on a pronghorn population-viability analysis by Daniel Goodman that concludes that the probability of pronghorn extinction in the area in the next 100 years is 18%. This is well within the range in which a species is considered severely endangered. WHAT CHANGES SINCE PARK ESTABLISHMENT? The table 9.1 ratios for early pronghorn numbers, based on historical data, place them at 71 and 50 per 100 elk. Thus the speculative numbers at park establishment, again assuming elk numbers on the order of 5,000, are 3,550 and 2,500. With current wintering pronghorn numbers at ~200, this implies population decline since park establishment on the order of 94% and 92%, respectively.

SYNTHESIS Altered Numbers and Guild Diversity The speculative estimates on the number of northern range ungulates in each species at the time of park establishment, calculated in preceding sections, and the contemporary numbers in each, can now be brought together for a synthetic analysis of the entire upland, ungulate guild. I have summarized these in table 9.2, using the ratios from table 9.1 based on Kay’s (1990) data, and show then graphically in figure 9.6. The approximation of 5,000 elk in 1872 was derived in chapter 3, and the other species estimates follow from this and the table 9.1 ratios derived from the Kay and Schullery and Whittlesey data. Clearly these are highly speculative and approximate estimates, but they point to profound changes in the magnitude and composition of the ungulate guild. The results in table 9.2 imply a 35% increase since park establishment in the total number of ungulates on the northern range. But this innocuousappearing statistic fails to disclose the profound nature of the changes that have actually taken place, including the 2.3× increase in total ungulate biomass. Elk numbers more than tripled 3.2× between park establishment and the 1990s, whereas the other four species together declined 67%. Two species, bighorns and pronghorns, have been reduced to or below minimum population viability. If the Houston (1982) surmise of 12,000–15,000 elk at park establishment were valid, the ratios in table 9.1 would produce an even more extreme decline in the other species than those depicted in table 9.2 and figure 9.6. What once was a guild of five species with low to moderate numbers, evenly distributed among them (figure 9.6), has become a group massively dominated by a single species with the others driven to low numbers. Mule deer, bison, pronghorn, and bighorn together constituted 65% of animal numbers in the guild at park establishment, but their numbers in the 1990s were 16% of the total. This concentration of numbers in a single species of the guild constitutes a sharp

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Table 9.2 Speculative Estimates of 1872 Northern Range Ungulate Numbers and Biomass, and Their Mean, Censused Numbers and Biomass for the 1990s 1872 Species Elk Mule deer Bison Pronghorn Bighorn Totals

Nos.a

1990s Biomass (kg)b

Nos.c

Biomass (kg)b

5,000 3,550 350 3,550 1,900

1,200,000 213,000 157,500 177,500 114,000

16,222 2,043 549 383 105

3,893,280 122,580 247,050 19,150 6,300

14,350

1,862,000

19,302

4,288,360

aSee

chapter 3 for estimate of elk numbers. Estimates of the other species are based on the elk estimate and ratios from Kay (1990) data in table 9.1. bIndividual weights from Houston (1982:157). cNumbers from Bishop et al. (1997) with faulty 1990–91 elk census omitted; northern range bison numbers from Meagher (1993); 1999 pronghorn census from Thomas Lemke (personal communication, May 20, 1999) included.

reduction in evenness, and because the latter is one parameter in diversity indices, the result is a marked reduction in the guild’s diversity. Competition by a large elk herd in all probability explains these changes. As broad-spectrum feeders, elk utilize forage resources used by all of the other species and are implicated in vegetation changes described in the previous three chapters. As highly mobile animals, they access most of the topography used by the other species. The population evidence for competition was persuasive for bighorns, and somewhat less so for mule deer and pronghorns, but nevertheless suggestive.

Figure 9.6 Estimated numbers of upland ungulates in the northern range in 1872 at park establishment (light bars), and average numbers counted in censuses of the 1990s (dark bars). See table 9.2 for sources.

Influences on Upland System Structure IV: The Ungulate Guild

Shifting Spatial Distributions The evidence indicates that the spatial distribution of wintering ungulates on the northern range has gone through three phases since before park establishment. Keigley and Wagner (1998; see chapter 3) concluded that the early evidence indicates fall movement of elk out of what is now the park northern range, down the Yellowstone River valley to lower elevations some distance north of the present park. According to the early reports, this occurred despite the apparently smaller herd size less than one-third the numbers in the 1990s (table 9.2). There may also have been some exodus of bison and pronghorn, at least in extreme winters, according to Meagher (1973) and Barmore (1980). As discussed in chapter 3, according to early accounts the elk sensed the protection afforded by the park soon after its establishment. This began the second phase, a century-long period in which a greatly enlarged elk herd wintered inside the park except for occasional limited sallies in extreme winters (Houston 1982). Reports are more fragmentary on the other species, but as cited early superintendents’ reports recounted large numbers of pronghorn, bighorn, and mule deer within the park. By the 1920s, the early censuses were reporting their numbers, by then already somewhat reduced from prehistoric numbers. The third phase may have begun in the 1940s to 1960s with the increasing tendency of the majority of mule deer to winter outside the park. This was followed in 1988–89 when roughly two-thirds of the elk moved out and on average a third continuing to do so each year. The most recent emigrations are the bison movements, which occur when the park population rises to 3,000 or more (Cheville et al. 1998). Thus, a major fraction of the wintering ungulates now move out of the park in winter, ironically finding a more favorable environment outside than inside. The motivation is in all probability a quest for forage because of depletion inside the boundaries. As discussed in chapter 7, if this pattern continues and intensifies, it could result in a new phase of somewhat reduced herbivory, at least in the upper parts of the northern range.

Magnitude of Herbivorous Pressure The biomass values in table 9.2 provide a better index of herbivorous pressures on the northern range than animal numbers alone. These values imply a 2.3× increase in total ungulate biomass between park establishment and the present. But once again this value for the totals does not disclose the details of the change. Except for a slight increase in bison biomass, the increase in the total is entirely produced by the 3.2× increase in elk. The total for the remaining species declined by 40%. Even these values are conservative as indices of herbivorous pressure. The third of the elk herd, which, on average, now leaves the park in winter, still goes no further than the Dome Mountain area at the northwest extreme of the winter range. Hence the 1990s ungulate numbers and biomass are valid approximations

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of the herbivorous pressure applied to the northern range. Most of that is still exerted within the park because two-thirds of the elk do not exit outside the boundaries, on average. But as discussed in chapter 3, the historical evidence indicates that the northern range elk herd migrated north considerable distances beyond Dome Mountain. Thus the biomass of animals remaining within what is now considered the northern range prior to park establishment must have been substantially less than the amounts shown in table 9.2, perhaps no more than a third. Thus the herbivorous pressure applied to the northern range in the 1990s must have been several times that of 1872 pressure, perhaps around six times. Pressures of that magnitude were applied for most of the 1900s, almost certainly an unprecedented force on the area’s ecosystem.

Apparent Competition and the Role of Carnivores Classical interspecific competition occurs when one or more species use a resource to the detriment of one or more other species needing that resource. A signature criterion of competition is the decline of one species associated with rise of a putative competing second species. I have applied this criterion to bighorn sheep and mule deer. Holt (1977) proposed the term “apparent competition” when the increase of one species depresses the numbers of an ecologically related species through the intermediary of predation rather than resource use. Two ecologically similar species may both be preyed on by the same predatory species. If an increase in one of the prey species contributes to an increase in predator numbers that then prey more heavily on the second prey species and reduce its numbers, apparent competition has occurred. One such case may involve coyotes, which, prior to wolf introduction, traditionally occurred on the northern range at unusually high densities for the species. That density was attained in part through protection from artificial predator control, which has been ubiquitous throughout western United States. But protection alone does not achieve such numbers, and the key additional variable is an abundance of winter food when the females are pregnant (Knowlton 1972; Knowlton et al. 1999). Murie (1940) early reported that elk carrion, available from the yearly winter elk mortality, is the dietary winter staple for coyotes on the northern range. Moreover, elk calves are a major component of spring and early summer coyote diets when the animals are feeding litters. Hence the large elk population in all probability has been a major determinant of the highdensity northern range coyote population. That population, in turn, preys on the other ungulate species. Houston (1982) examined a sample of dead mule deer fawns, and found that coyotes had killed 44%. He, O’Gara (1968), and Barmore (1980) all implicated coyote predation as a source of pronghorn fawn mortality and even suggested that it may have been one cause of the slow population recovery of pronghorns following cessation of culling in the 1960s. At the nearby National Bison Range, which

Influences on Upland System Structure IV: The Ungulate Guild

also has a high-density ungulate guild and a large coyote population, pronghorn fawn mortality averages 90% per year, and the major mortality source is coyote predation (Byers 1997). Thus elk may act as apparent competitors of mule deer and pronghorns on the northern range as well as direct competitors via the forage resource. A second case of apparent competition may develop with the reintroduction of wolves, according to park investigators. As the large elk population facilitates build-up of wolf numbers, the wolves conceivably could place additional pressure on some of the other ungulate species, particularly deer, pronghorn, and bighorn. Pronghorn might also benefit if, as appears to be occurring (see Mlot 1998; R. L. Crabtree, personal communication, January 20, 1999), wolves reduce the coyote population. In the latter case, elk would serve as apparent competitors with coyotes. Finally, a substantial literature discusses the aggressive interactions between the North American medium-sized carnivores (Wagner 1988). Several species are ordered in a dominance hierarchy of wolf–coyote–and a combination of fox, bobcat, and possibly others. Park investigators now suggest that wolf reintroduction and consequent coyote reduction could permit increase of foxes (see Gese et al. 1996). In such a case, foxes would become “apparent commensals” with elk. In total, build-up of the large elk population has resulted in numerous changes in the northern range biota from its structure at park establishment, largely in the direction of impoverishing it. Reestablishment and increase of wolves could effect further changes. Wolves were present in the Yellowstone area in prehistory, but the evidence indicates low numbers reflecting the lower numbers of ungulates (see chapter 3). A large, contemporary wolf population would not be likely to restore prepark conditions in the near future.

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Influences on Riparian System Structure

10 It is interesting to contemplate a tangled bank, clothed with plants of many kinds, with birds singing in the bushes, with various insects flitting about, and with worms crawling through the damp earth, and then reflect that these elaborately constructed forms, so different from each other and dependent upon each other in so complex a manner, have all been produced by laws acting around us. —Charles Darwin

INTRODUCTION The Yellowstone River flows northwestward through that portion of the northern range inside, and over half of the portion outside, the park (refer to figure 1.1). Smaller streams join the Yellowstone in the park and outside YNP as the river turns northward. There are also a number of springs and seeps that create small, wet pockets in the landscape of the northern range and a number of small ponds and lakes (see chapter 12). Houston (1982:86) placed “riparian shrub” vegetation within the park portion of the winter range at 300 ha, or about 0.4% of the northern range inside the park. Bishop et al. (1997:46) report the amount of willow in the northern range historically at 0.8%, and Meager and Houston (1998:246) place the percentage “prior to the 1930s” at “about 1 percent.” The research focus on the riparian zone in the northern range has been largely within the park, and that focus will be maintained in this chapter except for a brief comparison with the type on the outside. 172

Influences on Riparian System Structure

There are a number of parallels between the research treatment of aspen woodland and the riparian zone. The aspen focus dwelt largely on the species itself with little attention to the system of interacting species that comprise the aspen woodland. Similarly, the riparian focus has largely been on willow. Little research has addressed the remaining components of the riparian system except for cottonwood and beaver. Hence a major part of this review will concentrate on willow. As with aspen, there can be some confusion between discussion of the areas occupied by willow growth and the numbers of willow plants. And, as with aspen, discussion on plant numbers can fail to make a clear distinction between root suckers less than 1 m tall and fully grown plants 2–4 m in height. Like the treatment of aspen, the willow literature contains contradictions on chronological trends in the type and among numerous hypotheses of causes precipitating change, often discounting effects of herbivory. These commonly postulate reasons why willow should have been vulnerable to elk browsing in recent decades when supposedly it had not been in early decades of park history, always implicitly crafted for consistency with Houston’s paradigm of large elk numbers in the early years and historic evidence of largely unbrowsed willow.

WILLOW Chronological Trends Willow Abundance at Park Establishment All observers agree that willow was more widespread in the northern range in the early decades of park history than it is today, essentially occupying most if not all moist sites with habitable substrates. It is necessary to rely on the early photographic record and anecdotal observations for reconstructing the early trends. These are consistent and form a coherent picture. The early photographic record shows the type widely distributed through the area: 1. Meagher and Houston’s (1998) photographs show robust willow growth along the Lamar River in 1898 (their figure 59.1) and Soda Butte Creek in 1871, 1884, and 1885 (their figures 60.1, 61.1, 62.1, respectively). 2. Kittams (1948) commented in the caption for his figure 3, “The bottom near Soda Butte was virtually a bed of willows in about 1895.” For his figure 7: “Willows were numerous and extended well above the snow at Yancey’s in 1887.” 3. The abundance of willow in Yancey’s Hole reported by Kittams is also shown for 1893 in figure 2a of Keigley and Wagner (1998). 4. Kay (1990) shows a lengthy strip of willow extending across Soda Butte valley in 1896 in his figure 28a. 5. Kittams’s (1948) figure 11 shows “a dense stand of tall willows [that] adjoined the Gardner River” in 1905.

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Kay (1990:229) cited Captains Barlow and Heap, who toured the Yellowstone area in 1875 and commented on “thickets of willow along the river banks.” He also cited the 1880 report of Norris, the second park superintendent, who noted that the park was “well supplied with rivulets invariably bordered with willows.” Smith et al. (1915:20) commented on the basis of their 1914 survey, “Willows of many species are an abundant source of food supply along many of the valley streams and meadows . . . there are extensive willow bottoms and numerous brushy gulches which afford excellent winter food.”

First Reports of Change The first reports of browsing impacts on willow were in 1915: 1. Smith et al. (1915:20) commented “Wherever they occur on the wintering grounds the willows are trimmed off to mere stumps during the winter.” 2. Bailey (1930), in his figure 20, shows “Willows in summer, showing effects of winter browsing by elk, near Yancey’s, July 15, 1915.” His figure 19 shows hedged willows “near the Tower Falls Ranger Station” in a March 10, 1916 photo. Although there is convincing evidence of browsing impact on willows in the second decade of the 1900s, willows were still a conspicuous component of the vegetation in the 1920s: 1. Figure 10.1a is a 1921 photo of the Lamar River showing robust willow growth on the south shore, reproduced by Kittams (1948) as his figure 1. 2. 1921 photographs of beaver ponds on Elk Creek near Tower Junction were reproduced by Kay (1990:figure15a) and Meagher and Houston (1998:figure 49.1) and show willow growth on their fringes. 3. Meagher and Houston (1998:figure 63.1) show an extensive 1921 willow community across the Pebble Creek floodplain in the northeast corner of the park. Their figure 45.1 also shows what appears to be a thin fringe of willows around a beaver pond near Wraith Falls in 1924. 4. Kittams’s (1948:41) photo taken “near the mouth of Soda Butte Creek in 1926”(his figure 5) shows willows forming “a heavy cover in the bottom.” His Figure 9 (p. 45) of the “Gardner River above Mammoth” shows “several willow bushes produced rank growth along the right bank.” Yet there was growing evidence of willow decline in the 1920s. In the caption for his 1926 figure 5 mentioned above, Kittams (1948:41) commented on the willows near the mouth of Soda Butte Creek, “Note the numerous dead willow branches . . . suggesting prior abuse.” And elsewhere (pp. 21–22) from his observations in the 1940s: typical of the upper Lamar Valley . . . This virtual annihilation of willows can be attributed primarily to elk. . . . The lack of dead branches suggests that the die-off was completed some years ago. This seems to preclude the drought of the 1930s . . . dead willow branches ordinarily stand several years before falling. . . . severe use occurred prior to 1926.

Influences on Riparian System Structure

Figure 10.1 (top) Willow growth along the Lamar River in 1921 (from Kittams 1948). (bottom) Kittams’s (1948) retake of the same site in 1948.

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Meagher and Houston’s (1998) figure 57.1 shows a 1923 scattering of willows in the lower Slough Creek valley. The shrubs have numerous dead branches. Although the authors do not consider them to have a hedged appearance, the low, clumped stature of the live portions of the shrubs suggests classical hedged structure. Wright et al. (1933), in their figure 50, show a 1929 photo of the Lamar River along which the stream bank shrubs appear punctuated and reduced in stature by comparison with the 1921 photo in figure 10.1, top.

1930s to Herd Reductions There were very few actual observations in the 1930s. Houston, who conducted his research in the 1970s, has stated three times (1976:97, 1982:129; Meagher and Houston 1998:230) in almost verbatim the same statement: “Mortality was most conspicuous on upper flood plain terraces and occurred mainly during the drought of the 1930s.” But no evidence has been provided to support the statement. Yet park publications repeatedly assert that most of the decline occurred in the 1930s (see Bishop et al. 1997:47; Meagher and Houston 1998:246). The one published willow photo for the decade was taken in 1932 near its beginning, and therefore reflects influences for some years previously. It is reproduced in Wright and Thompson (1935). Its caption: “Willow thickets killed along the Gardner River by overbrowsing. This is characteristic of the plight of willows over the elk winter range generally.” Park biologist Walter Kittams (1948, 1952) began the first park repeat-photo series by photographing scenes in the latter 1940s that had originally been taken by photographers from the late 1800s to the early 1900s. Hence they provide a basis for observing change by the 1940s from the conditions in the early photographs. The captions for his six 1948 pictures state: The [Lamar] River bank is now covered by grass in contrast to willows shown in Fig. 1. [His figure 2 is shown here as figure 10.1, bottom.] near Soda Butte Creek . . . the bottomland cover near Soda Butte is now grass-sedge. [His figure 4, a retake of the 1895 photo.] Little remains as evidence of the willow thicket. [His figure 6, a retake of a 1926 photo of Soda Butte Creek listed as (4) above under 1920–29.] much of the remnants of willows at Yancey’s are dead branches, much different from the picture of 1887. [His figure 8, a retake of the 1887 Yancey’s Hole photo showing willow growth.] only a few of the willow plants are alive now. [His figure 10, repeating a 1925 photo of the Gardner River, no. (4) above under 1920–29.] Practically none of the willows, shown in Fig. 11, remain. [His figure 12, also of the Gardner River, is a retake of a 1905 shot.] The captions for Kittams’s (1950) photos taken in 1949 comment in a similar vein: “Willows [near Pebble Creek] have a hedged appearance as though picture was taken at beginning instead of end of growing season” (his figure 20, 1949). Kay (1990:161) reproduced this same photo, and when he rephotographed the same scene in 1988, almost no willows showed in the picture: “In

Influences on Riparian System Structure

25 years willows along river and stream in distance have become fewer and shorter [Kittams’s figure 32, 1949]. Park visitors now see tall, dead willow branches. [Kittams’s figure 34, 1949].” Cahalane (1943), commenting on northern range conditions in 1943, remarked: “However, much of the edible portions of shrubs and trees, especially of fir, juniper, willow and aspen, were already stripped as high as the elk could reach.” I have not located photographs of willow taken in the 1950s. Kittams’ set was taken at the end of the 1940s. Houston’s (1982; Meagher and Houston 1998) repeat series was not started until he arrived in the park in the 1970s, and for some reason did not include the early Iddings photographs, used by Kittams, for comparison. Kay (1990) did not begin his photo archive search and taking repeat photos until the 1980s. However, Kittams (1959) commented at the end of the 1950s: “Now willow thickets are almost non-existent, but dead or badly mutilated bushes evidence their former state.” Park publications have commented that the decline in willow ended in 1959 (see Anonymous 1992): “The willow of the Northern Range have not changed appreciably since about 1959, when the declines ceased.” No evidence for this view has ever been cited unless it is based on the earlier reports that willow had been almost completely eliminated as a shrub type by 1950s, and no further decline was possible. The first half of the statement cites a paper in a conference proceedings (Despain 1994), but the paper does not appear in that document. There is contrary evidence. Barmore (1980:353) commented: “Thus . . . ungulate browsing suppressed the aerial cover and height of Salix sp. from sometime prior to 1958 through 1965.” Subsequently (pp. 357–58): Other photos . . . show that Salix sp. was taller, less heavily hedged, and less decadent, and Salix communities were more widespread on parts of the winter range prior to about 1930 than during the early 1960s. . . . Houston (1976) suggests that reduction of soil moisture during the 1930s drought, more than increased browsing by ungulates, may have been responsible for the decline of Salix sp., but other evidence suggests that increased browsing by ungulates (primarily elk) was more important. Barmore was commenting in the context of his measurements inside and outside the large exclosures, discussed in previous chapters, that were established in 1957 and 1962. These measurements included numbers and heights of plants and cover and have resulted in confusion over statements about willow trends when the parameter under discussion was not specified. Up to this point in this chapter, the comments by early authors on willow trends have referred to the areas occupied by stands of shrubs. But beginning in the 1960s, the literature changes largely to discussion of plant numbers or density, aerial cover, and height irrespective of whether they are full-statured shrubs or root suckers <1 m tall. Richard Keigley (personal communication, July 7, 1999) has shown me copies of photographs in park files taken in the 1950s of willow transects on south-facing hillsides in the Lamar River valley. These showed scattered, heavily hedged willow plants. Inspection of the same sites on July 7, 1999, showed

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fragmentary remnants. Some plants in the original photos had either disappeared, or were present as dead stems.

Changes During the Herd Reductions Barmore (1980:353–56) measured willows on belt transects inside and outside the 2.1–ha exclosures in 1958, 1962, and 1965 at the first four exclosures and in 1962 and 1965 at the second four. Both cover and plant height increased by 1962 inside the first group and by 1965 in the second. But these parameters did not change statistically outside the exclosures. The distribution of willow plants in the transects inside and outside the exclosures did not change, and Barmore concluded, “The increases [inside the exclosures] in aerial cover was [sic] due to the enlargement of existing plants rather than establishment of new plants.” These measurements were repeated by other park investigators in 1974, 1981, 1986, 1989, and by Kay (1990:142–50) in 1988. The trends shown in Kay’s tables 33 and 35 (see also figure 10.4 later) continue the pattern observed by Barmore with little or no significant increase outside the exclosures but marked increase in canopy cover and plant height inside. Barmore’s measurements showed continued high browse-utilization rates on willows during the elk reductions of the 1960s, and park publications (Anonymous 1992:22; Singer et al. 1994; Bishop et al. 1997:48) have implied that there was no willow response during the period. But Barmore (1980:357) stated: By the late 1960’s [note this is after the 1965 measurements but before those of 1974] the growth form and condition of Salix sp. on most of the winter range began to more closely resemble the less heavily browsed conditions of the late 1800’s and early 1900’s. This change was associated with major reduction of the northern Yellowstone elk herd. Bart O’Gara (personal communication, February 8, 1996), a doctoral student at Montana State University conducting research in YNP in the early 1960s, told me that he observed a response in willow growth within 2 years after herd reduction. He completed his degree in 1967, left the park, then returned in 1972. At that point the herd had recovered to nearly 10,000, and willow growth was once again suppressed. Patten (1968) studied willow ecology along the Gallatin River in the northwest corner of YNP in 1963. He reviewed 1924 photographs of the area and observed significant reduction in willow growth by the time of his study, which he attributed to elk browsing. When he published in 1968, he observed some recovery from the low point reached at the beginning of his study. Kay (1990:152) shows a 1965 park photograph of a belt transect outside the Junction Butte exclosure. It depicts a dense stand of willow plants perhaps 0.5–1 m high. Kay’s retake of the site in 1988 shows almost complete elimination of willows and conversion to grasses and perhaps sedges. The abundant 1965 growth may have been a response to the herd reduction, and the barren 1988 photo clearly represents willow decline after 1965.

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Keigley (1998; Keigley and Frisina 1998, and see cottonwood section below) has shown that a woody plant’s browsing history can be reconstructed by characterizing its architecture and aging its growth shoots, either by coring or crosssectioning. Although he did not conduct intensive research on northern range willow, his observations of the species in the area (figure 10.2) prompted the following comment (R. B. Keigley, personal communication, 1994): Throughout the northern range, it is common for willows to have tall, unbrowsed branches growing next to shorter, heavily-browsed branches. . . . this combination suggests an increase in herbivory that abruptly arrested the height growth of those shorter branches. . . . Branch architecture clearly shows that the reduction program allowed willows to grow taller. . . . One can surmise that browsing on willow has increased substantially within the past few years. No data suggest that willows declined up until 1959 and then stopped declining.

Postreduction Trends The periodic belt transect measurements inside and outside the large exclosures summarized in an unpublished report (Singer 1996a) included in the material

Figure 10.2 This willow in the Lamar River valley has existed under at least two periods of browsing intensity. In an earlier period of light or no browsing, the tall branches were able to grow up through the browse zone, extending annual growth shoots out of reach of browsing elk. A recent period of heavy browsing maintained the lower branches around the base of the plant in hedged form. The three low hedged plants beside the observer are probably younger plants that have spent their entire lives under the heavy browsing of recent years. Photograph taken July 17, 1997.

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submitted to Congress, has been the basis for park statements on willow trend up to the present. But those made in Bishop et al. (1997) and Singer (1996c), and between the text and figures of the latter source, are contradictory and do not allow inference of what long-term trends the studies really show. Moreover, the confusion is compounded by failure of some of the statements to specify the measurement parameters to which they refer. Thus Bishop et al. (1997:47) state, “Since . . . [the 1930s drought] there had been little change in willow status.” And two pages later, “Willow status on the northern range was relatively stable or even improving during the past three decades.” Singer (1996c) commented, “Numbers of above ground willow clumps declined at two browsed sites and increased at two other browsed sites, 1957– 89.” But on p. 279 Singer (1996c) stated that “willow heights . . . increased on browsed transects during the period 1957–89 . . . but the height increase (only 6cm) was probably not biologically important” (emphasis added). Yet page 287 states, “Average heights of willow clumps declined 6cm on the average, suggesting a slow decline in stature” (emphasis added). A following sentence remarks, “Willow cover increased on browsed transects 220% over the time period, primarily due to root suckering of browsed clumps, but the browsed clumps were very small above ground clumps.” On page 285 the author states, “Total canopy area averaged . . . more on unbrowsed than on browsed sites.” a result consistent with Barmore’s (1980:334–35) and Kay’s (1990:tables 33, 35) findings. But figure 2 on the same page shows canopy cover higher outside the exclosures than inside. Singer (1996c:279) concludes: The continued persistence of height-suppressed willows is problematic, however, because average heights decreased slightly (11%) over the study period . . . numbers of willow clumps declined at two of four browsed sites over the 32–year period, many other stands have disappeared, and a general lack of seed production and new seedling establishment exists across most of the Yellowstone winter range. Kay (1990:159) commented on the Tower Junction exclosure. Constructed in 1957, the willow inside the fence increased in height and canopy coverage by 1962, and further by 1965 (Barmore 1980:354). The exclosure was dismantled in 1971, and the willow was browsed down by 1973.

Summation on Chronology Full-statured riparian shrub communities, dominated by willows but containing other shrubby species, were clearly a conspicuous component of the northern range ecosystem in the early years of park history. The total area on the northern range was minor, perhaps making up no more than 1%, but they contributed to the habitat, landscape, and ecosystem diversity of the area. The first report of browsing impact was by Smith et al. (1915). But in view of the authors’ statement (p. 20), “Wherever they occur on the wintering grounds the willows are trimmed off to mere stumps,” significant impacts must have occurred for some period prior to their survey. Bailey (1930) also photographed

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hedged willows in 1915 and 1916. The fact that there were no earlier reports may be due to the fact that there were no earlier professional investigators to observe and report on the matter. Although there are photographs of robust willow growth still in the 1920s, there was widespread browsing suppression by this decade. Thus Kittams’s (1948:21) remark about “virtual annihilation of willows . . . typical of the upper Lamar Valley,” which he concluded to have occurred before the 1930s. There is no significant record for the 1930s, but Kittams (1948, 1952) abundantly documented the now pervasive impacts and decline of willows by the 1940s. The park contention that decline ended by 1959 may simply reflect nearly complete elimination to the point where no significant decline was any longer possible. Yet there was, in fact, further decline in those limited areas where some willow was still present: Willow that had recovered to some degree during the herd reduction, and the new growth inside the Tower Junction exclosure, disappeared with its removal. Most park publications maintain that there was no willow recovery during the herd reductions. But Barmore (1980), O’Gara (personal communication, 1996), Patten (1968), and Kay (1990) all report evidence of some response. That it was not more marked was doubtless due to the short period during which the herd was reduced, and the fact that willow had been depleted to such low levels that even a sharply reduced elk herd could continue to exert significant pressure on this highly palatable class of browse. Once the herd recovered, the response was obliterated. It is impossible to portray the chronology of willow decline quantitatively with the evidence available. Park publications state that most occurred during the 1930s, but there is no clear evidence of this. It does appear that decline in an absolute sense was pronounced in the 1920s, 1930s, and 1940s, but there is no basis for judging the decline rate to have been any higher in the 1930s. Park publications, starting with Houston (1982), have claimed 50% reduction in willow. But this point is confounded by estimates of the percentage reduction in full-statured shrubs versus the area occupied by the species, even if only by root suckers that may be no taller than the grass and sedge matrices in which they occur. Kay (1990) estimates 95% decline of what he called “tall willows.” In view of the photographic record (see figure 2b in Keigley and Wagner 1998, figure 10.1 here), the elimination of willow shrubs in the northern range must approach Kay’s percentage more closely than the park’s 50%. To what extent the area occupied by the species, irrespective of plant stature, may have changed is impossible to say.

Causation General Considerations The science of ecology operates on a premise of mechanism. Phenomena are produced or determined by causal factors or “determinants,” which operate in hierarchies of influence, some with major effect, some with minor (Walker and Noy-Meir 1982). Causation also operates in chains and networks of determi-

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nants. A given phenomenon may be driven or induced by one or more first-order factors that impinge on it directly. In turn, each first-order factor and its degree of effect is driven or affected by one or more second-order factors, they in turn by third-order factors, etc. A given factor or determinant may exert influence through more than one chain and may operate in more than one ordinal position. Both positive and negative feedbacks add to the complexes. Ecology strives to elucidate these complexes by identifying the causal factors, ordering their positions in causal chains and networks, and quantifying their effects. Identifying and listing determinants of a phenomenon are a necessary first step. But ecological understanding is not significantly advanced until an investigator delineates the ordinal positions of factors and, to the extent possible, at least qualitatively, assesses the relative importance of each factor’s effects. The ultimate goal is simulation models that represent these relationships quantitatively and depend on quantification of the state variables, process rates, and effects of constraints on the processes. I pose this conceptualization of ecological causation at this point because eight or more causes of willow decline have been proposed by different investigators. In many, perhaps most cases, there has been no attempt to suggest their relative importance within complexes of factors. For some there is no evidence, or what is offered as evidence does not really serve that function. Some are, if real, lower-order causes with no suggestion or evidence of the causal chains that connect them to first-order factors affecting willow. My purpose here is to place these hypotheses in ecological context in order to more effectively evaluate them. Ideally, the processes causing decline in northern range willows should be synthesized into a model representing the interactions quantitatively. However, the change has occurred in the past with only limited measurement of some of the constituent processes, state variables, and causal factors. Thus no wellquantified model is possible at present. But the existing evidence and hypotheses can be combined into a conceptual model that itself becomes a hypothesis. Validity of the proposed relationships can be assessed by weighing the supporting and contrary evidence for each, and ultimately a model representing the best pro tem judgment of the phenomenon remains. That the judgment should be pro tem is not a deterrent to formulating it. As discussed in the preface, science never gives certainty. All scientific conclusions are pro tem. Although there is only limited evidence with which to construct a model of the extremely complex set of relationships constituting the effects of ungulates on the northern range, the profession has an obligation to do so to evaluate the effects of existing management programs and illuminate, to the degree possible, contemplated decisions on management alternatives. Figure 10.3 is an oversimplified conceptual model of the overall process of willow decline on the northern range showing hypothesized factors, both those for which there is and is not significant evidence. The boxes represent state variables, the arrows processes, and the single bow tie a constraint or factor affecting the rate of the elk herbivory process.

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Figure 10.3 Simplified conceptual model of state variables (boxes), processes (arrows), and regulating factor (bow tie) shown or hypothesized to induce willow decline on the northern range. Arrows show directions of effect. List at bottom classifies ordinal positions of influences in causal chains.

As represented here, elk, moose, insects, habitat, and fire are first-order determinants that impinge directly, through their processes, on willow abundance. Climate and beaver, in affecting the first-order determinant willow habitat, are second-order determinants. Climate is also postulated to function as a third-order influence on elk herbivory by producing willow habitat that affects willow health; that in turn is said to affect the plants’ ability to produce secondary compounds that are hypothesized to influence the rate of elk herbivory. Thus, elk also function as a fourth-order factor by directly reducing willow needed by beaver for building material and food, thereby contributing to willow decline, and ultimately reduction in beaver-constructed willow habitat. This scheme is highly abstracted and oversimplified. Some hypothesized factors for which there is no evidence are not included. Any effort at developing

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a realistic simulation model of the complex would involve reduction to numerous lower-level processes not included. Thus the herbivorous removal of plant tissue would affect plant physiological and demographic processes, such as photosynthesis, growth, reproduction, and survival, with ultimate effects on plant biomass and numbers. Climate would operate through a number of hydrologic processes to affect willow habitat, which would in turn affect willow physiology and demography. But with the limited amount of information available, such detail is neither possible nor desirable at this time.

Hypothesized Individual Factors ELK HERBIVORY. Several lines of evidence point to elk browsing as a major firstorder influence in the northern range willow decline. One is the chronological correlation between elk numbers and willow impacts. The earliest photographs and anecdotal reports indicate robust riparian communities in moist areas of the northern range. These coincide with and follow a low elk population that migrated out of the park area in winter prior to 1872 (chapter 3) and continued into the latter 1800s at low levels with partial migration. Bishop et al. (1997:48) make the quite valid point that the northern herd had risen to substantial numbers (e.g., 10,000+) by the latter 1800s, yet photographs and observers’ comments through the remainder of the century and into the early 1900s indicated healthy willow stands in many areas. However, as discussed in previous chapters, there was a lag of 1 to 2 decades in the response of virtually every sector of the biota to major changes in elk numbers. Particularly in the early stages of park history, it may well have taken this length of time for even a large herd to make significant inroads into a lush vegetation and the effects to become discernible in photographs and early observations. Moreover, there were no professional observers until 1914, and hence it is not known whether some effects were evident prior to this date. The evidence is persuasive indicating a continuous willow decline between the first Smith et al. (1915) report of heavy impact and the major elk reduction in 1961–62. This was the final 46 years of a 7-decade period during which censused elk numbers were 10,000 or more, at times 2 to 3 times this number. Although a number of park publications have stated that most decline occurred during the 1930s, no evidence has been shown to support this assertion. The evidence does show decline underway prior to and after that decade. Park publications have repeatedly stated that there was no willow response during the herd reduction. But I have summarized evidence, assuredly limited, to the contrary above. Again, the low-population period was only ~ 11 years. Given the response lags in virtually all parts of the system, this period may have been too short for a significant recovery of willow still subject to some degree of ungulate use. Moreover, the tall shrub form of these highly palatable species had been virtually eliminated. Conceivably the low population level to which the elk herd had been reduced was still sufficient to deter or at least slow recovery. A second line of evidence pointing to elk herbivory as a determinant of willow decline is the exclosure evidence (Barmore 1980:354–55; Kay 1990:149). Con-

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struction of the 1957 and 1962 exclosures in moist areas—Mammoth, Junction Butte, Lamar East, Lamar West—was followed by immediate willow growth inside but no significant response outside (figure 10.4). Today these contain dense thickets of willow shrubs 2–4 m high (figure 10.5). Thus climate and site conditions at these sites have been conducive to robust willow growth over the past 40 years as long as elk have been excluded. All of the sites are upland areas with

Figure 10.4 Trends in willow canopy cover inside (a) and outside (b) northern range exclosures, 1957–88. Data from Kay (1990:table 35).

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Figure 10.5 Willow growth in Junction Butte exclosure. Photo taken August 15, 1998.

springs or water seeps, considerable distances from any streams. At each site the surface moisture extends outside the exclosures, but no significant willow growth has occurred in the absence of protection from the elk herd. Most recently, Ripple and Beschta (2004) show willow regrowth along Blacktail Creek in 2002, which the elk now avoid because of the presence of wolves. The third line of evidence is the condition of willows within or near the northern range but outside park boundaries. Outside the park where substrates are suitable, willow thickets appear along the banks and bottoms of the Yellowstone River. These become dense and extensive north of Yankee Jim Canyon as the Yellowstone Valley (Paradise Valley) widens. In the adjacent Tom Miner basin, dense willow growth, similar to that shown in early photos of Soda Butte Creek and the Lamar River bottom, fill the flood plain (figure 10.6). In total, these lines of evidence along with nearly a century of anecdotal information and a longer photographic record leave little doubt that elk browsing has been a major force in the northern range willow decline of this century. Bishop et al. (1997:56) concede its importance: “There remains no question that ungulate browsing is the immediate cause of the decline of aspen and willow on the northern range.” But the authors follow with a caveat: “but there is considerable uncertainty over why that browsing has a different influence now than it had historically.” This is based on the premise that elk were abundant in the region prior to park establishment.

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Figure 10.6 Photos of willow growth in (top) the Tom Minor basin northwest of Park’s northern entrance by Charles E. Kay, September 3, 1983, and (bottom) the Lamar River bottom, July 17, 1986.

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MOOSE HERBIVORY. Park publications (Houston 1982; Bishop et al. 1997; Meagher and Houston 1998) have invoked moose herbivory as one influence in willow decline. Except for a Chadde and Kay (1988) study conducted outside the park and an unpublished report in YNP files cited by Bishop et al. (1997:86), there is no evidence to evaluate the effects of moose on the northern range. However, comparison of moose and elk numbers provides some basis for comparing relative effects. Moose were extremely rare in the Yellowstone region in prehistory and in the 1800s. No moose remains have been found in archaeological sites in the Greater Yellowstone Ecosystem (Kay 1990:297), and only one of 23 explorer and military parties traveling through the region during the 1800s reported in their journals seeing a moose (Kay 1990:273). The species “colonized” the Yellowstone area by the 1870s and began appearing on the northern range about 1913 (Houston 1982:158). Northern range moose counts varied largely between 37–100 from 1968– 69, when they were begun, and 1989–90. But numbers declined in the 1990s, averaging 19 in five censuses of the decade. Elk have therefore outnumbered moose on the northern range by 2 to 3 orders of magnitude for the past 30 years, and probably throughout the twentieth century. Thus, there is no basis in the available evidence for inferring that moose have played a significant role in the northern range willow decline by comparison with the elk impact. The difference in effect between the 2 species is probably somewhere near the difference in numbers: approximately 2 to 3 orders of magnitude. INSECT HERBIVORY. Houston (1982:131, 134) describes an outbreak of the beetle Disonycha pluriligata in 2 consecutive years in a “stand” of willows along the Gardner River. The beetles defoliated the plants and killed them. This incident has been cited by him and subsequent park publications (Bishop et al. 1997; Meagher and Houston 1998) as an example of the complexity of factors contributing to the willow decline. One can grant the validity of these observations, but there is no evidence to suggest that such outbreaks were frequent, rangewide phenomena of the northern range nor for assuming that this was more than a localized episode. I do not suggest that it might not have occurred at other locations or at other times. But willows have been measured by park personnel numerous times since construction of the 1957 and 1962 exclosures, and by Kay (1990), Kay and Chadde (1991), and Chadde and Kay (1988). No one has reported a similar outbreak. Hence there is no basis for assuming that it was other than an infrequent, localized phenomenon; there is no indication that this form of herbivory has played a significant role in the willow decline. CLIMATE CHANGE AND PLANT VIABILITY. Climate change has been invoked in a number of park publications as a cause of willow decline on the northern range. In all cases the hypothesized effects operate through the plants’ habitat

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quality and either affect their viability directly or weaken them in some way to affect their susceptibility to herbivory. Two climate changes have been implicated: the drought of the 1930s and the purported century-long temperature rise and precipitation decrease discussed in Chapter 6. Thus, Houston (1982:131) commented: “browsing may not have increased but . . . soil moisture decreased on flood plains with the drought of the 1930s, thereby reducing the plants’ ability to survive browsing . . . the hypothesis that . . . changes in distribution of willow were mostly climatically determined.” And in Bishop et al. (1997:47): other suggested causes include . . . climate variations, especially the drought of the 1930s, which, through changes in groundwater levels or other processes, might have either directly killed the plants or reduced their ability to produce secondary defensive chemical compounds that make them less palatable to grazers. [emphasis added] Meagher and Houston (1998:236) continue: “The relatively recent occurrence and locations of this browsing impact suggest that changes in soil moisture relationships adversely affected the willows’ abilities to withstand browsing. . . . Changes in soil moisture may have resulted directly from the climate events of the 20th century” [emphasis added]. Three separate sets of processes are contained or implied in these hypothetical statements. One is simply willow death as a result of falling water tables and/ or soil moisture, which render sites incapable of providing the species’ water needs. Browsing impact is not included conceptually. Declining habitat quality is the first-order determinant of willow decline, and climate change, as it affects willow habitat, is a second-order factor determining willow habitat quality. A second set of processes implied in the quotes is the physiological inability of plants to withstand browsing when hypothesized declining water tables and/or soil moisture reduce the habitat’s water availability below the species’ requirements. A third set is the physiological inability of willow plants, subject to moisture inadequacy again in their habitat, to synthesize secondary compounds that allegedly would deter elk browsing. In the latter two models the first-order determinant of willow decline is ungulate browsing, whereas habitat decline is second order, and climate change third. Kittams’s (1948:22, 23) skeptical comment on the possible role of the 1930s drought and falling water tables indicates that this causation of willow decline was suggested at an early date. But there were no consistent measurements of soil moisture and water tables before, during, and after the 1930s drought, or of willow growth and abundance which would permit correlations. And there were certainly none throughout the 1900s during which there have been small changes in temperature and precipitation discussed in chapter 6. Bishop et al. (1997:47) comment that “willow stands and individual willows died in the summers of 1988–89 apparently as a consequence of the severe drought of 1988.” When measured in 1988, these had extremely low water vapor pressure.

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But beyond this, the case for climate change as a major direct force on willow decline on the northern range is largely surmise, and there is contrary evidence. Kittams (1948:22, 23) commented: Study plots maintained through the drought period seem to minimize [it] . . . as a factor in decline on good sites. . . a change in water table is sometimes suggested as being responsible for death of willows [near Yancey’s Hole]. The rank growth of willows within an exclosure in a similar site little more than a mile up Lost Creek obviously depicts the benefits of protection from browsing and suggests that the inability of the plants at Yancey’s to maintain and reproduce has been due to browsing. Houston (1982:131) reports a 1934 measurement of the Gallatin River showing flow reduction to 37% of normal. The Gallatin River bottom and surrounding mountain slopes in the northwest corner of the park support a small wintering elk herd held at much lower levels than the northern herd by hunting outside the park boundary. Hedged less severely than northern range willow, the species have survived up to the present in the river bottom. And an exclosure alongside the stream supports a dense growth of tall willows (photo by Charles E. Kay, July 20, 1986). Despite the reduced 1930s stream flow, willows did not disappear from this riparian zone. Kay (1999) and Kay and Walker (1997) show matched 1910 and 1994 photographs of Miners Creek 100 km west of YNP, within the GYE, and at an elevation comparable with the northern range. The 1994 distribution of willows appears to be identical to that of 1910, except that a road was cut through one narrow strip at some date prior to the 1994 picture. Houston (1982) commented on widespread disappearance of willows along western streams at the time he wrote. But it is not clear on what this was based. Willow suppression does occur in western areas subjected to heavy ungulate use, both wild and especially domestic (Shute 1981). But except where subjected to such use, Intermountain streams today are characteristically bordered by healthy riparian zones, their extent determined by stream size and morphology. There is no evidence of massive West-wide willow disappearance comparable to what has occurred on the northern range, except again under heavy ungulate use. CLIMATE CHANGE AND BROWSING SUSCEPTIBILITY. This model implied by the Houston and Meagher and Houston quotes, and the model discussed next, tacitly acknowledge that elk browsing has been the dominant, first-order force reducing willow abundance on the northern range during the twentieth century. Both seek to reconcile willow decline with the Houston paradigm that elk were abundant and wintered on the northern range in prehistory and the early years after park establishment, yet the photographic and anecdotal evidence shows robust willow growth throughout the northern range. Climate change is implicitly a second-order factor modifying the effects of herbivory. A more subtle and perhaps often unrecognized implication of these authors’ writings is that northern range willow must have been subjected to herbivorous pressure in the

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1800s and before according to the model considered here and may not have been subjected to that pressure in the model considered next. Here again, this model is almost pure surmise. At the time Houston (1982) proposed it, there had been no ecophysiological research to demonstrate decreasing ability of willow to compensate for browsing under moisture decline or any other observations over time that would permit correlation tests of changing susceptibility and climatic variation. The closest evidence was reported by Patten (1968), who observed that two willow species at lower levels along the Gallatin River, and nearer the stream, produced more postbrowsing shoot growth than species at higher levels with lower soil moisture. Two lines of evidence cast doubt on the validity of this model as a general explanation of the northern range willow decline. One is that heavy browsing impacts on willows were observed as early as 1914 by Smith et al. (1915), and decline was well under way by the 1920s, well ahead of the 1930s drought, and barely into the putative temperature and precipitation changes of the twentieth century. The second line is the stature of northern range willow today in moist areas unprotected from browsing. Willows have disappeared from much of the Lamar River banks (figure 10.1b). The water table is at the surface outside the Mammoth, Junction Butte, and Lamar East exclosures with willows growing in sites with surface moisture. Yet these plants are suppressed to a few centimeters in height. The abundance of moisture affords them no significant resilience against the browsing pressure to which they are subjected. Thus there is little (if any) evidence supporting this climate-change model that northern range willow was somehow able to withstand herbivorous pressure by a large elk herd under an unspecified, more favorable climate prior to and in the early years after park establishment and maintain robust growth. The photographic and anecdotal evidence attests to that robust growth. But the evidence cited in chapter 3 indicates a low elk population largely migrating out of the park in winter, and thus exerting very little browsing pressure in the early years. CLIMATE CHANGE AND SECONDARY COMPOUNDS. The third climate change hypothesis was proposed by Singer et al. (1994) and reiterated by Cates et al. (1999). The first authors measured tissue tannin content, above-ground production, and browsing intensity (% leader use) in eight species of willow over an elevational range of >600 m in the park’s northern range. They also measured tannin content of plants inside and outside exclosures. Willow shrubs on average were taller, were less intensively browsed, had higher tannin content, and had higher above-ground production at higher elevations (e.g., >2,200 m) than in “height-suppressed” willows at lower elevations (<2,200 m). Plants inside exclosures had higher foliar tannin content than those on the outside. The authors concluded that willow plants on marginal sites could not produce as much foliar tannin as those on more favorable sites, and this in turn attracted heavier browsing pressure and shrub hedging. The inference of lower

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tannin production and heavier browsing on suboptimum sites was extended to a hypothesis that a less favorable climate of the twentieth century had impaired northern range willows’ ability to produce tannins. In turn this hypothesized decline in tannin content rendered the willows more palatable and attracted heavier browsing in the 1900s, leading to the willow declines now acknowledged by all observers. By implication tannin contents had been higher in the 1800s and browsing intensity lower. Wagner et al. (1995b) questioned this model primarily on the grounds that causal inference could more reasonably be reversed. Snow depth and winter severity are a function of elevation on the northern range, and elk tend to seek the lower elevations in winter. Hence low-elevation willow would sustain heavier browsing pressure and height suppression than plants at higher levels. This expectation is borne out by the higher utilization rates measured at lower elevations by the authors and was considered by Wagner and colleagues to be the more probable cause than lower tannin levels at lower elevations. Moreover Wagner et al. (1995b) suggested that rather than site conditions, it was the heavy utilization rates of the lower-elevation shrubs that impaired their ability to produce higher tannin levels. This is a well-recognized phenomenon in the secondary-compound literature (e.g., Rhoades 1979; Cates et al. 1983; Louda and Rodman 1983). This suggestion was supported by the higher tannin levels in willows inside exclosures than outside all at lower elevations. Because the hypothesis of lowered tannin levels in the 1900s was posed by analogy with the inferred relationship between site favorability and tannin levels, and this inference is now clearly uncertain, there is little or no basis of support for a historic decline in tannin levels. In any case this remains a hypothesis because no secondary compound measurements were made before and through most of this period of surmised decline. As a follow-up study, Cates et al. (1999) tested the hypothesis that unbrowsed willows growing on the northern range in purportedly suboptimum conditions would not be able to respond to clipping treatments by increasing production of defensive compounds. They clipped 10–15 shoots on plants (exact numbers were not specified) in northern range exclosures (again, which ones were not specified). The plants were clipped at 50% and 100% of current annual growth with a matching control set left unclipped. They were clipped in the summers of 1993, 1995, and 1996, and in winters only of the first 2 years. Total phenolics, tannins, and soluble carbohydrates were measured in the clipped material from each sampling. Two willow species, S. bebbiana and S. pseudomonticola, were studied. The results were mixed and the series short, so generalization is risky. The pattern that clearly stands out is the steep rise in summer phenolics in both species in 1996 over the 2 previous years, and in all treatments. There appeared to be a slight tendency in this direction in the tannins. In 4 out of 6 cases, S. bebbiana phenolics increased summer–winter, but in all cases they decreased in S. pseudomonticola. In all treatments in both species, tannins decreased summer–winter, another clear pattern.

Influences on Riparian System Structure

The authors comment on an increase in precipitation in the region in “the late 1990’s” and suggest that this is associated with the large increase in summer phenolics in both species in 1996. (I review these results in more detail shortly, but I observe here that summer phenolics declined in all treatments in S. bebbiana from 1993–95, and in the controls of S. pseudomonticola. Moreover the tannins declined from 1993–96 in all treatments in S. pseudomonticola, and in 2 of the 3 treatments in S. bebbiana.) Cates and colleagues then hint, without saying so explicitly, that the climate changes in the twentieth century are partly responsible for declining habitat quality of willow, and “suggest that browsing may not be the only factor in willow declines.” I do not question the possibility that willows growing in suboptimal habitat (especially moisture conditions) will not function at their full physiological potential. But once again, I question that the evidence presented supports that conclusion or the extrapolation that the changes in temperature and precipitation in the twentieth century played a significant role in the northern range willow decline. My demurrals with the evidence and logic presented here are as follows. 1. First, the evidence does not support the authors’ unqualified statement that the clipped willows “were not able” to increase phenolic and tannin content. It is true that neither species increased summer tannin content over the 3 years. Although summer phenolics decreased in S. bebbiana willow from 1993–95, the clipped false mountain plants increased slightly over the same period; all treatments in both species increased twice as much or more in 1996. Moreover, the intimation that 1996 may have been a more favorable year for the plants allowing increase in secondary metabolite production is based on the phenolics increase. In 5 out of 6 cases, tannins decreased between 1993 and 1996. No weather data were provided to demonstrate that in fact 1996 was a more favorable year. It has not been established that these two species do increase secondary compound levels under herbage removal. The clipped plants in this study behaved much like the unclipped controls. In their introductory comments, they cite Singer and Cates (1995) as reporting “tall productive willows growing within the upper reaches of the drainages where the habitat is more favorable for growth were browsed seven to nine times more than were suppressed willows at lower elevations” (emphasis added). This is not a result consistent with the hypothesis that willows in favorable habitat increase secondary chemical production under herbage removal and deter browsing. Moreover, as already commented, the preceding study measured lower tannin content of browsed willows outside exclosures than unbrowsed plants inside. 2. Despite the contentions of both studies, it is not clear that the exclosures are situated on unfavorable sites. The first study clipped shrubs in the Mammoth, Junction Butte, and “Lamar” (not specified whether East or

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West or both) exclosures. Since the Gardiner exclosures do not have willow, and the Blacktail exclosures were swept by the 1988 fires, the ones used in this study are almost certainly the same as those used in the first. As commented in Chapter 6, all of these exclosures have the water table at or near the surface. It is the reason why willows grow on the sites. Moreover, they are situated on upland areas with spring seeps, well away from streams that might have down-cut, or had reduced flows. 3. Beyond these uncertainties, there is no evidence that elk, hard-pressed for winter forage, will not use the willows with slightly higher tannin or phenolic content. There have been no observations of foraging behavior in the wild or captive feeding trials that would address this question. Willows are generally recognized to be among the most palatable browse species for browsing ungulates. And, the heavily browsed plants are not unprotected. The foliar tannin content of the heavily browsed, heightsuppressed plants in the Singer et al. (1994) study was 85% that of the tall, less-browsed, higher-elevation plants. 4. Finally, the same contrary evidence applies to this model as to the other climate-change hypotheses. Heavy impacts on willows were reported in 1915 and were well under way by the 1920s, barely into the centurylong climate change hypothesized to be the second- or third-order cause of willow decline, and well in advance of the 1930s drought. A considerable amount of research has been conducted on this question. And it has tended repeatedly to discount elk browsing as the major source of northern range change. Despite the lack of tangible support for the secondarycompound model, which was pointed out in the Wagner et al. (1995b) critique, park publications have continued to advocate its merit. Thus, “a likely answer to this question.” (Bishop et al. 1997:48). And: “the most promising theory” (Schullery 1997a:229). BEAVER DECLINE AND WILLOW HABITAT. Singer et al. (1998b) clipped current annual growth (CAG) shoots of willows inside 3 exclosures on the YNP northern range and 1 in Rocky Mountain National Park (RMNP). In each park, a third of studied plants were unclipped, a third were clipped at 50% of growth shoots, and 100% of shoots were clipped from a third. The authors measured plant height, sex, total number of stems, canopy diameters, basal diameters of shoots, and secondary metabolite production and estimated above-ground biomass. They also sampled CAG of willow plants in 15 9.3 m2 plots inside (unbrowsed) and outside (browsed) exclosures. Production increased, although to a declining degree, with increasing intensity of clipping in RMNP. Production decreased at all clipping levels (including 0) in YNP during the 2-year study period. Plant height increased at all clipping levels in RMNP during the study period but decreased at all levels in YNP (again including unclipped plants). Plants produced catkins at declining rates with increasing intensity of clipping in both parks but at higher levels in RMNP. Willows in RMNP produced higher levels of tissue tannins and phenolics at 50%

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clipping than at either 0% or 100%. But tannin and phenolic levels did not vary significantly with clipping level in YNP. In the 9.3 m2 plots with unbrowsed and browsed plants, those of Geyer willow (Salix geyeriana) had higher production, shoot length, and shoot weight in RMNP than plants in YNP. Mountain willow (S. monticola) in RMNP showed higher production and shoot length than any of the other species measured (data for three species are shown for each park). But production of Bebb willow (S. bebbiana) in YNP produced at higher levels than any of the other species except mountain. And unbrowsed false mountain willow (S. pseudomonticola) in YNP produced at higher levels than planeleaf willow (S. planifolia) in RMNP, although the reverse was true for browsed plants of the same species. Thus, only Geyer willow was common to both parks. Singer et al. (1998) concluded from these results that “growth conditions” for willows were more favorable in RMNP than in YNP because of higher effective precipitation, more beaver activity, more beaver dams in drainages, and “likely” higher water tables near stream sides in RMNP. They suggested further that a relatively larger beaver decline in YNP may have exceeded a threshold value for willow persistence and recruitment. They concluded that “elk herbivory alone does not influence willow persistence or recruitment as Wagner et al. (1995) concluded.” They misquoted Wagner et al., who commented in their paper, “we do agree that beaver decline could have been a cause of willow decline.” In their reply to Wagner et al. (1995b), Singer and Cates (1995) imputed assumptions and implications that the latter authors never made. Singer et al. (1998b) went on to state: In Yellowstone NP, a high density of elk were clearly contributing to the decline in willow . . . but a negative elk influence on those factors was not conclusively verified in Rocky Mountain NP where growth conditions were better. We agree that climate, beaver, large predators, and possibly floods and fires, in addition to elk, also influence willows [sic] abundance. . . . Some willow . . . declines should have been expected this century in both parks, regardless of the parks’ policies on elk, due to the long-term trends toward aridity. If climate changes toward aridity accelerated beaver declines, then the effects of aridity would have been exacerbated. The authors then propose an “ecosystem macrohypothesis” for YNP “to demonstrate the complexity of factors that influence plant communities” although their figure 2 to which the reader is referred is entitled “factors that might influence elk and willow populations.” The conceptual model in their figure 2 lists 23 interacting factors. No data were presented in the publication for over half of these, and several had not been previously mentioned. No direct effect of elk on willow is shown in the model, the only elk influence operating through a causal sequence of less aspen, beaver decline, riparian changes, and less willow. Changes in willow secondary compounds are not included in the model although three of the authors were among the authors of the two secondary compound papers just discussed.

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I will return to further consideration of this model later. For now I wish to focus on the beaver aspect and the inferences drawn from the evidence presented. First, the general inference from the evidence presented that RMNP is a more favorable environment for willows is problematic. The clipping studies investigated different willow species in the two parks. The studies of browsed and unbrowsed plants compared Geyer willow in both parks, but the other two species examined in RMNP were different from those measured in YNP, as mentioned. One cannot assume that the different species function similarly and can therefore be validly compared. Moreover the study sites were not comparable. Singer et al. (1998b) place the elk winter range in RMNP at 2,400–2,800 m; Houston (1982:6) reports the elevation of “the Lamar,” near the two YNP exclosures studied, at 1,973 m. The Junction Butte exclosure, the third one studied, is at some elevation lower than the Lamar. As Bishop et al. (1997) point out, willow is still abundant and robust at higher elevations in YNP, although this has to be attributed in part to the fact that these areas are above the elk winter range. As Singer et al. aver, precipitation is slightly higher at the RMNP site than at the YNP sites, but again possibly because of the differing elevations. As Keigley (2000) points out, the willow sites are subirrigated with water up to the surface at the YNP sites. Thus, it is not clear that the precipitation differential plays a significant role in any difference in favorability for willow growth. At best, one can only infer from the evidence that the conditions at the RMNP site might be more favorable for willow growth than those at the YNP sites. Secondly, the case for a beaver role in the inferred better growing conditions in RMNP does not follow from the evidence cited. Keigley (2000) cites a personal communication from D. R. Stevens, RMNP research biologist, who states that beaver have been absent from the vicinity of the RMNP exclosure since “at least” 1968. Similarly, Keigley points out that the YNP sites are at springs or seeps on hillsides and upland, considerable distances from any streams and not sites that would be occupied by beaver. Thus the sites in both parks provide no evidence that support inferences about the role of beaver in the comparative favorability of growing conditions of the two. In sum, the logical thread of the paper begins with (1) the inference that willow does not function as well in YNP as in RMNP, (2) that the difference is due to less favorable conditions in YNP, (3) that the condition differences are the result of differences in precipitation and beaver abundance, and (4) that the ubiquitous historical decline of willow in the YNP northern range was driven in part by declining growth conditions associated with climate change and a precipitous decline in beaver. Although I question whether the evidence presented supports any of these steps in the progression, I do consider, as did Wagner et al. (1995b), that beaver decline in YNP could have played some role in willow decline. In damming streams and forming ponds, often extensive ones, beaver increase the area of riparian zones. The result is an increase in willow habitat. As Singer et al. (1998b) comment, once the sites are abandoned and water levels drop, the bare, moist soils of the former ponds provide favorable habitat for willow

Influences on Riparian System Structure

seedling germination and establishment. As will be discussed, beaver that were abundant as recently as the 1920s on the northern range have largely disappeared. Kay (1990) calls them “ecologically extinct.” In all probability there has been a decline in riparian zone as a result and with it decline in willow habitat. But it is necessary to return to the question of ordinal positioning in the causal sequence. Although Singer et al. (1998b) do not show a direct effect of elk on willow in their conceptual model, park investigations (Singer et al. 1994; Bishop et al. 1997) now acknowledge that elk are “the proximate factor in the [willow] declines” (i.e., first-order factor). Decline of beaver-constructed willow habitat may also be a first-order factor. But the question needs to be asked why beaver declined so precipitously in the YNP northern range. As I will discuss later in this chapter, a number of authors have implicated elk removal of aspen and willow, both beaver food and building material, as the cause of beaver decline. Thus beaver emerge as a second-order factor affecting willow abundance by creating willow habitat, and elk serve as a fourth-order influence in the elkwillow-beaver–willow habitat–willow abundance causal sequence. In the Singer et al. (1994) terminology, elk are both a proximate and ultimate factor in northern range willow decline. Moreover, there is considerable question as to the relative importance of the beaver link vis-à-vis the elk direct effect. As commented, the first willow impacts were observed in 1914, and decline was well under way by the 1920s when there were still large numbers of beaver on the northern range. Kay (1994b) surveyed willow stands and beaver numbers on the lower reaches of Odell Creek and its tributaries in the U.S. Sheep Experiment Station, located in the Centennial Mountains 100 km west of YNP and within the GYE. The area has no wintering elk and only a few moose. Kay documented robust willow stands and nine beaver colonies. In 1996, Smith et al. (1997) found only one colony on the entire YNP northern range during an aerial survey. Hence willow and beaver persist in comparable areas not impacted by large elk numbers. Despite these uncertainties, and the substantial evidence of an overriding elk effect, one of the authors of the Singer et al. (1998b) paper was quoted on the February 10, 1999, Internet page of Salt Lake City’s Deseret News as saying the decline of the beaver population may be “the most important factor in the demise of willow [on the northern range]” (Bauman 1999). Interestingly, the same authors of the Singer et al. paper commented 4 years later: “Elk herbivory [in RMNP] appears to be the dominant force determining vegetation productivity in willow sites, but the effects may be exacerbated by lowered water tables” (Zeigenfuss et al. 2002). FIRE SUPPRESSION. As discussed in chapters 6 and 7, Houston (1973) concluded from examination of fire scars on 40 trees that the northern range had experienced a 20–25-year fire frequency prior to park establishment. Several authors (Houston 1973, 1982; Bishop et al. 1997; Singer et al. 1998b) have invoked fire suppression during park history as one cause of willow decline. These tend to be brief suggestions, usually with no evidence and little if any rationale

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on the complex of causal mechanisms by which the absence of fire might have contributed to willow decline, or conversely how prepark fire frequency might have maintained willow abundance. Houston commented (1982:124): The addition of big sagebrush canopy [from reduced fire frequency] on formerly open grasslands and an increase in density and distribution of forest types could have changed the hydrologic characteristics of smaller streams. . . (Houston 1973). Fire suppression could have altered the hydrologic characteristics of the smaller streams on the area and reduced stream cutting . . . consequently, the amount of gravel bars and other pioneer habitat suitable for colonization by willows and alder has declined. These speculative comments do not specify what hydrologic processes are implied. The reference to increase in woody vegetation perhaps implies increased run-off associated with a decline in herbaceous vegetation. But the second part of the quote implies a reduction in stream flow, presumably associated with reduced run-off from unspecified cause. Alternatively the reference to increased woody vegetation might imply increased evapotranspiration and consequent reduction in stream flow, an effect widely demonstrated in watershed studies (e.g., Bosch and Hewlett 1982). But the evidence of this and previous chapters has shown sharp reduction in such woody vegetation as sagebrush, aspen, and willow, which could well have reduced moisture loss to evapotranspiration. The effect would be to augment stream flow. Sagebrush has also been shown to accumulate drifting snow and shade it (see Blackburn et al. 1990; Welch 2004). The result of the latter is to delay spring snowmelt, reduce evaporation, and facilitate infiltration into the soil. Sagebrush removal can thus have a xerifying effect, although the net result must be balanced against reduction in evapotranspiration and effects of converting to herbaceous vegetation. All of this is speculative and has not been measured on the northern range. Consequently there is neither evidence nor stated rationale for how fire suppression might have operated through trends in woody vegetation to affect hydrologic processes and ultimately willow abundance. The second Houston quote suggests a shortage of germination sites for seedlings due to reduced stream flow, flooding, and formation of bare sand and gravel bars. Despain (1989) suggests a shortage of wet, mineral seedling sites that would be produced by more frequent fires. However Beschta (1999) compared sequential northern range aerial photographs of the Lamar River and Slough Creek from 1954 to 1991 and observed increased lateral stream movement that produced more meanders and increased area of gravel bars. The area for seedling germination and establishment appears to have increased. He attributes the change to the removal of woody, bank vegetation and loss of bank and channel stability. Rosgen (1993) had previously measured high levels of bank instability on the northern range and attributed it to removal of streambank vegetation.

Influences on Riparian System Structure

Some authors have pondered the role of fire in riparian plant community composition and succession. Kay (1990:157) challenged Houston’s (1982:131) statement that riparian vegetation succeeds to grassland and forest, but in fairness Houston explicitly referred to riparian succession outside the winter range, not the northern range. Bishop et al. (1997:47) observed that some riparian areas burned in the 1988 fires were converted to grasses and forbs. Chadde et al. (1988), identified eight willow community types on the northern range. All are underlain by sedge and/or grass understories. The authors did not discuss the role of fire, but commented that heavy browsing could eliminate the willows and convert the communities to stands of graminoid species. Finally, Norland et al. (1996) examined the effects on survival and growth patterns of willows in four northern range sites swept by the 1988 fires. Protein levels and digestibility of burned willows increased, as did shoot lengths and weights and leaf surface areas. And although primary production increased in half of the burned stands, it decreased in the other half. Shrub heights in all burned stands declined, apparently due to heavier browsing by elk attracted to the more nutritious plants. The authors commented, “Nearly all of the willow plants in other stands died or their growth was greatly suppressed by the fires.” In total there is no clear evidence of how fire interacts with willow on the northern range. Nor is there an ecological rationale for the causal mechanisms by which fire frequency might have perpetuated robust willow growth before park establishment or how fire suppression contributed to willow decline during the park years. There is an immense amount of information on the effects of ungulate use, especially domestic, on vegetation and stream-channel stability in the extensive riparian literature, but curiously very little on fire effects. This may be due to the moist nature of the type and relative invulnerability to burning. There are a few references to the stimulation of willow growth by fire (see Wolff 1978). But the response of the northern range willows to the 1988 fires was varied and did not generally follow the stereotype. At this point, there is no persuasive evidence or ecological rationale suggesting that fire suppression during park history played a significant role in willow decline during most of the twentieth century.

Causation Synthesis Complex ecological processes are seldom determined or influenced by single factors, and ecologists seeking an understanding of such phenomena appropriately investigate a range of factors that might be involved. But the science also recognizes that relevant factors vary in degree of influence and moves toward the general paradigm that a major part of the behavior of complex systems is governed by a small subset of the full array of interactions (Holling 1995). Ecology also operates on a premise of mechanism, recognizing that environmental factors affect phenomena in chains and networks of causation. Significant understanding of a phenomenon depends on a knowledge of the causal factors, their relative importance, and their ordinal positions in causal sequences.

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All observers now agree that there was major decline during the twentieth century in the area of the northern range occupied by full-statured willow shrubs. Estimates of decline range between 50% and 95% with the actual value probably closer to the larger number. Severely hedged plants and root suckers still occur, but there are no reliable estimates of the remaining number of plants irrespective of stature, and of the remaining area occupied by the species. Because root suckers commonly occur outside the exclosures, and many undoubtedly sprout from roots of full-statured plants inside the exclosures, it becomes a matter of definition whether they should be considered individual plants, much like the question of ramets in aspen clones. There clearly has been extensive plant mortality, with dead plants reported during the twentieth century, and stream banks previously occupied by dense willow growth now devoid of the species. Thus although no estimates are available, there almost certainly has been significant decline in the area occupied by the species. Park publications have tended to emphasize the reduction in plant height and numbers (see Singer et al. 2003), but no work has attempted to measure the proportion of riparian zone, originally cloaked with willows, from which the species have been totally eliminated (see figure10.1). By far the overriding bulk of evidence points to direct elk herbivory as the main first-order force driving the willow decline. Moose may have played a small role, but I have suggested as a first approximation that elk influence exceeds that of moose by 2 to 3 orders of magnitude. Insect outbreaks may have had a small effect, but there is no evidence suggesting that they were anything more than occasional localized phenomena. The photographic and anecdotal record points to robust willow growth in the park’s early years, which must have coexisted with its insect fauna. Three variants of a climate-change hypothesis have been proposed. All conjecture implicitly that climate changes during the 1900s reduced the quality of willow habitat which affects the species’ physiology and ultimately their survival and/or reproduction. One variant surmises that habitat decline has reduced plant viability directly, in which case habitat change has been the first-order factor and climate change second-order. The other 2 climate-change hypotheses acknowledge elk herbivory as the first-order factor contributing to willow decline. Climate change is hypothesized to operate sequentially through reduction in willow habitat quality, the plants’ ability to withstand browsing, and ultimately their susceptibility to elk herbivory. In these cases, climate change acts as a fourth-order factor causing changes in the herbivory rate. There is not enough evidence to accord the climate-change hypotheses a significant probability of being valid. Several realities cast doubt on all of them. 1. As discussed in chapter 6, park investigators have based climate-change hypotheses on Houston’s (1982) analysis of YNP weather records showing twentieth-century temperature increase and slight precipitation decrease and those of Balling et al. (1992) showing temperature increase

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and decline in summer precipitation. But authors of the NRC report (Klein et al. 2002) could not discern a century-long trend in the data, and concluded that the twentieth-century variations were slight and within the bounds of natural variability. Most recently, two independent analyses of twentieth-century climate trends in the northern Rockies (Kittel et al. 2002; Baldwin 2003) have inferred slight temperature rise and precipitation increase during the 1900s. Thus the changes have been slight and, if anything, toward more mesic conditions. 2. Willow impacts were observed as early as 1914. Decline was under way by the 1920s, barely into the minor century-long climate changes, if any by then. 3. Willows grow in many moist areas in the northern range where they are subirrigated and it is not clear how much, if at all, they are affected by slight climate change. The reality is that willows today maintain strong growth inside exclosures on the northern range and in favorable sites outside the park if not subjected to heavy ungulate use (figure 10.5). I also find no significant evidence supporting the fire-suppression hypothesis, or a clear ecological rationale for the nature of the interaction. In this case fire would be a first-order determinant. I do accord a tangible probability to the beaver-decline hypothesis, not on the basis of the evidence provided to date but on the general knowledge of beaver influence on stream hydrology. But this is clearly not the only (or even major) process producing willow decline: (1) Decline began while beaver were still numerous. (2) Decline has occurred around springs and seeps on hillsides and along small first-order trickles too small and major rivers like the Lamar and Yellowstone with reaches too large to attract beaver impoundment. Moreover, evidence indicates that the beaver decline was induced by elk removal of aspen and willow needed for food and building material. Elk herbivory then becomes a fourth-order determinant of willow decline through elimination of beaverconstructed willow habitat as well as its first-order browsing effects. In sum, the overwhelming evidence indicates that elk herbivory is the major determinant of willow decline, functioning both as first-order and higher-order determinants. A small number of other factors have had a relatively minor role by comparison. After a half-century of research, we surely understand the general structure of the phenomenon. This paradigm is essentially similar to the one posed by park investigators up to the 1970s. The other hypotheses have been put forth by more recent park personnel since enunciation of the natural-regulation hypothesis and adoption of the natural-regulation policy. Houston (1982) initially discounted elk herbivory declaring (p. 131), “changes in distribution of willow were mostly climatically determined.” As increasing numbers of observers have pointed to the decimated status of willow, obviously produced by heavy browsing, park personnel could no longer deny the effects. But the concessions are almost perfunctory nods. Thus Bishop

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et al (1997:47) comment, “Elk are usually the proximate factor in the decline of willow stands.” But two paragraphs later they state, “Since then [the 1930s] there has been little change in willow status, perhaps lending credence to the idea that climatic forces rather than ungulate browsing, may be the most important factor in determining the fate of the willows.”

COTTONWOODS Introduction Chadde et al. (1988), in their classification of northern range wetlands, list two community types dominated by two species of cottonwood trees: black cottonwood (Populus trichocarpa) and narrowleaf cottonwood (P. angustifolia). Although it is a very minor component of the northern range, I devote space here to P. angustifolia for a number of reasons. It provides an element of landscape and habitat diversity to the northern range ecosystem. It is part of the riparian subsystem, and the evidence indicates that it, like the rest of the subsystem, has been in decline. Recent research on the species gives evidence of correlations between trends in yet another component of the biota and those of the northern herd. The methodology in that research has promise for documenting the browsing history on other woody species in the northern range. As with the other vegetative types, early accounts described heavy browsing pressure on cottonwoods and decline. Reporting observations in the 1920s, Skinner (1928) commented, “Elk even stood on their hind legs to reach and pull down the lowest branches of . . . cottonwoods.” Kittams (1948), in his figures 5 and 6, compared a “dense grove” of cottonwoods in an 1896 photo of Soda Butte Creek with his 1948 retake of the site that showed “Only a few trees of the original cottonwood grove . . . are left.” He commented further: “A cottonwood grove between Mammoth and Gardiner . . . has virtually disappeared. Close field examination shows that reproduction is present but is mostly of bush form due to browsing.” Thus, as with the other components of the vegetation, concerns developed early among park biologists on the effects of elk browsing on cottonwoods and what that portended for the future of the species.

Browsing History and Trends in the Species In an innovative piece of research, Keigley (1997b, 1998) characterized the architecture of cottonwood plants in four categories, and associated these with browsing patterns: 1. He termed the architecture of a typical tree with nonforking trunk (figure 10.7, left) “uninterrupted growth type.” Such a plant was considered to have grown during its early life stages in the presence of light to moder-

Influences on Riparian System Structure

ate browsing, which allowed a central stem to escape browsing and grow beyond maximum browsing height (e.g., 2.5 m). 2. Chronic, intense browsing of a young plant kills the terminal leader and any lateral branches that turn upward and attempt to assume the function of a main stem. The plant is maintained in a shrubby form, and no stems escape to become a tree (figure 10.7, right). He called this “arrested-type” architecture and attributed it to intense browsing throughout the plant’s life. The plant’s lower portion is able to survive because it is protected from browsing in winter by snow, perhaps to a height of 50 cm. But any shoots that grow above that height during the growing season are browsed off in the following winter. 3. Keigley’s third architecture type, called “retrogressed,” was also produced by a two-stage browsing history. Early in the plant’s development, light to moderate browsing permits a stem to grow taller than arrest height. When browsing increases to intense levels, the stem protruding above the snow is killed. 4. A plant with hedged growth form, but a single stem that grows out of a thicket of browsed stems and reaches tree form beyond maximum browsing height is classified as “released-type architecture.” Such a plant (figure 10.7, right) is considered to have spent its early years under intense browsing, which maintained it in arrested or retrogressed form.

Figure 10.7 Cottonwood “trees” showing Keigley’s (1998) (left) uninterrupted growth type; arrested growth type (hedged plant in foreground of right); and released type architecture (small tree with crooked trunk in the midground to the right of the skis and behind the arrested plants on right). Photos by Charles E. Kay, March 16 and March 10, 1980, respectively.

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But at some point the browsing pressure eases to light to moderate, and a stem escapes and is able to grow to tree height. Keigley reconstructed the dates of browsing intensity experienced by each plant through dendrochronology. He cored the plants at 0.3 m above ground, which he considered reflected the ages and therefore dates of establishment of each. Uninterrupted-type plants must have become established and lived for a sufficient period of early life to grow beyond the browse zone during a period of light to moderate browsing. That period could be light to moderate determined by ascertaining the establishment date of the plant. Similarly, ascertaining the establishment date of an arrested-type plant indicated that it had endured intense browsing from that date to the present. The browsing history of a released-type plant could be reconstructed by coring at 0.3 m to determine its date of establishment and coring the released stem at 1.0 m to ascertain its age and in turn the date on which it was released. Thus an early period of intense browsing followed by a period of light to moderate use could be dated. Establishing the browsing history of a retrogressed-type plant was less precise. Its date of establishment could be determined by the 0.3 m core, and the length of life of the escaped stem could be determined by cross-section. But it was not possible to determine exactly when during the life of the plant the stem escaped or when it was killed. If inspection of the stem showed it to have died in the recent past, the return to intense browsing must have been relatively recent, and the stem’s life as a tree could be placed approximately within the plant’s lifetime. If the stem had obviously been dead for some period, this implied that the return to intense grazing had occurred some time in the past. The younger the plant, the smaller the error in making these approximations. Keigley’s method makes it possible to establish the age (except for trees with heart rot) of all of the plants and thus their dates of establishment. By aging stems that had grown to tree stature, it is possible to establish dates on which this stem growth began and the early years during which that growth carried the stem beyond browsing vulnerability. By implication, these dates coincided with lightto-moderate browsing. In the case of uninterrupted-type growth, plant- and stem-establishment dates are the same. In the case of released-type plants, a measurable period occurs between plant establishment and growth-to–1m dates. This period is assumed to coincide with intense browsing followed by light-to-moderate browsing, enabling the stem to escape and assume tree form. In the case of arrested-type growth, only the plant establishment date is ascertained with the entire lifetime of the plant occurring under intense browsing. Plant establishment is made possible by winter snow that enables it to begin life and grow to a height of winter snow depth before it is vulnerable to winter browsing. Keigley (1998) examined plants in three adjacent stands near the confluence of Soda Butte Creek and the Lamar River in 1992. He chose this area because of the diversity of age classes not found elsewhere on the northern range where

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cottonwood plants are either tall, advanced-age trees of uninterrupted growth form (figure 10.7, left), or heavily hedged, younger plants of arrested-type architecture (figure 10.7, right). His results were as follows: 1. Stand 1: 14 “very large” trees of uninterrupted-type architecture, up to 20 m high, and averaging 1 m dbh. All had heart rot, making it impossible to age them precisely. He counted 140–150 rings in the core of one tree before reaching the rotten center and concluded that its establishment “preceded the 1840s–1850s.” Because all 14 trees were of a comparable size, he reasoned that all 14 were established in approximately this period. 2. Stand 2: 45 trees all of uninterrupted-growth type, up to 15 m in height, averaging 30–50 cm dbh, “about the same size.” Most had heart rot, but he was able to age four and determined the establishment dates in 1877, 1892, 1893, and 1894. Keigley concluded that “all 45 trees were likely established within this approximate span of years.” 3. Stand 3: 53 trees in all four architectural types, and approximately 3,000 stunted plants of the arrested and retrogressed-type growth form. Keigley was able to core 51 of the 53 trees successfully at the 0.3 and 1.0 m heights. The approximate stem establishment dates of the trees in the 3 stands can be arrayed chronologically as follows: <1840s–50 ~1877–94 1894–34 1934–51 1952–62 1963–74 1974–92

14 45 1 33 0 17 0

These results are generalized to the entire northern range with some uncertainty, given the small number of trees aged in the first 59 examined, the consequent approximate age and date assignments, the moderate sample size in total, and the fact that the tree stands occurred in one localized area. Given the consistencies between the distribution of these dates and the trends in the elk population, and to some degree the Romme et al. and Ripple and Larsen trends in aspen trees, the chronological parallels must be pointed out. The first 14 trees probably began life prior to park establishment when, as inferred in previous discussions, elk in the area were present at low densities. The uninterrupted growth of the plants implies a period of light-to-moderate browsing up to the point when they were no longer vulnerable. The 45 cottonwoods in Keigley’s stand 2 were established in about the same period as a major fraction of the aspens aged by Romme et al. (1995). Again one must infer that these 45 trees existed in their early years during a period of light to moderate browsing. And in total, the establishment of 59 trees during a period

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from before the 1840s to the 1880s or early ’90s is consistent with a low-density elk population up to the latter part of this period. By the 1890s, the herd had reached high numbers, which continued until the major herd reduction in 1962. Keigley’s cottonwood establishment dates show only 1 tree established between about 1894 and 1934. Establishment of the 33 trees between 1934 and 1951 is the one part of the data set that does not coincide with the elk population trends. Whether this was a group of plants that avoided browsing by some means, perhaps through changes in the channels of the streams during this period or through changes in elk movement or distribution patterns, is impossible to say. Elk numbers were high throughout the period. However, the years 1952–62, also a period of high elk numbers until the final year, resumed the patterns of the first 34 years of the century. No plants escaped to tree form. Thus, except for the 1934–51 period, only 1 tree established during the first 61 years of the 1900s, a pattern also observed by Romme et al. (1995) and Ripple and Larsen (2000) for aspen. Seventeen trees became established in the period 1963–74, thus experiencing light to moderate grazing. This was the period of low elk numbers produced by the herd reduction. Romme et al. and Ripple and Larsen did not observe any aspen trees established during this period. But I cited limited evidence of aspen response in chapter 6. Keigley (1998) found no trees established after 1974, by which time the northern herd had returned to high numbers, and despite the fact that he observed (3,000 arrested-type plants in stand 3 that could have sent stems up to tree stature if released from browsing. He considered the larger trees in stand 1 to be at or near their maximum longevity. With continued high elk numbers, he predicted that cottonwood will be eliminated from the northern range. The few remaining old trees on the northern range (figure 10.7a) are in sharp contrast to the often dense stands of mixed-age cottonwoods along the Yellowstone River bottom north of the park in the Paradise Valley.

THE RIPARIAN FAUNA Beaver Beaver is one of the northern range species about which a great deal his been written. But there are few hard data on the species in the area. Most of the discourse has been based on anecdotal information and on implications of changes in other components of the system that are assumed to affect beaver. As with other components of the northern range, there was an initial paradigm posed by earlier park investigators that was changed to a current interpretation advocated by post–natural regulation park personnel. I will review the early park perspective, the current park position and the evidence invoked to support it,

Influences on Riparian System Structure

and then propose my own interpretation of the evidence and the trends in the northern range population. The current park position on beaver is largely set forth in Schullery and Whittlesey (1992), Bishop et al. (1997), and Schullery (1997b). These authors (see Bishop et al. 1997) concluded from historical accounts that beaver were “abundant,” “common,” and “distributed in appropriate habitats in Yellowstone National Park in the middle and late 1800s.” Kay (1990:165) cited Osborne Russell (1965), who spent “several days” to nearly 3 weeks each in 1835, 1836, and 1837 trapping beaver in northern range streams. And park Superintendent Norris (1880:613) commented that trappers took “hundreds, if not thousands” of beaver skins from the park each year of his superintendency in the late 1870s. Norris stopped beaver trapping in the park in 1883. Bishop et al. (1997:96) concluded that by this time “it seems probable that the beaver population was reduced significantly . . . beaver numbers may have been kept low by continued poaching well after sanctioned trapping terminated in about 1883. Beaver may not have been free of this trapping until the early 1890s.” The authors then referred to “a sudden growth of beaver numbers beginning about the turn of the century, as the beaver population responded to improved protection and the increase in available food” (p. 95). This surge in aspen escapement was the 1870–90 aspen cohort inferred by Romme at al. (1995) and discussed in chapter 6. Bishop et al. (1997) conclude that beaver reached an ephemeral high density in the 1920s, and then began a decline that coincided with the 1930s drought. Today beaver are scarce on the northern range. Smith et al. (1997) found only one colony on the northern range in the course of a parkwide aerial survey. Bishop et al. (1997) generally dismiss the suggestion, begun in the 1930s by early biologists, that the decline and current low numbers are primarily the result of aspen and willow elimination—resources needed for food and building material—by a large elk herd. Rather, they associate the decline with a drying climate that they assumed made aspen more vulnerable to elk browsing. One can accept parts of this scenario. But it dismisses out of hand the views of some of the earlier biologists. It also bases much of the argument on conjecture, unproven hypotheses (e.g., changing susceptibility to browsing), and a scenario that is not supported by the evidence (the 1870–90 aspen surge). The end result is an ill-supported paradigm that, like so many other natural-regulation arguments, strives to discount the effects of elk on the northern range. In the process, part of the argument is inconsistent with another tenet of the park’s northern range paradigm. I agree with the park authors that the evidence indicates beaver abundance prior to and at the time of park establishment. Beaver were numerous throughout Western North America at the time of European arrival, and a number of early reports indicate this to have been the case in the Yellowstone area (Schullery and Whittlesey 1992).

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Trapping in the early years of park history may have reduced the population somewhat. But it is a matter of conjecture as to whether beaver numbers were “reduced significantly” or “kept low.” Superintendent Anderson (1892) commented in 1892, “Of late I have seen evidence of great numbers of beaver.” He commented again in 1893 on beaver abundance, and in 1897 he remarked that beaver were “numerous” and listed 18 streams and all creeks that drain into the Gardner that had beaver (Anderson 1897). Meagher and Houston (1998:134) remarked on the basis of comparative photos that “beaver had constructed an extensive series of ponds on the floodplain of Soda Butte Creek by the time of the 1900 view . . . the ponds apparently remained until the early 1920s.” The evidence does indicate high densities of beaver in the 1920s. Skinner (1927:176) reported beaver on “practically every stream” as did Bailey (1930). Warren (1926) began a study in the early 1920s because there was concern that the large beaver population was cutting down too many aspen trees. He studied a small area on the northern range near Tower Junction and counted 12 colonies with 236 beaver. Bishop et al. (1997) infer a major population increase to achieve the 1920s densities (“a sudden growth of beaver numbers . . . the beaver population irruption”), and attribute this to protection from trapping, predator control, and the “20- to 25-year surge in aspen escapement.” There may have been some increase as a result of the protection from trapping and predator control, but the extent is uncertain. As already noted, there is evidence that beaver were still abundant in the 1890s. And, as discussed in chapter 6, the evidence fails to support the inference of an 1870–90 surge in aspen escapement. The evidence of beaver abundance at the time of park establishment is also inconsistent with the park position that elk were numerous at the time (see chapter 3). The large beaver numbers must have been supported by a vegetation that supplied ample food and building material, namely, aspen and willow. Yet Bishop et al. (1997:52) comment, “The work of Romme et al. (1995) and Warren (1926) suggests that in the early 1800s, aspen were typically unable to grow to tree height.” The large beaver numbers at park establishment have implications about the nature of the vegetation. In a real sense, as Kay (1990) points out, high ungulate numbers, especially elk, and high beaver numbers are mutually exclusive. The northern range is the prime laboratory pointing to this conclusion. All observers agree that beaver began declining soon after the high populations of the 1920s. While commenting on their abundance in the northern range in the 1920s, Bailey (1930:114) observed, “On account of the extensive herds of elk, which keep down aspen growth over most of the Park, there is not an abundant food supply for a large number of beavers .” Five years later, Wright and Thompson (1935) wrote that beaver (as well as other species) “have been more or less sacrificed to the maintenance of the elk herd.” It is generally agreed that the northern range beaver population had declined sharply by the 1950s. Jonas (1955) conducted a survey of beaver numbers over the entire park, and he found no animals in the area in which Warren (1926) had counted 12 colo-

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nies and 236 beaver in the 1920s. Kittams (1959) commented on the disappearance of beaver and attributed it to elk impacts on willow and aspen. Houston (1982:183) remarked that the available evidence does not support earlier interpretations of competitive exclusion of beaver by elk. And Bishop et al. (1997:95–96) chided Chase (1986) for claiming that Jonas (1955:166) had attributed beaver decline “solely” to an overpopulation of elk, pointing out that Jonas had suggested a number of factors. Indeed he had, but the authors neglected to cite Jonas’s statement: “The unfortunate food situation at the time of this survey was a result more from the overpopulation of elk than from any other single cause. . . . Generally, conditions favoring beaver activities were considerably better in those regions where elk browsing was not severe.” Further discounting the role of elk in beaver decline, Bishop et al. (1997) also cite Houston’s (1982) comment that the post-1900 beaver increase occurred during 1900–1910 when elk numbers were extremely high. But beaver populations must respond not to elk numbers per se but to the condition of the vegetation. We have seen in this and previous chapters that vegetation impacts were first reported in the second decade of the twentieth century. But the photographic record shows a considerable amount of aspen and willow vegetation still surviving in the 1920s. Several decades of high elk numbers were required to reduce the vegetation to its present condition and inability to support a significant beaver population. Though finding only one beaver colony on the northern range, Smith et al. (1997) counted 48 other colonies in the park during the course of their aerial survey. All were associated with willow vegetation. As already commented, Kay (1994b) observed nine beaver colonies along Odell Creek west of the park and robust willow stands along the creek. In summary, beaver were numerous in streams of the Yellowstone area from the early 1800s through the early years of park history. This abundance implies a vegetation capable of supporting the populations and in turn is one more piece of circumstantial evidence pointing to low elk numbers in those years. Beaver remained numerous in the northern range through at least the 1920s. There may have been some variation in the population level through this period, perhaps declining somewhat in the 1880s, and reaching highest numbers in the 1920s. The population then began the decline that has taken it to the point of being, in Kay’s (1990) term, ecologically extinct on the northern range today. All observers agree, if somewhat tangentially, that the decline was induced by elimination of food and building material. Houston (1982:182–83) attributes the food decline to reduction in aspen by fire suppression, drought, and browsing; Bishop et al. (1997:96) attribute it to a drying climate that made aspen more vulnerable to elk browsing. Thus the answer resolves back to the first-order question of what factor(s) produced aspen and willow decline. I concluded in chapter 6 and here that although there may have been some influence of other first-order factors, by far the most influential is the high elk density. Because there was no demographic, behavioral, and physiological research on beaver during the decline period, all evidence is circumstantial and inferential, and any conclusion remains a hypothesis.

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One is left with judging the probability of this and other hypotheses. I consider this the more probable because it accords with all of the evidence and is consistent with the changes in the other components of the system. Whatever the correct hypothesis, the decline of beaver from the riparian zone of the northern range streams constitutes a profound reduction in diversity. And that loss goes beyond just the loss of the species. It includes the biota and ecological processes associated with the expanded riparian zone created by the beaver. Kay (1990) speculates on lowered water tables and down-cut streambeds in areas previously occupied by beaver colonies and ponds. Schullery and Whittesey’s (1992) historical accounts of river otters (Lutra canadensis) and flocks of waterfowl, infrequently seen in numbers on the northern range today, tend to support this image of more extensive riparian and aquatic zones in the northern range during the park’s early history, and what that contributed to the species, habitat, and landscape diversity of the system.

Riparian Ungulates Whitetailed Deer Whitetailed deer (Odocoileus virginianus) in the western United States are essentially a riparian species, perhaps as a result of evolutionary partitioning of the habitat and terrain resources with the upland mule deer. Commenting on their former distribution in YNP, Skinner (1929) observed, “They were by nature inhabitants of the bushy areas along rivers and streams.” Schullery and Whittlesey’s (1992) historic accounts include occasional reports of the species attesting to its presence in the park in the 1800s. But it apparently was never abundant. Earliest reports of their numbers estimated the population at around 100 animals about the turn of the century, roughly similar to the contemporary number of bighorn sheep. Park publications have commented on a population increase up to this point, but no basis has been provided for this assumption. The accounts of the population continued into the early 1920s, but apparently with declining numbers until the last observations were made in the winter of 1923–24 (Skinner 1929; Bailey 1930; Murie 1940). The population occurred largely along the Gardner River north to the park’s boundary. Judging by the small numbers and higher elevations in the park and the fact that it is more abundant north of the park along the lower-elevation stretches of the Yellowstone River, I agree with Bishop et al. (1997) that the species was probably at the upper edge of its elevational range in the park. Murie (1940:9) observed that the whitetailed deer range along the Gardner River was “heavily browsed” by elk and that “the vanishing of a suitable winter habitat for this brush-loving . . . species was probably the basic cause of its disappearance.” That view was generally held by park biologists until questioned by Houston (1982) on the grounds that about 100 whitetails occurred in the park around 1900, when elk numbers were quite high. Here again, pressure on putative competitors is not a function of elk numbers per se but the condition of the vegetation at any given point in time which elk created over some preceding period. We have seen in previous

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chapters that some of the first impacts on vegetation were observed in the 1910s and 1920s. The whitetails occurred largely in the lower-elevation northwestern portion of the northern range where, as evidence presented in previous chapters has shown, the northern herd tends to concentrate and exert heaviest pressure. It is probably here that the first significant impacts occurred. Whitetailed deer occur commonly to the north of the park in southwestern Montana. On the basis of their in-depth study of the species along the lower Yellowstone River, Dusek et al. (1989) commented: Relative density of [white-tailed deer along the lower Yellowstone River] . . . varied directly with abundance of riparian forest and shrub cover: a strong linear relationship suggested a continuum of density distribution of deer with that of riparian cover . . . Any significant alteration of river flow and dynamics that resulted in a reduction of riparian tree and shrub cover would ultimately result in a reduction of base densities of white-tailed deer. As with beaver, it will never be possible to prove conclusively that elk competitively excluded whitetails from the northern range. The existing evidence is circumstantial and inferential. But based on (1) what is known about white-tailed deer ecology in the region, (2) the common ecological tendency of species to be most sensitive to environmental changes at the fringes of their ranges, and (3) the known changes that occurred in the park at the time, the hypothesis that elk competitively excluded whitetails from the northern range has to be accorded significant probability. It is one more aspect of the profound reduction in ungulate diversity, discussed in chapter 9, by a burgeoning elk population. An occasional whitetail has been seen in the park in recent years. But these tend to be itinerant individuals.

Moose Moose are well known to prefer riparian habitat and willow for food in western North America. But they are wide-ranging animals that can and do occupy mature forest at higher elevations. As reported in the willow discussion, northern range moose counts starting in 1968–69 varied from 37–100 over the ensuing 20 years but declined to an average of 19 in the 1990s. Thus they are now uncommon, if not rare, denizens of the northern range. The few animals that do occur in the area tend to winter in higher-elevation coniferous forest (Maxwell 1994; Tyers and Irby 1995). In essence the species is no longer a significant occupant of the northern range riparian zone. Kay (1990) hypothesized that this is another example of competitive exclusion by the northern herd elk population, and Singer et al. (1998) now agree that this may be the case.

Riparian Avifauna Riparian zones in the northern Rockies, along with aspen woodland, are known to have the highest avian diversity of any other major vegetation types in the region. Hansen (1997) observed that cottonwood, aspen, and willow have roughly

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twice the avian species richness and abundance as in lodgepole pine, sage, or grassland types in the GYE. Bottoroff (1974) recorded the highest mean densities and the second highest mean number of avian species in cottonwood habitat of the nine uncultivated, nonurban vegetation types he surveyed in northern Colorado. Douglas et al. (1992), working in the vicinity of the Centennial Mountains west of YNP, found that avian species vary with subtle variations in riparian vegetation structure. Differences in riparian vegetation structure were strongly correlated with proximity to moisture, and presence or absence of beaver play a significant role (second-order in this case) in determining the nature of the avifauna. Medin and Clary (1990) had earlier observed exactly this effect in the same state. Douglas et al. concluded that “subtle changes in riparian areas (e.g., from cattle grazing, timber harvest, drainage, etc.) may have severe impacts on the bird community.” Jackson’s (1992, 1993) work in and around YNP provided strong supporting evidence for the Douglas et al. hypothesis. Jackson studied avian densities and species composition on eight sites falling along a gradient of browsing intensities as measured by the percentage of willow shrubs severely browsed. Three of these areas were within the northern range, and these were the most severely browsed. Three areas were in the park but outside the northern range, and two were outside the park, one of them as far away as the Red Rock Lakes National Wildlife Refuge west of Yellowstone. Jackson observed 26 species in all of the areas during her 2 years of summer observations (her table 4). The number of species among the eight areas was inversely correlated with browsing intensity: 18–19 observed in the two areas outside the park and the 3 outside the northern range, and 7, 9, and 9 on the three northern range sites. Number of species and population densities were roughly (but not linearly) correlated among the sites. Avian densities averaged ~25 birds/ha at browsing intensities from 0–65% of willows severely browsed (her figure 6). But this dropped off sharply at higher browsing intensities, with only ~5–8 birds/ha in the most severely browsed sites. Jackson selected stream reaches that still had some willow shrubs, even if heavily hedged, for her “severely browsed” sites. Undoubtedly those reaches on the northern range which have completely lost their shrub strips (see figure 10.1b) have also lost the shrub-inhabiting species entirely. There is little doubt that the ubiquitous decline of full-statured riparian, woody vegetation, including cottonwoods, has sharply reduced the riparian avifauna on the northern range.

Invertebrate Fauna There probably are several hundred species of invertebrates in intact riparian zones in the Yellowstone region. Debinski (1996) found 9 species of butterflies alone in moist areas of the northern range. On the basis of work in Glacier National Park and the GYE (Debinski and Brussard 1994; Debinski et al. 1999), she and co-workers observed that habitats that had high avian diversity also tended to have high butterfly diversity. Thus one might surmise that the heavy elk impacts on northern range riparian zones that Jackson (1992) has shown to affect avian diversity also affect butterfly diversity.

Influences on Riparian System Structure

Except for surveys of species distributions, little research has addressed the ecology of the invertebrates on the northern range. However, Debinski’s (1994, 1996) study of a single butterfly species, the Yellowstone checkerspot (Euphydryas gillettii), gives insight into the delicate sensitivities with which a single species maintains a tenuous hold on existence, even in intact riparian zones. The checkerspot is declining in the region and may need investigation to determine whether it should be listed as a threatened or endangered species under the Endangered Species Act. The female butterflies lay their eggs only on the growing tips of a single, shrubby species, the black twinberry (Lonicera involucrata). Twinberry occurs on the northern range, although it is apparently not common. Kay (1990:144) observed it in the Lamar East exclosure but not in the Lamar West. The shrub is obviously browsed by elk. Kay measured mean plant heights at 91.8 and 34.0 cm (p < 0.01) inside and outside the exclosures, respectively. Plant cover was 4.2% and 2.0%, respectively, though the means were not significantly different. However, Debinski (personal communication, October 4, 1998) commented that the outside plants had probably been browsed down to the point where the tips were no longer functional as ovoposition sites. Debinski (1994) observed that the checkerspot exists in metapopulations, the subpopulations occurring in patches of riparian habitat. The species’ mobility is also quite low. Repopulation of patches in which the species has been extinguished, through dispersal from occupied areas, depends on proximity of the latter. Hence wide distribution of habitat patches, or any factor, such as elk browsing, which eliminates patches (see figure 10.1b), and certainly Lonicera involucrata, threatens persistence of the species in the area. This is just one small example of what is doubtless a heavily affected invertebrate fauna in the altered, northern range riparian zone.

THE RIPARIAN ECOLOGICAL SYSTEM Like the aspen and conifer woodlands and the sagebrush-steppe, the riparian type functioned in early park history as a discrete system with its own plant and animal species interacting among themselves through a myriad of processes. The six species of willow (Chadde et al. 1988) discussed collectively as “willow,” and the two species of cottonwoods are but a fraction of the total riparian plant species. The Chadde et al. list of five other shrubby riparian species, Debinski’s (1994) mention of Lonicera involucrata, and Chadde’s (1989) citation of historic accounts reporting “large quantities of chokecherries” (Prunus virginiana) along streams add seven woody species to a list that is not exhaustive. I know of no concerted effort to develop a full list of herbaceous species and their interactions. Hence, we have only fragmentary knowledge of the northern range riparian vegetation and its interactions. Likewise our knowledge of the fauna is restricted to the few species that have attracted research attention. Little more can be said about the invertebrates,

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and I am aware of no concerted attention to the small mammals. Geier and Best (1980) observed clear preferences for riparian habitat types among Midwestern small mammals. Thus, we do not begin to have a thorough insight into the original structure and function of the northern range riparian ecosystem. By the same token we have only a superficial and fragmentary understanding, based on trends in the few components that have been studied, of the changes that must have occurred in the organisms and processes of the system. The riparian type has also been one of the several subsystems that interact among themselves and collectively form the northern range ecosystem. It has provided food, habitat, and migration corridors for the more wide-ranging animal species that use several of the subsystems. The cottonwoods must provide perches, nest sites, and observation posts for raptors, ravens (Corvus corax), and magpies (Pica pica) that forage along the streams and over the uplands. The shrubs obviously provide browse for the more wide-ranging ungulates. Kay (1997) points out that riparian zones are essential security cover, travel lanes, and foraging sites for grizzly bear (Ursus arctos horribilis) in several areas of Montana. Thus the riparian type has added to the species, habitat, and landscape diversity of the northern range. The riparian type has also functioned as a buffer on hydrologic processes, not only within the northern range but also on the interaction between the higher elevations of the park and the northern range. The surface hydrologic network of the northern range is a dendritic conveyance system, collecting water both from higher elevations of the park, and the intervening ridges and plateaus of the northern range itself and carrying it downslope to the major streams. The associated mesic or aquatic vegetation—in some cases, narrow strips in drainage ways and along stream banks, in some cases broad floodplains, depending on the topography—has served as a buffer on surface and stream bank erosion, thereby providing a measure of stability on surface soils and stream morphology. In turn the streams and their aquatic ecosystems have been provided a measure of stability. Gregory et al. (1991) pointed out that riparian zones are the interface between terrestrial and aquatic ecosystems. The plants’ root zone is an important determinant of the character and functioning of nutrients passed on to streams in the groundwater. The zone is also a source of abundant and diverse allochthonous food resources passed into streams. This is an important determinant of the abundance and diversity of aquatic macroinvertebrates, and they in turn affect the abundance and diversity of fish life. In total, I concur with Gregory et al. (1991) and Chadde (1989) that the ecological importance of riparian zones far exceeds their minor proportion of the landscape. Clearly this entire complex of organisms, processes, and interactions has been profoundly reduced or eliminated, perhaps under the influence of several factors, but largely and overwhelmingly by massive reduction of riparian vegetation by elk browsing. The full extent and complexity of the changes will never be known, although there is little doubt that they have occurred.

Influences on Ecosystem Function I: Erosion FREDERIC H. WAGNER AND RICHARD B. KEIGLEY

11 Everything in the world is strange and marvelous to well-open eyes. —José Ortega y Gassett

SURFACE EROSION Background On the basis of his northern range studies between December 1, 1928, and April 1, 1932, W. M. Rush (1932:64) commented: “On more than half of the range sheet erosion has taken place to a depth of one to two inches which in soil of this type is a serious loss. In some few minor areas gully erosion has begun.” Three years later, Wright and Thompson (1935) stated: “Erosion is underway. . . . It is now stripped naked, and the fertile topsoil is wasting away. . . . With the original character of the northern range becoming more obscure each year, the urgency of this [proposed] investigation cannot be overemphasized.” These authors’ comments were not based on well-designed studies of soil erosion but rather on their observations of pediciled plants, exposed tree roots, and surface gullies. And they coincided with the early paradigm of large increases in the northern herd following park establishment. Houston (1982) stated, “The available evidence does not support interpretations of widespread or accelerated erosion in the area.” But the park had still not conducted formal research on erosion at the time he wrote, and he based his views on inspection of early photographs and his own 1970s retakes of the same sites. He concluded that any early photos suggesting erosion were taken 215

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during the 1930s drought when surface soils were poorly covered with vegetation, or on highly erosive, steep slopes that were inherently subject to erosion. In fact, Rush’s observations were made before the 1930s drought and over “more than half of the range” including the Gardiner area, the Blacktail plateau, and the Lamar River valley. Houston’s inferences were in accord with his own paradigm of large elk numbers on the northern range prior to and through most years of park existence. The park has conducted only a few short-term studies of erosion since Houston’s book. These have generally concluded that the evidence did not indicate accelerated erosion since park establishment. Park public-information documents (e.g., Bishop et al. 1997) have generally reiterated these conclusions. Implicit in the Rush and the Wright and Thompson comments are certain complexities of the erosion issue that have been largely overlooked. Rush commented on the loss of 1–2 in (2.5–5.1 cm) of topsoil by 1931; Wright and Thompson expressed concern over whether the “original character” of the northern range could ever be determined. The questions then arise as to the chronology and extent of erosion, if it occurred. By 1931, the northern range had been used by a vastly increased elk population for some 45 years, which impacted a soil that had formed over preceding centuries or millennia without heavy ungulate use. The initial surface erosion rate could have been substantial. Any topsoil loss could have left behind a gravel residue surface that armored the soil and reduced current erosion rates. Any early losses would be unmeasurable. Low contemporary losses, if any, would understate earlier loss rates. As Wright and Thompson feared, it might never be possible to understand the original character of the northern range. Some of these questions can never be answered. But with reanalysis of data from previous studies and evidence to be presented in the next chapter, we infer that surface erosion has accelerated during park history up through the 1900s.

Recent Studies of Sheet Erosion Prevalence of Surface Rock Pediments Shovic (1996) conducted a soil survey over a 4,065-ha study area located north and northwest of Mammoth Hot Springs. He classified sites by (a) soil productivity, (b) erosion potential, and (c) site productivity. He described no specific evidence of elk-related erosion, but noted: “High wildlife usage undoubtedly contributed to average vegetative coverage, vegetative species dominance, and erosion status. Its importance, however is confounded by inherent ecosystem properties and past disturbances.” If sheet erosion occurred during the twentieth century, rock fragments should have been concentrated at the soil surface once the fine-grained component was removed. Data from Shovic (1996) indicate that over 62% of his study area, more than half of the soil surface was covered by rock fragments. A more extensive soil survey by park personnel suggests that surface rock fragments are common over an extensive area of the northern range (figure 11.1). Data from

Influences on Ecosystem Function I: Erosion

Figure 11.1 Across much of the northern range, the soil surface is covered by coarse material, gravel-sized and larger. Although frost heaving could have played a role, early park observers spoke of the erosional loss of one to two inches of topsoil. Coin in the photograph is a quarter.

this study provided us by Ann Rodman of the Yellowstone Center for Resources show soil-surface rock fragments (>2 mm diameter) covering 0–65% of the surface, with vegetation, litter, and soil clasts >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.

Influences on Ecosystem Function I: Erosion

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

Influences on Ecosystem Function I: Erosion

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.

Influences on Ecosystem Function I: Erosion

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.

Influences on Ecosystem Function I: Erosion

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

Influences on Ecosystem Function I: Erosion

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 <.01 .03 —

270 210 390 250

20 20 — —

9.4 8.9 — —

430 260 — —

— 5.8 — —

— 8.0 — —

— 29 — —

— 58 — —

— .08 — —

— .05 — —

600 850 — —

Slide-deposit Lakes B.S. M.R. V. Buck

15/8/74 30/7/74 23/8/77 8/65

Kettle Lakes F.I. B.T. S.C. B.F. aData

7/66 19/8/77 — —

from U.S. Fish and Wildlife Service files now maintained by the National Park Service, except for conductivity, which is from Engstrom et al. (1991). T is temperature Celsius; AlkT is total alkalinity summing carbonate and bicarbonate; KN is Kjeldahl nitrogen; NN is nitrate nitrogen. All concentration in ppm. Lake name abbreviations are explained in figure 12.1.

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

sion of small trees, as indicated by limited dendrochronological data. That comment also showed that the portion of land in the northern range experiencing increased transport of silt and organic matter to lakes, as represented by percent elevation range associated with accelerating bulk-sediment rate, rose from less than 40% prior to park establishment to over 80% by the turn of the century. It then declined, reaching its earlier low value approximately in the late 1950s, only to increase sharply above 70% after the late 1960s. Although the accuracy of this interpretation is certainly subject to refinement, I do not agree with Engstrom et al. (1994) that it is an “arbitrary treatment of the data” to draw attention to secular trends that approximately mimic elk population numbers while differing markedly from climatological trends shown in their figure 2. Moreover, figure 1 in Hamilton (1994a) clearly shows that 6 of the 8 lake records exhibit an increasing secular trend in bulk-sediment accumulation rate since 1872, possibly affirming the elk population hypothesis. Foster and Big Trumpeter are the only exceptions. There is no way of knowing whether or not the interpreted change in Big Trumpeter from Wright (1987) to Engstrom et al. (1991) resulted from analysis of a different core or from some other factor. The first interpretation showed an increasing secular trend. In the reanalysis below, there was not sufficient time to calculate rate data for each sample in each core. Instead, with the exception of allogenic silica, I rely only on the sample dating nearest to 1872 for comparison with the most recent core increments, and I again recommend that Engstrom et al. perform the rigorous reanalysis of the data.

Allogenic Silica Engstrom et al. (1994) commented that I “ignored geochemical evidence and disregarded the fact that sediments are derived from multiple sources.” These were the parameters I had hoped they might reassess, having access to the original numerical data that would simplify such analysis. To assess the deposition rate of allogenic silica, thought to be a good measure of silt, I applied their silica data in their figures 10, 11, and 12 (expressed in mg/g dry weight of total sediment) to the total sediment data in their figure 6. Sampling increments did not match between the 2 figures (suggesting lumping to increase sample size for certain determinations), so I interpolated to calculate values of allogenic silica. My use of numerical data derived from their plots indicates that 6 of the 8 ponds show an increase in siltation rate since 1872 by a mean factor of 2.1. When Big Trumpeter and Floating Island Lakes are excluded (as described), all but Foster Lake show increases. The mean increase is a factor of 2.5. To the extent that the beginning and ending values represent the change since park establishment, the data once again appear to support the elk population hypothesis. Figure 12.2 illustrates the calculated allogenic silica deposition rates in the cores with lakes ordered by elevation. The figure shows that the changes between beginning and ending values discussed derive from relatively monotonic, secular increases in deposition rates with episodes of rapid increase and decrease

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Figure 12.2 Plots of allogenic silica deposition rates over time. Rates are calculated from graphed data in Engstrom et al. (1991) and essentially represent annual silt deposition as discussed in the text. Lakes are arranged in order of increasing elevation from left to right. Linear regressions are shown by solid lines. R2s from left to right are 0.80, 0.86, 0.70, 0.50, 0.47, 0.64, 0.25, and 0.78.

superimposed. At 6 of the 8 lakes, some systematic factor has been responsible for increased siltation since establishment of the park. Four of the curves exhibit large increases in the most recent decades, and five show declines in rate during the time of elk reduction in the 1950s and 1960s. Clearly there is no support here for the report’s conclusions that: “Increased elk populations during the last two decades have apparently not resulted in increased erosion of the lake catchments . . . [and] The case for accelerated erosion in the northern range resulting from ungulate grazing is not supported by the stratigraphic studies.” The causes of these trends were not adequately addressed by Engstrom et al. (1991), nor was their existence clearly reported (see also figure 12.3).

Organic Sedimentation Using the Engstrom et al. (1991) plots for loss-on-ignition values, I applied combustible organic concentration (percent) in the 1872 and the surface layers to their plots and arranged them by elevation (figure 12.4). All of the lakes showed an increase since park establishment, on average 20–30%. The deposition rate of organics also increased (figure 12.5) by a mean factor of 2.4. When Big Trumpeter and Floating Island data were excluded, the increase grew to a mean factor of 2.8. Thus Engstrom et al. may have neglected to look carefully at this parameter. Over the time interval since establishment, organic deposition-rate change appears partially to meet the requirement of the fourth element of the hypothesis.

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

Figure 12.3 1872 (dashed lines, open symbols) and ca. 1985 (solid lines, closed symbols) values of allogenic silica deposition rates at lakes versus elevation, calculated from data in Engstrom et al. (1991). The greatest increase since park establishment is seen at lakes in landslide terrain (excepting Foster Lake). Relatively little change occurred at glacial kettle lakes.

This result is consistent with my observations in late winter and early spring at ponds near Tower Junction. Lakes and ponds are frozen over in winter, and with the approach of warm weather elk tracks and numerous melt pits containing fecal pellets are seen on the ice surface. The tracks indicate that the elk were feeding on sparse, exposed stems of shrubs and sedge. After the ice melts, the pond bottom near shore is littered with thousands of soft fecal pellets which roll back and forth, disintegrating under the influence of wave action.

Biogenic Silica A parameter whose increases Engstrom et al. (1991) suggested would be indicative of a eutrophic state, biogenic silica concentration increased in all cores from the level in 1872, with the exception of Buck Lake (figure 12.6). Deposition rate of biogenic silica increased by a mean factor of 3.5 over the interval, with a decrease only at Big Trumpeter Lake (figure 12.7). Element four of the hypothesis is strongly supported by this secular trend.

Phosphorus All eight cores exhibit a significant, relatively monotonic increase in phosphorus concentration in the upper portions (figure 12.8). There is no discussion on the extent to which this represents normal phosphorus dynamics at the anoxic

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ELK EFFECTS ON ECOSYSTEM STRUCTURE AND FUNCTION

Figure 12.4 1872 and ca. 1985 values of organics concentrations (percent) at lakes versus elevation, from Engstrom et al. (1991). A majority of lakes record an increase. In general, kettle-lake (intermediate-elevation) sediments contain higher concentrations than those in slide-block terrain. A weaker elevation trend is also evident.

boundary or whether it might represent some increase in phosphorus flux. Iron and organic carbon, both probably more mobile than phosphorus (Baccini 1985), exhibit correspondingly less enhancement in the upper 10 cm of each core. Furthermore, the apparent mean increase in phosphorus concentration is by a factor of 2.5, and in most lakes the greatest apparent increase has occurred since the 1960s. This increase exceeds that reported by Baccini (1985) in mesotropic and eutrophic subalpine lakes affected by urban and agricultural activity. The mean change in deposition rate (again, all cores record an increase) from 1872 to ca. 1980 (excluding the surface layer) is by a factor of 4.8 (figure 12.9). The phosphorus data suggest the possibility of real secular increases in the rate of phosphorus deposition, as might be attributed to leaching from bone, antler, and fecal material in the catchments. As useful as the phosphorus (and iron) data might be in assessing redox in bottom water, the limited coverage it is given by Engstrom et al. (1991) is puzzling. The only reference I can find occurs in their Geochemistry section in the results portion for the first lake (Foster) as follows: “authigenic P = 1 mg/g (Fe

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

Figure 12.5 1872 and ca. 1985 values of organics deposition rates at lakes versus elevation, calculated from data in Engstrom et al. (1991). All lakes record an increase since establishment of the park.

Figure 12.6 1872 and ca. 1985 values of biogenic silica concentrations versus elevation, from Engstrom et al. (1991). All lakes except Buck show an increase since park establishment. A negative relation to elevation is evident over the entire range in both time periods. 241

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Figure 12.7 1872 and ca. 1985 values of biogenic silica deposition rates versus elevation, calculated from data in Engstrom et al. (1991). Seven of the eight lakes show an increase since park establishment. A negative relation to elevation below ~1,870 m is evident.

Figure 12.8 1872 and ca. 1980 values of phosphorus concentrations versus elevation, from data in Engstrom et al. (1991). All lakes record an increase since park establishment, with greatest departures at higher elevations.

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

Figure 12.9 1872 and ca. 1980 values of phosphorus accretion rates versus elevation, calculated from data in Engstrom et al. (1991). All rates have increased since park establishment, with greatest enhancement recorded at lakes in landslide terrain. As with biogenic silica, there is a negative correlation with elevation below ~1,870 m in the early record.

and P are generally similar among all sites).” It is not mentioned again, except in passing in the diatom results.

Diatom Stratigraphy and Paleoecology Engstrom et al. (1991) reported taxonomy and relative abundance of diatom taxa from only five of the eight lakes. Modern pond-water chemistry was partially reported for six ponds, because chemistry, among other factors, affects diatom species abundance. Relative abundance of Stephanodiscus species was used as an index of eutrophication following, for example, Brugam (1982, 1983) who found a generally positive relationship between S. hantzschii abundance and phosphorus concentration in 48 alkaline ponds in Minnesota and used that observation to characterize conditions in the lake predating Kirschner Marsh there. Indeed, the Engstrom et al. (1991) data show a positive relationship between phosphorus concentration and total Stephanodiscus abundance in the surface layer at the four lakes for which data are available; but a stronger negative relationship (not mentioned) is evident at all five lakes between conductivity and Stephanodiscus abundance (figure 12.10).

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Figure 12.10 Abundance (percent) of Stephanodiscus sp. in uppermost samples from Engstrom et al. (1991) versus conductance and phosphorus concentration (see table 12.1). Stephanodiscus abundance appears positively related to phosphorus, in accordance with the model used to assess trophic change, but limitation of the model is suggested by inhibition of Stephanodiscus in waters of higher conductance. Lakes are Buck (B), Foster (F), Slough (S), Middle Rainbow (MR), and Big Slide (BS).

This suggests that to be a reliable indicator of paleochemistry in these lakes, diatom stratigraphy should have been augmented with a more complete understanding of species tolerance to water chemistry in some detail and additional modern water chemistry data including pH, dissolved silica, and nitrogen. Stephanodiscus abundance following the alkaline water, phosphoruslimited model suggests that 2 of the 5 lakes (Middle Rainbow and Slough) have become more eutrophic since the nineteenth century; two (Big Slide and Foster) have become more oligotrophic; and 1 (Buck) has remained about the same. Two lakes (Big Slide and Middle Rainbow) show large increases in relative abundance of another taxon, indicating eutrophic conditions (Cyclostephanos invisitatus). Thus, my evaluation of the plots of diatom data using the Brugam (1983) model indicates that 3 of the 5 lakes studied show evidence of increased eutrophication. The finding for Foster Lake is puzzling because Stephanodiscus declined over time and was absent in the surface layers there, yet the lake water phosphorus concentration is reported as 36 ppb, a level well above the value Brugam (1982) found limiting for S. hantzschii. The result for Big Slide Lake was even more problematic because its phosphorus concentration was reported as 143 ppb. Their discussion does not mention this dilemma, the phosphorus budget, or mobility in any of these lake systems, pH, the role of elk in introducing phosphorus to the system, or the role of organic carbon in the sediments mobilizing stored phosphorus (see phosphorus section).

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

In their discussion, Engstrom et al. interpret the bewildering diatom data as follows: All five sites . . . were mesotrophic to eutrophic prior to the establishment of the Park . . . it is clear that none of the lakes has been significantly enriched in comparison with their “natural” pre-Park disturbance conditions, and that the magnitude of any enrichment by ungulate populations must be small in comparison with natural nutrient yields from the catchment. This stands in sharp contrast to the organic carbon, phosphorus, and biogenic silica increases since 1872 noted previously and to the increase in eutrophilic taxa at three of the five lakes discussed in this section. In their conclusions, Engstrom et al. acknowledged: An increase in lake productivity during the last few decades is suggested for Buck, Slough Creek, and possibly Big Trumpeter, Floating Island, Big Slide, and Middle Rainbow lakes by sedimentological increases in biogenic components and accumulation rates. The diatom profiles support this interpretation for the four [sic] lakes that were studied (Buck, Slough Creek, Big Slide, and Middle Rainbow). Tophic [sic] conditions today, however are not much different from what they were before the Park was established. Aside from overlooking the Foster Lake diatom results, their figures 15 and 16 (including Foster) do not accord with the conclusion as stated. Their figures 15 and 16 show clearly that diatom-inferred trophic conditions for all five lakes immediately postestablishment were indicative of “lower nutrients,” “low nutrients,” “lower nutrients?,” or “low nutrients?” whereas the uppermost sediments representing the most recent several decades were characterized in 4 cases out of 5 by “higher nutrients?,” “high nutrients,” or “high nutrients?” To the extent that the figures show prepark trophic interpretations, which is unclear, a high trophic state is weakly implied at only one (“Benthos Increase Slightly” between prepark and 1980 at Big Slide Lake)! Moreover, the figure captions call these tentative interpretations. The discussion identifies a shift at all sites from planktonic to epiphytic and benthic diatoms during the late twentieth century, and because the shifts were asynchronous, lowering of lake levels was ruled out. Their analysis failed to take into account the relative importance of groundwater versus surface-water recharge, and as will be shown the greater shifts occurred at kettle lakes (groundwater dominated) while smaller shifts were associated with slide lakes (surfacewater dominated). In summary, I believe that their interpretation of diatoms in these northern range lakes, though supporting increased eutrophication since establishment, sheds very little light on details of the history of elk in the catchments.

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POLLEN CHANGE Pollen types were identified and counted in all cores, and Engstrom et al. (1991) concluded that: The pollen evidence for ungulate grazing on trees and shrubs is weak: most sites do show a slight decrease in pollen profile for one or another hardwood tree or shrub (willow, aspen, alder, birch), but the dates for the decrease are highly variable from site to site, and in any case dry climatic conditions provide an alternative explanation. The bulk presentation of pollen data at different scales made it difficult to clearly focus on browse species preferred by elk as a test of the elk population hypothesis. I therefore regressed pollen profiles for willow, alder, and birch over time. Aspen was not considered because preservation of its pollen is still considered problematic (Holloway 1984; Davis and Botkin 1985). Similarly, cottonwood pollen was excluded from this reanalysis because mature trees that produce pollen often have bark too thick to be palatable to elk. As shown by Wright (1987), willow pollen, the species most preferred by elk, declined markedly between 1920 and 1935 at Foster Lake and remained at historic lows after 1947. This fits almost exactly the photographic documentation and chronologic interpretation of willow loss from elk browsing presented by Kittams (1948) for the valley floor a mile from the lake catchment. Indeed, 5 of the 8 profiles—Big Slide, Slough Creek, Buffalo Ford, Floating Island, Foster—exhibit secular declines in willow abundance since park establishment (and earlier). R2s for these five tests are, respectively, 0.67, 0.22, 0.22, 0.39, and 0.25. A sixth profile, for Big Trumpeter, also suggests a decline since just after the earliest part of that record. As will be discussed, the record at Middle Rainbow Lake must reflect early grazing by cattle followed by partial recovery possibly associated with artificial perturbation. Following that history, the Middle Rainbow Lake record also shows a decline. This leaves only Buck Lake, the lake farthest removed from the northern winter range, with willow pollen showing no trend over time. Pollen of less palatable alder similarly declines at five lakes since establishment, with a sixth (Big Trumpeter) showing a possible decline since shortly after the earliest sediment sampled. The least preferred browse species, birch (Richard Keigley, personal communication), exhibits a decline at only 3 of the 8 lakes. The principal representative on the northern range is water birch (Betula fontinalis). A decline in pollen representing trees and shrubs requiring proximity to the water table, where that is the principal variable, would most likely support a climatic interpretation. However, with the data showing a strong relationship to palatability, elk number rather than climate emerges as a more likely controlling factor. The pervasive decline in willow is probably not attributable to beaver, whose populations have declined drastically owing to the increasing difficulty in finding alternates to their preferred food sources.

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

The Engstrom et al. pollen discussion contains a number of speculative statements that are not supported by evidence. Nearly a quarter apply to discussion of aspen. One, concerning Populus (where it is not apparent whether aspen, cottonwood, or both are implied) states: “Its decline probably reflects the restriction of Populus with the onset of warmer and drier conditions and development of better soils.” Another example follows discussion of the claim that willow pollen (in addition to birch, alder, and aspen) failed to show recovery “at any of the sites” during the 1960s when elk numbers were low. “It seems likely that the decline in shrubs is a response to episodic drought in the northern range and shifts in the patterns of animal activity.” In my judgment, the pollen data appear to show increase (albeit modest) in willow, alder, and birch pollen in the 1950s and 1960s at half the sites for each species. More to the point, the report neither claims to test the climate-change hypothesis nor accomplishes that goal. The insertion of such unsupported statements within the discussion can mislead the casual reader to the view that there is a body of knowledge that clearly indicates a reduced importance of elk and elevates the importance of climate in causing decline in the browse species when there is no such body of evidence. The assessment of weed pollen might have more effectively tested the elk impact hypothesis if it had focused on exotic species presently abundant in heavily browsed lake catchments. These would include timothy, beggar’s lice, toadflax, and Canada thistle.

ANALYSIS OF PRE-1872 DATA The greatest difficulty in assessing pre-establishment core data comes from the dating uncertainty inherent in the Pb-210 method. Age uncertainty implied in figure 5 of Engstrom et al. (1991) in immediate prepark sample increments ranges between ±15 and ±112 years. For earliest prepark samples, the uncertainty ranges between ±32 years and ±125 years. In addition to placing doubt on prepark sedimentation rates and rate changes, these inherent activity-counting errors make absolute dating of events represented by pollen and diatom counts quite unreliable. Engstrom et al. (1991) acknowledged this limitation reservedly, with one notable exception in presenting the results of the Middle Rainbow Lake core: “However, dating precision, particularly before 1900, is very poor . . . and sedimentation rates are in effect unknown.” As shown in figure 12.2, allogenic silica deposition rates increased from deepest portions of the cores to the time of park establishment as indicated by Pb-210 dates at 6 of the 8 lakes. Those were the same sites at which the increases persisted up to the 1987 layer. Estimates of the size of the northern elk herd before establishment and shortly thereafter made by Keigley and Wagner (1998; see chapter 3), suggest that elk numbers were low on the northern range before 1872. They report evidence that after park establishment elk were increasingly occupying “safe” habitat

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ELK EFFECTS ON ECOSYSTEM STRUCTURE AND FUNCTION

at higher elevations in the northern range in spite of deep snow cover to avoid hunting pressure caused by early settlers in the Paradise Valley along the Yellowstone River north of the park. This early history may agree with evidence of early fluctuations in pollen and indications of trophic disturbance offered by diatom data in contrast to the Engstrom et al. report’s conclusion that, for example, “it is clear that none of the lakes has been significantly enriched in comparison with their ‘natural’ pre-Park disturbance conditions, and that the magnitude of any enrichment by ungulate populations must be small in comparison with natural nutrient yields from the catchment.”

SUPPLEMENTARY INFORMATION ON LAKE SETTINGS AND HISTORY Middle Rainbow Lake I wish to supplement the Engstrom et al. (1991) setting for this lake in several important respects based on my fieldwork and literature search conducted in 1998. The catchment is formed along a northwest-striking depression within a large landslide block, and Rocky Mountain juniper and Douglas fir grow sparingly on the slopes of the recent outlet. Two of these trees, of about 100 years in age, have curvature reflecting slope instability that would have tended to raise the outlet during the lives of the destabilized trees, perhaps apace with outlet incision. Moreover, there is an artificial berm, now incised approximately 1 m, at the outlet. Two other artificial berms block old swales along the north shore. A lichen trim line, about 80 cm above water level, is evident on relatively stable boulders on the northeast shore, near a large patch of houndstongue (Cynoglossum officinale), an exotic. The trim line has been transcended, mostly by Xanthoria elegans, with less ardent lichen species absent. Field relations suggest that the trim line corresponds in elevation to an artificially high stand produced by damming. When I asked long-term resident Raymond Stermitz, whose family had grazed cattle there prior to 1932, if he knew any reason why the lake would have been dammed, he suggested the possibility of irrigation. The significance of that idea was immediately clear, because the northernmost berm occupies a low divide that faces downslope to the northeast. Damming would have raised the lake level exactly enough to permit diversion of water from the Landslide Creek drainage onto a slope leading to otherwise dry pasturage below. Middle Rainbow is fed in part by upper Landslide Creek, and an emergent delta is evident at the inlet. A possibly older delta is several meters higher, at a spring line evidently associated with coarse deltaic deposits. The actual inlets do not match the single inlet shown in figure 3 of Engstrom et al. (1991). The lake is also fed intermittently by the discharge of Upper Rainbow Lake. The outlet of the upper lake, which appears free of human disturbance, is (in 1998) rapidly incising at two knick points developed in clay-rich slide deposits heavily tracked by elk. Headwalls of the knicks are 1 m or more in height.

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

Although Middle Rainbow lies within the original 1872 area of Yellowstone, it is close enough to the 1932 addition (where the Stermitz ranch was located) that the area was used by nearby ranchers for grazing cattle. An artificial stock tank is still evident, constructed half a kilometer northeast, where it submerged and killed juniper trees. Lower Rainbow Lake, into which discharge from Middle Rainbow flows, was a source of ice for use in Gardiner early in the century (R. Stermitz, personal communication), explaining the road that ran to the site. Trout were also stocked in the lower lake in the 1930s, and it was a popular fishing lake until planting ceased in the early 1950s and the population declined ( J. Stermitz, personal communication). Finally, the slide block and other terrain within the catchment of Middle Rainbow Lake are of heterogeneous composition (various Eocene sedimentary and Tertiary igneous lithologies in sizable areas of outcrop) owing to the postglacial age of the landslide and steep slopes to the southwest (Anonymous 1972; Pierce 1973). For this reason, it would be quite reasonable to expect some change of elemental ratios over time in a sediment core from this lake owing to catchment adjustments whether related to elk or not. The sediment deposition rate peak between ca. 1900 and 1920 in figure 6 of Engstrom et al. (1991) might be attributed to the construction of the dams at the outlet. This date is compatible with the time when other dams nearby were built by a rancher named Chadburn (R. Stermitz, personal communication).

Big Slide Lake Supplementing the description of the setting given by Engstrom et al. (1991), I observe that Big Slide Lake occupies a depression formed in a landslide block developed in weathered and disturbed deposits of the Tertiary Lost Creek Tuff (Anonymous 1972). The outlet of this lake is now being rapidly incised. A beaver dam constructed of mud, Typha stalks, driftwood, and freshly gnawed rabbit brush (Chrysothamnus sp.) has raised the pond level some 60 cm above the notched outlet channel. Yet the lichen trim line on relatively stable boulders along the east shore lies approximately 40–46 cm above the 1998 water level. The trim line, separating Xanthoria and another lichen pioneer species from a thick, diverse lichen growth above, corresponds approximately to the level of a welldeveloped elevated terrace evident along the west and southwest shore. Live lichens on one stable boulder were inundated, suggesting the recency of the latest beaver activity, and the east shore exhibited evidence of new slumping in steep areas. The incised outlet is partly stabilized by a dense growth of Typha, heavily browsed wild rose (Rosa woodsii), and boulder armor. The lichen trim line is less distinct than those observed at closed-basin ponds near Junction Butte, where I used air photos (Hamilton 1994b) to associate them with high stands of the ponds between 1969 and 1971. For this reason, I suspect that the elevated shoreline at Big Slide Lake either dates earlier than that or reflects a fluctuating high stand in response to varying beaver activity. Beaver are notorious pioneers, and they will try out an area only to abandon it if conditions

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are not satisfactory. In view of the present absence of mature cottonwood or willow, and the few water birch, I expect that their future residency at Big Slide Lake may be brief. Amphipods (presumably Hyalella sp.) were present in 1998, having survived the interval from 1936 to 1952 when 66,800 rainbow trout were stocked in the lake (Pierce 1987). Trout no longer inhabit the lake. The U.S. Fish and Wildlife Service reported (Anonymous 1974) that pond algae were “dominated by Aphanozomenon, which inhibits all others.” This alga is reported to precipitate calcite (with phosphorus) in naturally eutrophic lakes (Murphy et al. 1983). It is instructive to speculate on the decline of calcium in the deep-water core from this lake at the time rainbow trout were introduced (fish were not mentioned by Engstrom et al. 1991), as well as the large increase in biogenic silica after extirpation of trout. The signs of the old road (unknown to Engstrom et al. 1991) descending to the lake from the unpaved Gardiner-Mammoth road above were still visible. The road was pioneered with minimal construction to provide easier access for anglers in about 1936 (W. S. Chapman, personal communication). This disturbance does not appear to be recorded by elemental ratio shifts as required by Engstrom et al. in the case of elk disturbance, but I observe an increase in silica deposition rate over this interval in their data. The road disturbance is far from theoretical, because I was a passenger in a truck on that dusty road in 1952. The road was abandoned after the trout died out shortly after the last year of stocking. Apparently spawning habitat was lacking. Exotic toadflax and houndstongue were abundant along the shoreline in May 1998.

Floating Island Lake This is a closed basin pond, situated alongside the heavily traveled north portion of the Grand Loop Road. Eocene volcanic bedrock outcrops nearby, but landslide structures are not present. The lake appears to occupy a depression in glacial deposits that thinly mantle the bedrock. It is unlikely that road fill blocked an early outlet, though along the north shore the lake laps against fill. Lichens on boulders extend to the high water line (April 30, 1998), somewhat unexpectedly following a low-snowpack winter. Because there is no culvert, I believe that lake level is controlled by a highly permeable zone in glacial till (or possibly talus from an adjacent Eocene tuff cliff) under the roadbed. Engstrom et al. (1991) emphasized the effect of road construction on allogenic sedimentation in this lake, suggesting that the observed increase in deposition in the late 1930s was a result of road dust and downplaying any role of elk or climate. The history of road construction at Floating Island Lake is more complex than might be inferred from reading Engstrom et al. The road was first constructed there in 1931 (W. S. Chapman, personal communication), and in 1932 following construction of a bridge over Tower Creek, travel heretofore mainly to and from mines in Cooke City became increasingly popular with park visitors (Culpin 1994).

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

A bituminous pavement was applied to this section in or just before 1936. This history makes it difficult to explain how the new road would increase deposition of total sediment detectible in the middle of the lake between 1910 and 1931 (their figure 6) while allogenic silica concentration actually dropped between 1925 and 1940 (their figure 11). The increase in potassium/silica ratio, which they used to associate road building with increased sedimentation, could also be construed only as a continuation of the secular trend, according to their data, that began at the time of park establishment and leveled off in 1940. Floating Island Lake was probably low during the drought of the 1930s, and when precipitation returned to normal the rising lake encountered the new roadbed increasing biogenic silica production while reducing carbonates. The Fish and Wildlife Service (Anonymous 1966) reported the amphipod Hyalella sp. at Floating Island Lake.

Big Trumpeter Lake Big Trumpeter Lake is located on a hummocky, glacial till surface pocked with kettle depressions near Junction Butte. It appears to occupy a kettle whose rim has been breached on the west. It is fed partly by an intermittent stream from the southeast but primarily by groundwater. The shoreline in 1998 resembled a wadi, in that it was heavily trafficked by elk and bison especially in areas along the north and west shores where sodium bicarbonate efflorescences on the muddy shore were being used by the ungulates. Game trails descending to the lake showed development of rills, with headward erosion evidenced by knick points of decimeter scale. A clear lichen trim line 15–23 cm above the water surface during moderate outlet discharge suggested a relatively stable outlet. The outlet itself was bounded by gentle slopes with no indication of recent incision. The pond into which Big Trumpeter drains, however, discharges through an outlet that is undergoing active incision. Situated at a major game trail, only the cobble armor is restraining a rapid lowering of pond level. There Xanthoria exclusively extends to within 5 cm of the pond level, with a better trim line located at 20 cm. Apparently this outlet incision occurred several decades earlier. The level of Big Trumpeter since 1954 (Hamilton 1994b) has responded partly to recharge dependent on previous winter precipitation but mostly to precipitation and runoff of the year of record being lowest and without discharge (during intervals represented by air photos) in fall 1988. The 1988 low stand was probably 1–2 m below the discharging stage, based on the bathymetry shown by Engstrom et al. (1991).

Foster Lake Foster Lake occupies a depression formed in landslide terrain that was derived from failure of steep Eocene volcanic outcrops, with some involvement of glacial valley fill that contains diverse lithologies. It is a closed-basin lake, appearing to

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discharge in the subsurface (evidenced by lower elevation seeps nearby) only through permeable clastics that may underlie the silty slide deposits. On September 5, 1998, I observed a distinct lichen trim line on stable Eocene agglomerate rock outcrop along the east shore at an elevation of 25–46 cm above the water. The only possibility of an early abandoned outlet (showing no evidence of incision) is located at the southwest end of the lake 2.7 m above the lake. Field evidence suggests the possibility of blockage of the outlet by a relatively recent debris flow. Vegetation at the shore included Douglas fir, spruce, wild rose, sedge (Carex sp.), occasional heavily browsed willow, timothy, bulrush (Scirpus sp.), thistle (Cirsium sp.), and common juniper. Chara and water-milfoil were unusually abundant in the lake, and Canada geese were present. Heavily browsed aspens were located in the upper part of the catchment near the source of a dry tributary channel. In 1931 2,500 rainbow trout were stocked in Foster Lake. Reportedly, none survived, so stocking was discontinued (Pierce 1987). The catchment of this lake lies within an area currently closed at times owing to wolf winter habitat. In the 1930s it was similarly a popular area for coyotes (W. S. Chapman, personal communication). This may derive from the abundance of prey, as indicated by numerous burrows in the silty landslide deposits. The level of Foster Lake has responded to precipitation timing and amount since 1954 (Hamilton 1994b). Limited records show low stands in fall 1954 and 1971 and a higher stand in fall 1962. Lake stage appears to depend more strongly on rainfall and runoff during the year of record and may be less groundwaterdependent than the closed-basin kettle ponds at Junction Butte. The most unusual feature of Foster Lake, in comparison with the others I have visited, was the thick growth of submerged aquatic vegetation. It appears that it could greatly retard transport of sediment to the deep-water portion of the lake.

Results of Lake Setting and Water Chemistry Synthesis Engstrom et al. (1991) did not relate lake settings clearly to the terrain of the park. Figures 12.3 through 12.8 show for each lake the change in major parameters since establishment, both in terms of concentration (excluding allogenic silica) and deposition rate, as functions of elevation. The figures make it immediately apparent that this exercise could have been carried out just as clearly by plotting those changes as functions of distance up the Yellowstone River and thence up the Soda Butte Creek valley (see figure 12.1). The figures illustrate that there is a significant difference between the lowest and highest pairs of ponds on the one hand and the four at intermediate elevations on the other. The characteristic in common among those farthest up and farthest down valley is their occurrence in slide deposits (Pierce 1973; Prostka et al. 1975) associated with Eocene volcanic terrain (Anonymous 1972). The characteristic

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

in common among the other four is their occurrence in kettle depressions in glacial till (figure 12.1), which is dominated by a granitic lithology (Anonymous 1972). Recognition of these morphological, lithological, and hydrological characteristics makes it easier to explain why certain changes differ in one set of lakes from those in the other. The kettle lakes all contain twofold higher concentrations of organics in their sediments than those of the slide lakes. This pattern combines with an interesting elevational trend, with lakes at higher elevations having higher organic sediment concentrations. On the other hand, increase in deposition rate of organics since park establishment has been much greater in the slide lakes, particularly those at lower elevations. The distribution of organics also helps explain that of phosphorus in the upper sediments. Phosphorus accretion rate (deposition rate being an inappropriate descriptor owing to some loss rate from the surface layer due to reducing conditions) in the 1980s was much higher in the slide lakes than in the kettle lakes, owing probably to the enhanced organic carbon content in the kettle lakes, producing stronger reducing conditions. A weak elevational increase in phosphorus concentration also appears in the 1872 and 1980s sediments, but 1872 accretion rate of phosphorus is slightly enhanced at the lower elevation lakes and greatly increased at the slide lakes in about 1980. Biogenic silica concentrations in 1872 correlate well with elevation, with higher concentrations at lower elevations. Increased concentrations in the 1980s follow the same trend. Deposition rate however shows an elevation trend only below an elevation of 1,870 m approximately, where it is greatly enhanced in the 1980s layer. Allogenic silica (silt) deposition rates in relation to elevation at establishment are almost an exact replica of rates for organics, with greater rates at slide lakes and proportionally greater increases in the 1980s in slide block terrain, especially at lower elevations. The greatest departures of all rate parameters from establishment values are at the slide ponds. The kettle ponds appear more complacent over time, even though they record increases as well. The difference in water chemistry between slide lakes and kettle lakes (table 12.1) strongly suggests that the latter are more strongly dependent on groundwater recharge, having the higher conductivities and alkalinities commonly associated with groundwater in northern Yellowstone (Kharaka et al. 1991). This water-chemistry characteristic may explain why Engstrom et al. (1991) noted that “The magnitude of the increase (in epiphytic and benthic diatoms . . . relative to plankton . . . ) is quite small in Buck, Big Slide, and Middle Rainbow lakes (surface-water dominated). . . . The shifts are more pronounced, however, in Foster and Slough Creek lakes.” Slough Creek Lake was the only kettle lake investigated for diatoms; Foster, owing to its closed basin, may have a chemistry significantly influenced by groundwater (table 12.1). As Engstrom et al. recognize, the onset of this change in the late nineteenth century with a variable history among the different lakes makes climatic drying,

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as a means of enhancing alkalinity or conductivity, seem an unlikely candidate. One could even speculate that the geothermal plumbing beneath the northern range contributed increased volumes of high-conductivity deep water to shallower aquifers at various times in the past. I suspect, however, that more needs to be learned about response of diatom flora to total water chemistry before any cause, including elk numbers, can be ruled in or out.

OVERALL ANALYSIS Data Trends Trends both suggested and clearly demonstrated by reevaluation of the Engstrom et al. (1991) data fall into several categories. The 1872–1980s starting and ending values of rate variables all suggest increased deposition rate for parameters selected by the investigators to “isolate elk as the sole agent of environmental change.” At the selected sites, siltation rate, represented by allogenic silica, increased by a mean factor of 2.1 ± 1.5. Organic carbon deposition rate, representing manuring, increased by a mean factor of 3.6 ± 3.2. Biogenic silica deposition rate, indicative of trophic state, increased by a mean factor of 3.6 ± 1.9. Phosphorus accretion rate, partly an indicator of fecal, bone, and antler contribution to the lakes contributing to eutrophy, increased by a factor of 2.6 ± 1.4 between 1872 and ca. 1980 (surface samples excluded to avoid most of the phosphorus that would later reenter the water column). Trends in allogenic sedimentation rate and bulk sedimentation rate, using data for all samples in each core, showed relatively monotonic increase up to the present, with superimposed episodes of smaller-scale increase or decrease in rates. These trends were attributed by the authors primarily to processes other than elk-caused erosion, because few of the changes in elemental-ratio values seemed to fit their understanding of elk behavior and sediment transport. The only elemental ratios selected for their evaluation (P/Fe, Fe/Mn, and K2O/SiO2) certainly exhibit secular trends, but the authors do very little to show how these particular ratios would be diagnostic in view of the given lithologies and differing degree of representation and exposure of soils of contrasting chemical signature in the various lake catchments. At Big Slide Lake, the existence of an unpaved access road used to travel to and from the lake between 1931 and about 1952, clearly equivalent to a major and prolonged increase in soil disturbance by modern Park Service standards, while associated with increased siltation rates, failed to produce a noticeable change in elemental ratios. At Middle Rainbow Lake, artificial elevation of lake level by a meter, undoubtedly resulting in transfer of freshly available soils into the lake by wave action, failed to cause a ratio change large enough to attract interest (their figure 12). I believe the small Fe/Mn peak at about 1905 might possibly reflect this event. The allogenic silica peak at 1910 in figure 12.2 here may also record the damming.

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

The lake catchments at Big Slide and Middle Rainbow lakes are probably as or more heterogeneous with respect to distribution of varying soil chemistries than any of the other lake catchments, and for this reason I strongly believe that the case for their ratio-change criterion required to test the elk hypothesis is diminished, if not dismissed. Therefore, with the exception of the Floating Island Lake record, which is almost certainly influenced by road construction, the responsibility falls back to the investigators to show why there has been a pervasive increase by a factor of 2 to more than 3 in the deposition rates of sediments indicative of increased elk abundance. With climatic variations excluded for the most part (I agree with Engstrom et al. on this), elk seem more irrevocably implicated. That relatively small declines are seen during the 1950s and 1960s, when herd size was reduced, could be attributed more to the fragility and slow healing of soils in this semi-arid region than to the failure of the sediment data to precisely track elk numbers. Once cryptogamic soil crusts have been destroyed, centuries may be needed for them to regrow (Jayne Belknap, personal communication). The failure of Engstrom et al. (1991) to more fully recognize the roles of elk in direct disturbance of lake margins diminishes the value of the attention they drew to focusing as a general explanation for the sedimentation rate increases, which they do acknowledge. Focusing, as a mechanism for redistribution of sediments from the shore to the depths of the lakes, is given an almost passive connotation by the investigators. Moreover, they did not seriously attempt to understand why focusing should increase over time, as clearly demonstrated here for siltation rate. Focusing, when attributed hypothetically to episodically increasing numbers of elk, seems to explain the sedimentation database more easily, hence elegantly, than any of the other proposed mechanisms. First, owing to the storage of sediment between the active soil surface and the inactive depths of the lakes, it should have a relatively large time constant. As a result, rather large increases and decreases of erosive activity should be represented in the deep-water sediment by a highly subdued signal, with the extremes smoothed out. Engstrom et al. (1991) attempted to explore this process by taking additional shallow-water cores at Big Slide Lake, but they seemed unaware of the probability that lake levels had been dropping there over recent decades in response to outlet incision, intermittently slowed by decreasingly effective reconstruction of dams by beaver. The declines in gross sediment deposition rates since the 1920s and 1930s shown by their shallow-water cores probably reflect increased scour of near-shore sediments by wave action, resulting in augmented rates of deposition in deep water. My observations at Big Slide Lake in 1998, presented earlier, convinced me that destruction of beaver habitat by foraging elk is a focusing mechanism at least as important as trampling in the shallows along the shore. The pollen data profiles exhibit less complacency than those for allogenic silica deposition rate; that is, there are more peaks in most pollen profiles for nearly the same number of samples. This observation too suggests that for some

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mechanism common to both, sediment lags somewhat in reaching the depths of the lakes. Storage of sediment near shore is also implied by the fact that the data for largely airborne willow pollen do reveal four increases that might be attributed to partial recovery during the elk reductions of the 1950s and 1960s, whereas the allogenic silica data do not. The unprecedented decline in willow beginning in 1921 at Foster Lake closely matches the estimated date reported by Kittams (1948) of willow destruction in the Soda Butte Creek area a mile away. The allogenic silica record there shows an historically high deposition-rate peak at about that same time. The artificial damming at Middle Rainbow Lake, in addition to possibly producing the pronounced peak in siltation, also coincides closely with the beginning of willow recovery indicated by pollen increases at that site. Increase in willow might reasonably be expected with a higher water table produced by a higher lake level. These examples of both positive and negative correlation between willow abundance and siltation rate, one associated with rising lake level and the other with increased browsing near lakes, may help account for the greater number of pollen peaks in relation to siltation peaks in the records. In broad outline, that is, excluding short-term peaks, the records of the following cores exhibit secular willow pollen decline contemporaneous with silt deposition increase: Big Slide, Middle Rainbow (after about 1916 when cattle grazing began to diminish there, possibly as a result of the change from Army to NPS management), Slough Creek (after 1860 when Pb-210 counting statistics become reliable), Buffalo Ford, Floating Island, and marginally (owing to rarity of willow) Buck Lake. The pre-1900 pollen record at Big Trumpeter was apparently excluded from the report, and willow pollen was so rare there that a comparison is not reasonable. Only the record at Foster Lake suggests decline in both willow pollen and silt deposition rate, perhaps because the unusually thick growth of submerged vegetation has impeded focusing at Foster Lake. By any accounting, this result challenges the claim that the record is inconsistent with the elk hypothesis in terms of willow loss and increased transport of sediment to the lakes. To the contrary, with 83–86% of the profiles exhibiting secular covariance and a number of cases in which covariance coincides with historical events of decade timescale, it appears that willow pollen and allogenic silica deposition rather clearly support the elk hypothesis.

Relations Warranting Further Study The data show that covariance between willow pollen abundance and allogenic silica deposition rate is statistically significant only at slide lakes and supportive of the elk hypothesis only at those at lowest elevation. Moreover, the data show that the number of catchments recording increasing siltation over time decreased until recent decades. However, siltation rates have risen at an accelerated rate. This needs further, updated investigation.

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

Confounding Analytical Methods Earliest data, illustrated in Wright (1987), indicate that sample intervals for elemental analysis (at Buck and Big Slide Lakes) were for the most part quite consistent (~3.8 cm). Intervals in the Foster Lake core were fine in the upper part (roughly 2 cm) and coarse lower down (6–10 cm). However, intervals for diatom analysis reported in Engstrom et al. (1991) were coarse in the upper and lower portions of the Buck and Big Slide Lake cores and relatively fine in the lower part of the Foster Lake core. More important, the chemistry intervals and diatom intervals suggested by the plotted data points seldom match, so it is not possible to determine whether or not chemistry changes are also related to diatom changes, especially in the younger portions of the Buck and Big Slide cores. The pollen data also suggest a different sampling interval from that for sediment chemistry (Engstrom et al. 1991). Moreover, there is no mention of the extent to which samples for pollen and diatoms are contiguous within each core. It is impossible to determine whether these analyses represent counts involving contiguous lengths of split cores or short, noncontiguous pieces or even duplicate cores. Whatever the actual situation, the differences in data coverage from core to core and from parameter to parameter make it difficult to reliably compare events recorded for any two variables.

CONCLUSIONS One park publication (Anonymous 1992:11) acknowledged the implications of Engstrom et al. (1991) in overarching terms as discrediting critics thusly: “the wholesale movement of sediment by ungulates across the Northern Range . . . or a sudden increase in erosion caused by a sudden increase in the intensity of elk grazing, as presupposed by many earlier observers and researchers, have been disproved.” In their reply to my earlier comment (Hamilton 1994a), Engstrom et al. (1994) stated: “We do not believe, however, that our results ‘disprove’ historic changes in elk impacts, only that they do not support the alternative hypothesis.” I interpret this as a partial acknowledgment that their conclusions could not be used to endorse the natural-regulation hypothesis or support the above park assertion. The Engstrom et al. (1994) reply continues: “To shed any further light on the subject would require many additional sites, multiple cores (from each basin), and multiple proxies (geochemistry, diatoms, pollen, dating—as in our original study), and a systematic evaluation of all factors affecting elk utilization of these catchments.” I conclude from this partial reanalysis of their existing data that such a large-scale effort, though certainly desirable, might be unwarranted until their original data are reanalyzed synoptically and hopefully by them. Only by incorporating regional and site parameters into the analysis is it possible to distinguish between variables of unifying significance and those of

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more local or time-limited influence. Elevation, possibly distance up-valley, terrain, lake level, and ground- and surface-water regimes are important variables that significantly relate to sedimentation rates, rate changes over time, and trophic parameters in ways that make sense. As I suggested earlier (Hamilton 1994a), other site parameters such as aspect, catchment area, vegetation, bottom profiles, and discharge should be more fully incorporated into the assessment. My reassessment, primarily dependent on numerical values scaled from graphed data, visits to five of the lakes, knowledge of geological settings based on field observations and published maps, and historical information, has revealed events and trends that are both important to long-term management of the northern range and strongly suggest that elk might be implicated in contributing to erosion, depletion of native vegetation, and eutrophication of lakes over time on that range. Until the pervasive secular increase in allogenic silica deposition rate (and gross sedimentation rate) is rigorously attributed to an agency other than elk, then elk will continue to be a likely cause. Unless it can be convincingly shown that widespread secular decline in willow pollen is due to some cause other than browsing, then the elk population will remain as the most logical causal agent. Without an alternative explanation for the >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., <10%) of the vegetation and most of the energy is eventually released by detritivores and microorganisms if not by direct oxidation. Grasslands with large populations of grazing ungulates have been considered to approach aquatic systems most closely because a larger fraction of primary production is grazed off by herbivores than in other terrestrial systems, and thus is a potential constraint on herbivore numbers and biomass (Sinclair 1975). And primary production is the key parameter measuring herbivore sustainability of the system. Because a major fraction of the northern range area is woody vegetation (41% of the park portion is conifers, as discussed in chapter 8), it would not contrib259

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ute to the purpose of this book to attempt a reconstruction of its entire trophic structure even if the data were available, which they are not. But the production and consumption of that part of the vegetation used and affected by the ungulates, a comparison of these with other grassland systems, and some exploration of ungulate use on the system’s productivity are relevant to this synthesis.

UNGRAZED PRIMARY PRODUCTION Annual Net Herbaceous Production Four studies measured annual net above-ground primary production (ANPP) or year-end biomass of ungrazed herbaceous vegetation on the northern range during the 1980s. These were short-term studies—one for 3 years, two for 2 years, and one for a single year—during the 1986–89 period. I have summarized three of these (one did not report the measurements) in table 13.1 along with comparable measurements in three other areas in the Yellowstone region and two sets of measurements in different areas of the African continent. Since primary production in arid and semi-arid regions is so tightly coupled to precipitation (Blaisdell 1958; Sinclair 1975; Le Houérou and Hoste 1977; Wagner 1980), production values cannot be meaningfully compared except at specified precipitation levels. Hence I have shown the approximate precipitation levels along with the production estimates in table 13.1. Several points emerge from table 13.1. Except for the Frank and McNaughton (1992) results, the ANPP values for herbaceous vegetation on the northern range and surrounding region fall roughly at or below 100 g/m2. They are substantially below the production levels of the Sahelo-Sudanian zone and the Serengeti Plains of Tanzania at comparable precipitation levels. Again except for the Frank and McNaughton values, they correlate roughly with precipitation levels. Why production in the North American areas is so much lower than the African values can only be speculated on. The differences in insolation and temperatures favorable for plant growth between the North American areas at 44–45° N latitude and the Sahelo-Sudanian zone at 10–20° N and the Serengeti at 1–3° S, must be a partial explanation, as are the altitudinal effects. Elevations of the North American areas range for 1,620 m at the west end of the YNP northern range to 2,500 m at the southwestern Montana sites. Although the Serengeti, at 1,230–1,780 m (Sinclair 1979a:32), is at substantial elevation, the equatorial latitude more than compensates. Sinclair (1979b:7) states that there is usually enough moisture for an 8–month period (November through June) on the Serengeti “to produce grass growth.” The growing season on the YNP northern range is one to four months on upland, up to 4 months in “mesic-wet valley-bottoms” (Frank and McNaughton 1992). Why the Frank and McNaughton production values in table 13.1 are so much higher than the other North American ones needs some consideration.

Influences on Ecosystem Function III: Bioenergetics

Table 13.1 Mean Annual Net Above-Ground Production of Ungrazed Herbaceous Vegetation Locale

g/m2 Prod’n.

Mean Annual ppt (mm)

Source

285a 258–320 ~356 257.6e ~400 ~600

Blaisdell (1958) Coughenour (1991) Evanko and Peterson (1955) Mueggler (1971, 1972) Frank and McNaughton (1992) Merrill et al. (1993)

300 300 ~200–1,000k

Le Houérou and Hoste (1977) Le Houérou and Hoste (1977) Sinclair (1975)

Northwestern U.S. Eastern Idaho YNP no. range SW Montana SW Montana YNP no. range YNP no. range

52.8a 80.3b 91.6c 110.6d 363.1f 101.1g

Africa Sahelo-Sudanian Zone Sahelo-Sudanian Zone Serengeti

300h 700i ~500–600j

a13-year

mean of grasses and forbs, precipitation. live and dead grass and forb biomass in four exclosures, 1987 and 1988. cMean production by four “dominant” grass species. d5-year mean herbage production, southwest and northwest exclosures combined. e5-year mean precipitation for May–October. fMean of seven winter-range samples on four sites in two years. gMean of green herbaceous production on 25 sites. hConsumable. iTotal production. jAnnual production of grasses over seven-year period. kPrecipitation during growing period. bMean

These are not unduly high values resulting from above-average precipitation in the years of study (1988 and 1989). The first year was a “drought” year, and the second was “near average.” The years were comparable with those of Coughenour’s (1991) study in 1987, a slightly below average rainfall year and 1988, and the Merrill et al. (1993) study conducted in 1987. However, the Frank and McNaughton (1992) study stands out from the others in several respects. One is the vegetation sampling procedure. All of the other North American samples were taken by vegetation clipping at the ends of the growing seasons. Frank and McNaughton (1990) took their samples with a canopy-intercept method except for one site with sedge vegetation. In calibration studies, metal pins were projected through a wooden bar over vegetation calibration plots at a 53° angle. The number of times each pin contacted vegetation was recorded, and the vegetation in the plots was clipped, dried, and weighed. Plant biomass was then regressed on number of hits per pin, and an equation derived that was used to interpolate the results of the more

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extensive canopy-intercept recordings into biomass values. This is a method previously used by McNaughton (1979) for measuring Serengeti vegetation. Sampling-site selection may also have played a role. Two of the authors’ four winter range sites were “mesic-wet” on river terraces with production measurements two or more times the levels on their other two sites. However, the mean production of the two years on the latter sites was still 173.7 g/m2, well above the other North American measurements. Finally, the plant species composition on the authors’ four winter-range sites were not representative of the northern winter range as a whole. Coughenour (1991) states that the upland steppe of the northern range is dominated by the 3 perennial grasses, Agropyron spicatum (now Pseudoroegnaria spicata), Festuca idahoensis, and Koeleria cristata. In 3 of the 4 winter-range sites studied by Frank and McNaughton, the “herbaceous dominants” were three species of exotic grasses—Phleum pratense, Poa pratensis, and Bromus inermis. These are agricultural grasses bred for use in pastures and hay. In 1 of the 25 plots (no. 3) studied by Merrill et al. (1993), measured production was more than three times the mean for the 25. They suggested that this atypically large value was attributable to the dominance of the site by Phleum pratense. Frank and McNaughton’s fourth winter-range site was a wet sedge meadow with the dominant species a sedge, Carex rostrata. Wet meadows are a minor fraction of the northern range area, and establishing one of four winter range sampling sites in this type gives it and its production more representation in the measurements than its contribution to the total northern range primary production. Thus, for the foregoing reasons, I do not consider Frank and McNaughton’s production measurements to be typical of the northern winter range. And what appear to be more characteristic ANPP levels are only a fraction (e.g., 0.1–0.2) of the levels on the Serengeti. By implication, the northern range vegetation can support sustainably only a fraction of the number of ungulates that the Serengeti can support, as Frank et al. (1998) point out. In more recent studies (1999), Frank et al. (2002) again measured ungrazed ANPP at nine sites in YNP at elevations from 1,635–2,370 m that included winter, transition, and summer ranges. They did not describe the vegetation measurement technique, nor specify which of their measurements were taken on the northern range, nor tabulate their results. But one can infer with visual extrapolation of the values in their figure 2a that measurements at 8 of the 9 sites exceeded 100 g/m2, in one case was ~ 250. No information was given on the precipitation level in that year, and once again 2 of the 6 grass species dominating their sites were exotics. There have been fewer measurements of below-ground production per unit area in ungrazed herbaceous vegetation on the northern range. Using harvest procedures, Coughenour (1991) measured root biomass (not production) at the end of the growing season of 1987 and 1988 in four of the large northern range exclosures. Despite the fact that precipitation in 1988 was well below that of 1987, mean root biomasses were 1,214.3 and 1,277.8 g/m2, respectively, for the 2 years, not significantly different.

Influences on Ecosystem Function III: Bioenergetics

Frank et al. (2002) measured annual below-ground production of grasses at nine ungrazed sites with mini-rhizotron tubes in 1999. Although they did not tabulate their results, their figure 2b indicates values between ~ 300 and ~ 1,150 g/m2, with eight of the nine falling between ~ 300 and ~ 750. That these values are lower than Coughenour’s may reflect the difference between production and total biomass measurements.

Annual Net Shrub Production As discussed in chapter 7, the nonforested portion of the northern range within the park, 53% of the total, is shrub steppe with varying amounts of herbaceous and shrubby vegetation. Sagebrush is by far the dominant shrubby component, and hence any approximations of ungulate forage production must include measurements on these species. Though shrubs made up only 7.8% of elk winter diet in one study (Singer and Norland 1994), they were 49.7% and 80.5% of mule deer and pronghorn diets, respectively. Two studies have measured net annual above-ground production of unbrowsed sagebrush on the northern range. Singer and Renkin (1995) measured ANPP of shrubs inside the eight large exclosures on the northern range in 1986 and 1987. Their mean results were 18.7 g/m2 (18.0 for sagebrush alone) in the two Gardiner exclosures, and 82.6 g/m2 (73.6 for sagebrush) in the six higherelevation areas. Wambolt and Sherwood’s (1999) study of sagebrush inside the higherelevation large exclosures, as discussed in chapter 7, allows calculation of ANPP per unit area for shrubs ≥15 cm in diameter (see table 7.1):

 Mean shrubs / 60m 2    gm 2  ×  = gm prod’n./ m   prodn./ shrub 60 and (41.4/60m2) (147.0 g) = 101.4 g/m2 Thus, on the basis of these measurements, sagebrush production inside the exclosures is at about the same order of magnitude as the ungrazed herbaceous production measured in several areas on and in the vicinity of the northern range (table 13.1).

CONSUMPTION Several attempts have been made to estimate ungulate consumption rates of herbaceous vegetation on the northern range (percentage of annual herbaceous production consumed each year). The results have varied widely. Using values for elk food consumption per unit of weight per day, elk weights, a population

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of 12,000–14,000, elk use from mid-November to May, and estimates of forage standing crop, Houston (1982:433) estimated consumption of the average standing crop at 9–10%. Houston also cited a 1972 unpublished report by W. J. Barmore, who developed a model for forage consumption by a population of 10,000 elk. The model projected 5% consumption of the standing forage on the winter range inside the park. Frank and McNaughton (1992, with their results also reported in Frank and McNaughton 1993 and in Frank 1998) estimated consumption during the “snowfree” seasons of 1988 and 1989 on the same four winter-range sites discussed. For reasons that I will discuss shortly, consumption measured during the “snow-free season” on winter range seriously underestimates total consumption of a year’s production. Moreover, the typical production measurements for the winter-range sites (table 13.1) raise further uncertainty about the consumption estimates. Frank et al. (2002) subsequently measured consumption rates on nine winter-, transitional-, and summer-range sites in YNP with temporary 1.5 × 1.5 m exclosures moved monthly through the “snow-free (‘growing season’)” of 1999. Consumption was measured by summing the monthly differences between vegetation standing crop inside and outside the exclosures. The authors graphed their results but did not tabulate them. Hence the numbers must be approximated visually from the graphs. They did not indicate which of the points on the graphs represented the winter, transitional, and summer ranges. Subject to these uncertainties, their estimated consumption rates varied between ~18% and ~42% for eight sites (a ninth was not shown). However, it does not appear possible to estimate the proportion of ANPP consumed on winter range with growing-season measurements because of elk seasonal movements. The animals remain on the winter range for a short period after snowmelt, and consume some of the early vegetative growth of year x. This consumption may be measured by the snow-free season measurements. But the elk shortly move upslope to summer range, and winter-range consumption abates except in the localized area where the northern range bison subpopulation congregates and remains for the summer. Plant growth continues during the spring and summer and restores the biomass removed in the short grazing period prior to upslope movement (Merrill et al. 1994). The elk return to the winter range in fall after the end of the growing season, and spend 5 to 7 months grazing off the now-dead production of year x growing season. None of this latter, which constitutes most of the year’s consumption on the winter range, would be measured by growing-season measurements alone. Thus, whichever of the eight values in the Frank et al. (2002) figure 3 are the winter-range measurements, they must significantly underestimate the year’s consumption. Some speculative, perhaps order-of-magnitude estimates of annual consumption rates can be obtained by estimating total rangewide forage production and comparing this with absolute forage consumption based on (1) per

Influences on Ecosystem Function III: Bioenergetics

animal forage intake rates, (2) population size and sex-age composition, (3) and grazing-season length. If an ungrazed, herbaceous net, annual above-ground production rate is set at 100 g/m2 (see table 13.1: a hypothetical average), the total herbaceous production for the 43,900–ha shrub steppe would be 43,900,000 kg of air-dried forage. Some herbaceous forage is produced in the understories of the 33,700 ha of coniferous types (Houston 1982:86). Houston (p. 433) considers 350 kg/ha of biomass to be a conservative estimate of the understory. Some of this must be woody vegetation and not all of it current annual growth. It seems unlikely that herbaceous production in this type would be as high as that in shrub steppe. I suggest a hypothetical value of 35 g/m2, which would total 11,795,00 kg. Combined with production in shrub steppe, the total annual herbaceous production on an ungrazed northern range becomes, hypothetically, 55,695,000 kg. Adams (1982) placed elk daily food consumption at 10°C temperature for a 100-kg calf, a 240-kg cow (see Houston 1982:432), and a 350-kg bull (Adams 1982) at 21–38 g/kg, 17–31g/kg, and 16–28 g/kg, respectively. I have used the midpoints of these ranges: 29.5, 24, and 22, respectively. For a censused herd of 15,000, application of the 0.75 sightability bias correction (chapter 2) produces a total of 20,000 animals. This number subdivides into means of 66.2% cows, 18.0% bulls, and 15.8% calves according to the percentages of each given in Houston’s table 5.3 for 1975– 79. Combining all of these values, and assuming a 180–day grazing season, produces total annual elk consumption at 20,394,720 kg by a high elk population. Combined with an additional 16% for northern range bison (see chapter 9), and dividing the total by the above 55,695,000 kg production, estimates a consumption rate of 42% of net annual above-ground herbaceous production in a hypothetical vegetation whose level was unaffected by grazing. Limiting the consumption estimate only to herbaceous production is based on the minor proportion of woody vegetation in both the elk and bison diets (Singer and Norland 1994) on the northern range. The calculated 42% consumption rate underestimates consumption by all herbivores. Other herbivorous fauna, such as insects and rodents, feed on the vegetation during the growing season. Hence the amount of herbage available to the elk herd when it arrives on the northern range in fall must be less than the original production. More important, I will discuss evidence suggesting that a large elk herd grazing on the vegetation over a period of years or decades may reduce ANPP by as much as half. Hence there is probably substantially less than 100 g/m2 available to the biota, including the ungulates. In fact, there is anecdotal and indirect evidence indicating that most of the remaining phytomass is consumed. Coughenour (1991), based on 1987 and 1988 late summer measurements, observed that elk grazing in the vicinity of the large northern range exclosures “reduced standing dead material and litter biomass at most but not all sites and dates.” Merrill et al. (1994) measured aboveground phytomass outside a 2-year-old exclosure during the 1990 growing

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season and observed that elk winter and early-spring grazing “removed essentially all of the standing dead plant material.” In chapter 9, I reviewed evidence of winter nutritional deprivation in both elk and bison on the northern range. The fact that population process of northern range elk and bison is density-dependent (chapters 2 and 9) indicates a quantitative food constraint. The increase in the nutritional deprivation indicators (e.g., UN:C ratios) through the winter points to tightening shortage during the season. It seems entirely probable that most of the net above-ground annual production of the northern range herbaceous vegetation is consumed each year by a large elk herd and substantial bison population.

EVIDENCE OF GRAZING EFFECT Addressing the question of grazing effects on primary production of northern range vegetation must examine the possibilities of production stimulation through a mutualistic relationship between herbivores and vegetation, of grazing as a negative force on plant function, and the possibility of no effect. Consideration of potential positive effects raises the complex issue of overcompensation, here defined as the situation in which grazed plants produce more herbage within a growing season than the same plants in the same environment would produce if not grazed. In examining the subject of overcompensation, I do not wish to venture at length into the extensive and contentions literature on the matter (see Belsky 1986; Briske 1991; and the symposium in volume 3[1], February 1993, of Ecological Applications) except as it relates to the YNP northern range. Much of the work on overcompensation in YNP has been conducted by D. A. Frank and co-workers. Wallace and Macko (1993), evidently referring to Frank’s work, stated that “most of the dominant graminoids” in YNP “compensate for herbivory.” One park publication (Anonymous 1992) proclaimed this “a milestone scientific finding” although McNaughton (1979) had been writing on the phenomenon for more than a decade. In fact the YNP evidence is equivocal at best. Frank and McNaughton (1993; Frank 1998) first inferred from their studies in 1988 and 1989 (Frank and McNaughton 1992) that northern range grasses overcompensate for elk grazing. The first publication (Frank and McNaughton 1992) reported herbaceous ANPP measured at 12 sites: 4 in winter range, 3 in transitional, and 5 in summer range. The 12 sites were measured in 1988 and 1989. One of the winter-range sites (W1) evidently was not measured in 1988. Hence the data consisted of 23 site measurements summarized in table 4 of their 1992 paper. No comparisons were made of production in grazed and ungrazed vegetation, and the table 4 summary does not specify either. But comparison with the figure 2 histogram in the 1993 paper shows that the 1992 measurements in table 4 were of grazed vegetation. Inferences on overcompensation were drawn, in Frank and McNaughton (1993) and Frank (1998), in both cases based on the original study. They were

Influences on Ecosystem Function III: Bioenergetics

drawn from only 6 of the 23 site measurements and only 4 of the 12 sites. The results were presented in figure 2 of the 1993 paper and in figure 2 of Frank (1998). Frank and McNaughton (1993) lumped the data for the 6 sites and calculated a significantly higher ANPP on grazed sites than on ungrazed. Frank (1998) graphed the same data for the 6 sites and drew the same conclusion. My uncertainty begins with the question of why only 6 of the 23 site measurements and data from only 4 of the 12 sites were used. It is true that there was no consumption on 6 of the site-year measurements, and hence the overcompensation hypothesis could not be tested on these. However, in 10 other cases there was ungulate consumption, but whether there was overcompensation was not stated. Hence the conclusion of overcompensation was based on less than half of the data. In fact, the conclusion of overcompensation rests on only 3 site measurements, and an equal number indicate no compensation. In the 1998 publication, Frank regressed grazed ANPP on ungrazed production for each of the 6 site measurements. Only 3 of the points fell significantly above the b = 1 slope, whereas the other 3 were virtually on the line. If the latter had error bars, they would almost certainly not differ statistically from y = x, and equality between grazed and ungrazed production. Moreover, the nature of the sampled vegetation in this test must be raised again. The site falling farthest above the b = 1 slope was w4, the wet meadow site with the sedge Carex rostrata the dominant species. The other two points falling above the b = 1 slope were the 2 years of measurement on t2, where the dominant species was the exotic agricultural grass Phleum pratense. The dominant species for the points falling essentially on the line were Festuca idahoensis on site w1, Poa nevadensis and Stipa occidentalis for the 2 years of measurement on site s4, all native species. In sum, the evidence from these studies of overcompensation in northern range vegetation comes from three site measurements where the dominant species were a sedge in a wet meadow and the exotic pasture grass timothy. The three site measurements on native grass species did not show such a tendency. One must also wonder what the unmentioned site measurements on native grass species showed. On the basis of this evidence, one would have to conclude that overcompensation does not occur in the native northern range species. Frank et al. (2002) repeated production and consumption measurements in 1999 at nine sites distributed over winter, transitional, and summer ranges. Ungrazed vegetation was measured with five 0.5m2 quadrats in 15 × 15 m exclosures erected at the sites in 1998. Production of grazed vegetation and consumption were measured with 5 or 6 1.5 × 1.5 m exclosures moved monthly during the “snow-free season.” The authors did not specify how vegetation was measured, but root growth and mortality in the top 20–30 cm of soil were measured with mini-rhizotron tubes. Six grass species dominated the sites, two of which were exotics. Grazed net above-ground production (NAP), below-ground production (NBP), and whole-plant production (NPP) were plotted as functions of ungrazed

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NAP, NBP, and NPP. In each case, grazed production appeared to exceed ungrazed in about 6 of the 9 sites. No error bars or variance measurements were provided, and it is impossible to judge the variability of means based on 5 to 6 replicate measurements in what is highly variable shrub-steppe vegetation. The authors calculated grazing “stimulation” of NAP, NBP, and NPP by subtracting production inside the permanent exclosures from production in the grazed, outside vegetation. Stimulation was regressed on consumption, both in absolute terms (g/m2) and as percentages. In all cases (only 8 of the 9 sites shown), stimulation was a significant, linear function of consumption, implying that the higher the consumption, the higher the stimulation within the range of ~ 18– 42% grazing intensity. This is persuasive evidence, but I find it difficult to evaluate because of the contrary evidence in the earlier study. It is also a short, four-page paper that does not fully describe the methodology or provide information on variances in a shrub-steppe vegetation with a great deal of spatial variation. And there is no indication which of the data points represent the winter-, transitional-, and summer-range sites. Singer et al. (2003) have discussed compensatory response of northern range plant species at some length, but I have difficulty evaluating the results critically. Only 10 of 128 plant species (8%) showed higher above-ground vegetative production on grazed sites than on ungrazed. Only six of these (5% of 128) “strongly compensated for intermediate levels of ungulate herbivory by greater aboveground biomass production.” Thus, at best, overcompensation is highly exceptional within the northern range flora. Beyond these general values, the authors show higher production per year in total grasses and total forbs in ungrazed sites than in grazed (their figure 13.4). Yet an adjacent histogram bar shows higher total herbaceous production in grazed sites. Singer et al. reported no compensation in browsed willows. But their figure 13.4 shows twice the current annual growth per unit area in browsed sagebrush as in unbrowsed. Yet the data in the extensive Wambolt and Sherwood (1999) studies summarized in table 7.1 produce estimates of 101.4 g/m2 production inside four higher elevation exclosures (see previous discussion) and 33.5 g/m2 outside, a threefold difference under grazing protection. Singer et al. do not cite the Wambolt and Sherwood studies. Finally, Singer et al. show four grazing optimization curves in their figure 13.5 for percent difference in seeds per plant, seeds per m2, annual above-ground production, and difference in density of plants, all as functions of “Mean % reduction in height of plants due to herbivory.” The species are not stated, nor is there any reference to the source of the data so that the methodology, research design, and the original data can be examined. Moreover, there are no data points in the range from 0 to ~15% reduction in height along the abscissa that would justify constraining the curves to the zero point on the ordinates. With the data points given, the more parsimonious fits would simply be inverse straight lines, implying inverse relationships between grazing intensity and the four plant variables.

Influences on Ecosystem Function III: Bioenergetics

Thus the evidence for overcompensation in northern range species in the Frank and the Singer et al. studies is equivocal and at best suggests the phenomenon in a minor fraction (8%) of the species. That evidence is even more difficult to evaluate in the face of the following studies showing no evidence of overcompensation. Coughenour (1991) clipped year-end phytomass inside and outside four of the large exclosures across the northern range in 1987 and 1988. He concluded that plant production “was not reduced” by elk winter grazing, but at another point stated, “Winter grazing reduced live grass biomass at only one site in 1987, but inhibited grass growth at three of the four sites in 1988.” Coughenour did not find any inside-outside differences in root biomass in the 2 years. That he found no significant inside-outside differences in vegetative production is perhaps not surprising because he did not observe statistical inside-outside differences in standing crop with the Parker transect measurements in 1986 and 1989 (Coughenour et al. 1996), nor did Reardon (1996) with 1986 and 1990 chart quadrat measurements (refer to figure 7.8). Merrill et al. (1994) measured year-end root biomass in two northern range grasses inside and outside a 2-year-old exclosure in 1990. There was an initial decline in root biomass following spring grazing. But by the end of the growing season, root biomass had regrown to equivalence with that inside the exclosure. Wilsey (1996) experimentally clipped Stipa occidentalis plants, a native northern range species, and observed increased above-ground production. However, these studies were coupled with urea fertilization studies. If there is any point on which there was some agreement among the diversity of papers on overcompensation in the February 1993 issue of Ecological Applications, it was that overcompensation is likely to occur in situations of nutrient enrichment. Wilsey commented, “In a defoliation experiment with two other Yellowstone species, I found that clipped grasses had less crown and root biomass than non clipped grasses.” In a later paper, Wilsey et al. (1997) compared response to defoliation of YNP Stipa occidentalis, and grasses from the Serengeti and Argentine pampas. They found that crown and root biomass remained constant or increased after defoliation in the Serengeti species, but both decreased in grasses from the other two ecosystems. The authors, one of whom had coauthored the earlier 1993 paper inferring overcompensation in YNP grasses, suggested “that the lower intensity and increased temporal variance in grazing pressure in Yellowstone vs. the Serengeti, selected for plants that shift allocation away from roots and crowns in order to compensate for above-ground herbivory.” In chapter 3, I cited research by Caldwell et al. (1981) who studied the ecophysiology of clipped Pseudoroegneria spicata. Following defoliation, the plants continued allocating carbon to root growth but not to foliar growth. As a result, vegetative production in this grazing-sensitive species was reduced. Nowak and Caldwell (1984) also concluded that “compensating photosynthesis [in P. spicata] does not appear to be an important ecological component of herbivory tolerance.”

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In the same studies, Richards (1984) studied root behavior following defoliation. With continued allocation of carbon, root growth continued “unabated,” but root mortality increased through the winter and into the next year’s growing season. In total, root length declined during these periods as did total net production following clipping. Thus a considerable amount of evidence fails to indicate overcompensation in northern range grasses. Several authors have interpreted this on evolutionary grounds because an infrequently mentioned but implicit underpinning of the overcompensation hypothesis is a proposed long coevolution between herbivores and vegetation that produces a mutualistic relationship. Milchunas et al. (1988) commented: “The bunchgrass steppe region of northwestern United States and the Patagonian steppe represent the semiarid grasslands with short histories of grazing . . . [These are] selected for their tolerance to drought . . . regrowth potential is low. . . . Feedback mechanisms are not well developed in systems with short evolutionary histories of grazing.” Mack and Thompson (1982) comment in almost exactly this same vain. Detling (1988) pointed out that most well-documented cases of overcompensation come from tropical African grasslands. Patten (1993) comments that “Theories of the herbivore optimization and overcompensation do not appear to be supported by the response of western rangelands to grazing by native herbivores.” Milchunas and Lauenroth (1993) reviewed 236 grazing studies and found evidence of overcompensation in only 17% of the cases. Turner et al. (1993) have been cited as recent evidence of overcompensation in tallgrass prairie vegetation. But the authors showed only slight compensation to moderate clipping in the first year following years of “light” grazing. Response each year thereafter was “undercompensation.” Aside from the issue of overcompensation, the question remains as to the broader effects of ungulate herbivory on the primary production of the northern range. Belsky (1986) and De Angelis and Huston (1993) point out that consideration of grazer effects on vegetation extends considerably beyond the question of ecophysiology and overcompensation in individual plants to a hierarchy of levels that includes species populations, plant communities, and ecosystems. I suggested in chapter 7 that the 1958–90 twofold variations in herbaceous canopy cover, measured with Parker transects and chart quadrats (see figure 7.8), roughly coincided with the marked variations in elk numbers (table 2.1), subject to a few years’ time delay and some involvement with sagebrush competition. The low point at the beginnings of the measurements in 1958 and 1962 followed 70+ years of heavy use by a large elk herd. And the measurements increased after the herd reductions and with protection inside the exclosures. The annual cover measurements were taken at the ends of the growing seasons and therefore must have borne some relationship to each year’s ANPP. Thus the total herbaceous ANPP of the northern range appears to have varied with the ups and downs of the elk herd; when the herd has fluctuated to its higher levels, it may have reduced herbaceous ANPP by half. Without ANPP measurements over time, this remains a hypothesis. But it would be surprising if it were

Influences on Ecosystem Function III: Bioenergetics

not the case, given the fourfold variations in the numbers of elk using that vegetation for prolonged periods. The situation with sagebrush is similar (refer to figure 7.4). The first measurements taken of shrub canopy cover in 1958 and 1962 were, as with the herbaceous measurement (figure 7.8), the lowest recorded. It increased in ensuing years to the highest measurements recorded in 1990 in the high-elevation winter range. But by 1994, after at least 13 years of high elk numbers (table 2.1), Wambolt and Sherwood (1999) found sagebrush canopy cover inside higherelevation exclosures 2.7× that on the outside (table 7.1), comparable with their 3× difference in ANPP discussed previously. Thus prolonged use by large elk numbers may reduce ANPP in the sagebrush-steppe of the northern range by half to two-thirds. Moreover, aspen, other deciduous species, and riparian vegetation—which together may have covered 10–15% of the area at the time of park establishment—have been reduced even more. With the possible exception of coniferous vegetation, there is little doubt that total, northern range primary production has varied markedly with fluctuations in the elk herd.

IS THE NORTHERN RANGE A NORTH AMERICAN SERENGETI? A number of publications in recent years have drawn analogies between Yellowstone and the Serengeti ecosystem and more generally placed YNP in a global expanse of semi-arid grasslands with large numbers of coevolved ungulates. Indeed there are similarities. Yellowstone and particularly the northern range today have the highest abundance and diversity of ungulates of any area in the western hemisphere. Frank et al. (1998) point out that unlike other terrestrial ecosystems, a major fraction of the above-ground primary production is consumed by herbivores, especially ungulates: ~40% in YNP, ~65% in the Serengeti. The animals in both systems undergo extensive seasonal migrations: from tallgrass woodlands in the northwest portion of Serengeti ecosystem to the shortgrass during the rainy season; and from the YNP low-elevation winter (northern) range to high-elevation, largely forested summer range. These movements follow seasonal development of nutrient-rich, newly growing forage in the Serengeti rainy season and in spring in YNP. The authors infer considerable “regulation” or “control” of plant growth that benefits the ungulates. Their evidence indicates increased vegetative density in response to grazing, which increases grazing efficiency for the herbivores. They also infer grazing-induced overcompensation in both systems. However, their evidence for overcompensation in YNP (their figure 9) is the same as that presented in Frank and McNaughton (1993) and Frank (1998), which I concluded does not support an inference of overcompensation “in the native, northern-range species.” Frank et al. (1998) acknowledge that there are differences between the two systems. The Serengeti is almost three times the size of Yellowstone, has nearly

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four times the number of ungulate species, and ~52 times the number of ungulates. Moreover, the Serengeti is entirely grassland, whereas the open portion of YNP (20%) is shrub steppe, in which production by woody vegetation equals or exceeds herbaceous production. Most of YNP (80%) is forested, high-elevation terrain through which the northern elk herd disperses in summer to assume low densities by virtue of the area’s size. The Serengeti is a tropical grassland with up to twice the precipitation of the northern range, and an 8–month growing season (Sinclair 1979a). The northern range is a high-altitude, temperate steppe with a growing season varying from 2 to 4 months (Frank and McNaughton 1992). The Serengeti grasses are warm-season species that function with the C4 photosynthetic pathway, and the northern range grasses are C3 cool-season species. The Serengeti plains are grazed during the growing season, whereas the northern range is largely grazed in winter. Thus there are numerous similarities, but also as many differences, between the Serengeti and the contemporary YNP. But the historic and evolutionary dimensions are major differences not generally acknowledged. The weight of the evidence points to relative scarcity of ungulates and large carnivores prior to YNP establishment (chapter 3).The northern range vegetation does not have the coevolved traits that enable it to “cope” (Caldwell et al. 1981) with herbivory. The large wintering elk herd can be considered an artifact of park protection and inhibition from long-range winter migration. The Serengeti and prehistoric Yellowstone were largely different prior to European arrival.

Influences on Ecosystem Function IV: Nitrogen Biogeochemistry

14 Any success in analyzing ecosystems is likely to be found in careful and explicit consideration of the organization underlying the complexity. —R. V. O’Neill, D. L. DeAngelis, J. B. Waide, and T. F. H. Allen

CONTEXT The seemingly straightforward question of how ungulates, primarily elk, affect the biogeochemical functioning of the northern range ecosystem is a simplistic abstraction of an inscrutable complex of entities and processes, both organic and inorganic, that function within the plants, animals, microorganisms, and inorganic components of the system. Frank (1998) lists 12 chemical elements that he extracted from the vegetation alone. Each must function within its own complex of components and processes. Analyzing the biogeochemical functioning of an ecosystem requires (1) delineating the limits of the system; (2) identifying and measuring the input processes that bring biogeochemicals into it; (3) measuring the distribution of the chemicals among the system’s components; (4) identifying and measuring the internal processes within the system that move the chemicals in space and time, and change their states; and (5) identifying and measuring the output processes that eliminate them from the system. Some chemicals and processes can be stimulating to components of a system, some can be inhibitory. But these do not determine the long-term trends in a system that are a function of the balance between input and output processes. The subject of ungulate effects on the northern range ecosystem addresses, of course, only a small subset of the biogeochemical complexity of the area. Only 273

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a small number of projects has addressed this subset, focusing on the hydrologic aspects reviewed in chapters 11 and 12 and on nitrogen (N). The latter have mostly been the excellent studies of D. A. Frank and co-workers. Nitrogen has received research emphasis in arid and semi-arid ecosystems, including Yellowstone’s northern range, because it is considered to exert some limitation on plant growth, perhaps second in importance only to water. Consistent with the major purpose of this book—to analyze the effects of elk on the northern range ecosystem—I review the studies to date on nitrogen processes affecting the states and content of N in the northern range ecosystem. The intent is to explore how the northern herd may be affecting the internal N dynamics of the system, and whether it is inducing any trend in the system’s N content by altering input-output processes.

NITROGEN BIOGEOCHEMISTRY ON THE NORTHERN RANGE Internal Dynamics Figure 14.1 is an oversimplified schematic of nitrogen dynamics in the northern range ecosystem. For the purposes of this analysis, I define the limits of the system as the upland terrain in the northern range area depicted in figure 1.1. These are represented schematically as the boundaries of the large box on the left side of figure 14.1. I will explain why I am restricting this to the upland terrain. Nitrogen is primarily usable to plants as nitrate (NO3–)and ammonium (NH4+) ions in the soil. Hence the northern range research has focused on: 1. Processes within the system that convert N in other forms (e.g., plant and animal tissue, animal excreta, soil organic matter, all called “Immobile N” in figure 14.1) to these ions (called “Mobile N” in figure 14.1). These processes are called “mineralization.” 2. Processes that add or remove N from the system, immobile or mobile, potentially inducing some trend in system functioning over time. These are shown as the arrows into and out of the total N box on the left of figure 14.1 3. The effects of the northern herd on these processes. The work to date indicates higher N mineralization rates on the northern range in the presence of ungulate grazing than in its absence. In studies at seven matched sites inside and outside four of the large northern range exclosures, average net N mineralization rates were twice as high outside the exclosures as inside: 3.8 g N/m2/year versus 1.9 g N/m2/year (Frank and Groffman 1998). With investigations inside and outside three of the same four exclosures, Augustine and Frank (2001) measured higher N mineralization potential as a function of microbial respiration inside two of the three, but not in the third (their table 1).

Influences on Ecosystem Function IV: Nitrogen Biogeochemistry

Figure 14.1 Simplified schematic of the northern range nitrogen budget. Boxes are nitrogen pools or compartments. Arrows represent input, output, and internal cyclic processes which bring nitrogen into, remove it from, and convert its chemical state in the northern range ecosystem.

Without grazing, herbaceous vegetation senesces at the end of each growing season and is gradually broken down physically to form duff, which is incorporated into the soil. Breakdown processes, both abiotic and biotic, gradually reduce the organic matter further to a state in which microbial action can mineralize the N content into the mobile forms. At this point the vegetation can once again take it up in subsequent growing seasons. Thus N cycles relatively slowly and remains in immobile states for prolonged periods of time in the absence of grazing. With grazing, ungulates consume some of the vegetation, convert some it to their own tissue, but excrete the remainder as dung and urine, in which state it is closer to mineralizable form than the original vegetation. Coughenour (1991) comments that the effect of grazing on the northern range is the redirection of

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biomass flow from decomposers into elk, accompanied by enhanced rate of N recycling back to elk. Frank and Groffman (1998) observe that herbivores improve the organic matter quality of soil by increasing the labile (i.e., mobile) N fractions and reducing the recalcitrant (to microbial breakdown) fractions (i.e., immobile). Several other processes contribute to the complexity of ungulate acceleration of the N cycle. Merrill et al. (1994) measured higher densities of microbe-feeding nematodes in the soil outside a 2–year-old northern range exclosure. The authors hypothesized that higher root mortality associated with grazing the two grasses on their sites stimulated microbial increase, which in turn supported the nematode increase. In turn, the nematode feeding released microbial-bound (immobile) N. The end result was to increase N turnover. Hamilton and Frank (2001) clipped foliage in laboratory-grown plants of Poa pratensis, a common northern range grass. The clipping stimulated higher exudation rates of carbon-containing fluids from the roots than the rates exuded by unclipped plants. Heterotrophic microbial populations increased in response to the elevated resource, and in the process released mobile N available for enhanced plant uptake and growth. The implication was that grazing would provide this same stimulus. Thus the northern range evidence indicates that ungulate grazing, primarily by elk and bison, accelerates N cycling within the system and provides mobile N for plant uptake. Possibly as a result, the N content of northern range vegetation in grazed areas has been shown in several studies to be higher than that of vegetation inside exclosures: Coughenour (1991) at the four large exclosures that he sampled in 1987 and 1988, Merrill et al. (1994) at a 2-year-old exclosure sampled in 1990, and Augustine and Frank (2001) at two out of three large exclosures sampled in 1998. In addition, ungulates fertilize plant growth directly with their excreta. Day and Detling (1990) measured higher root and shoot biomass in bison urine patches in their short-grass plains study area. Augustine and Frank (2001) observed plant responses to urine patches on the northern range. Despite these stimulatory effects, ungulate use does not in and of itself increase the total nitrogen content of the system. The high plant N content could result only from the more rapid cycling through the system and the high proportion of mobile N forms available for plant use. The amount of N in the system, and trends therein, are a function of input-output processes that bring it in and expel it. These can be significantly influenced by ungulate actions. Moreover, there is some indication that ungulate-induced internal acceleration of the cycle can increase the rates of N output (Lauenroth et al. 1994). When the vegetation is ungrazed, N is in the system in perennial plant tissue, duff, and soil organic matter for extended periods of time, thus conserved in the system in immobile forms. When the tissue is grazed and rapidly converted to mobile forms, it is vulnerable to loss by volatilization, denitrification, and surface erosion.

Influences on Ecosystem Function IV: Nitrogen Biogeochemistry

Input-Output Processes Nitrogen Fixation, Denitrification, and Volatilization Arid and semi-arid land soils are typically N deficient (West and Klemmedson 1978), possibly due in part to a dearth of plant species with symbiotic nitrogenfixing root nodules. Frank et al. (1994) found no nodulating species in one northern range area they studied. The low N levels may also be due in part to the high pH levels of the soils, which promote rapid denitrification and ammonia volatilization. Atmospheric N is fixed in Intermountain soils primarily by free-living bluegreen algae and by lichens that are symbiotic associations of blue-green algae and fungi situated in surface microphytic crusts (West 1991). Fixation largely occurs in spring pulses when soils are still moist from spring thaws and seasonal precipitation. Frank et al. (1994) comment that N mineralization rates were high in their study area when soils were moist after snowmelt. As the soils dry, N is lost through denitrification (Westerman and Tucker 1978) and to a lesser degree from ammonia volatilization (Klubek et al. 1978). Thus plants have a short window of time during which there is a significant amount of fixed N they can take up for growth. It is not known whether there is significant denitrification and/or ammonia volatilization of the nitrogen mineralized by bacterial action. Numerous studies have shown that the microphytic crusts can be broken down by ungulate trampling (West 1990). No systematic study has examined this effect on the northern range, but there is anecdotal evidence that suggests it may occur. Coughenour (1991) measured 2 to 3 times as much lichen biomass inside as outside two of the large exclosures, but higher biomass outside a third. From my own casual observation, microphytic crusts appear to be much more intact inside the Gardiner exclosures than outside. Urine deposition appears to have a mixed effect. As commented, there appears to be a direct fertilization effect on plant growth. But Frank and Evans (1997) report increased ammonia volatilization from urine patches and dung as do Frank et al. (1994).

Erosion Surface erosion has been shown to move soil N from higher to lower sites where it is no longer available for use by plants on the uplands. Chapters 11 and 12 presented evidence that surface erosion had intensified since park establishment, leading to deposition in small lakes and ponds. This was the rationale for my defining the limits of the northern range ecosystem, in terms of N cycling, as the upland terrain. Several studies have shown that surface erosion in grazed, hilly terrain moves N compounds downslope. Schimel et al. (1985) studied a hill slope in the shortgrass steppe of eastern Colorado that had undergone several cycles of erosion and deposition. The total mass of N, annual N mineralization rates, and plant uptake of N all increased downslope.

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Northern range studies point to the same effect. Augustine and Frank (2001) measured spatial variation in soil N and N mineralization on slopes inside and outside three of the large exclosures. On the outsides, both variables were higher on an upper bench than on the steep slope below. The authors suggested that the higher values on the bench may have been due to bedding site selection by elk. However, Frank et al. (1994) measured these same variables on a ridge top, upper bench, slope, lower bench, and streamside zone in the drainageway of Blacktail Deer Creek. The two bench sites had higher soil N and N mineralization rates than the ridgetop where elk also bedded and the slope sites. But inorganic N levels were highest on the riparian sites, more than could be produced by the relatively low mineralization rates. The authors concluded, “there must have been a supplemental source of N for the riparian site. . . . A large portion of the N budget at that site may have been exogenously provided by overland flow and/or subsurface leaching from upland habitat directly above or up-drainage from the site.” West (1990) reviewed numerous studies documenting increased wind and water erosion following trampling damage to microphytic crusts. That Augustine and Frank (2001) did not find the same correlation between topography and soil N levels inside the exclosures that they found outside could reflect more stable soil surfaces inside the fenced areas and the influence of ungulate presence on erosional loss on the outside.

Seasonal Elk Movements Frank et al. (1994) suggested that northern herd animals build body condition and weight on the high-elevation summer range, then lose weight on the winter range. I reviewed the evidence for that catabolic loss in chapter 9. To that tissue loss may be added the winter mortality of animals that summer at higher elevations and winter on the northern range. Singer et al. (1997b) measured 35% average winter mortality of calves alone. The effect in both cases is to carry N assimilated from summer forage to the winter range in the form of excreta and body tissue. There may be other seasonal transports possibly moving N out of the northern range. The cows are largely bred in late September and early October (Murie 1951). Within a month or two, they move onto the northern range where they spend most of the period of their 250–day (Raedeke et al. 2002) pregnancies. Their fetuses grow to birth weight despite the fact that the cows are losing their own body tissues. Where the calves are born depends on the timing of snowmelt (Barmore 1980). If it is early, they may be born on transitional or summer range. If it is late, many are born on the winter range. Singer et al. (1997b) measured an average 28% mortality of radio-collared calves during summer (based on his 0.72 summer survival rate) over a four-year period. Roughly half (54%) of these died between 3 and 10 days of age, with predation by far the major cause of death. Probably most of this occurred on the

Influences on Ecosystem Function IV: Nitrogen Biogeochemistry

winter range, which would simply retain the nutrients on that area that had been subsumed in fetal growth. But some fraction of the remaining half of calf deaths must have occurred on transitional and summer range. This would transport nutrients out of the winter range. In total, seasonal movements must provide both N input and output to the northern range.

DISCUSSION Clearly the subject of elk effects on the northern range biogeochemistry is an enormously complex one that has had only limited study largely focused on N. The evidence is persuasive that elk accelerate the N cycle by eating vegetation that would otherwise accumulate as duff and rough soil organic matter, which is slow to return to inorganic mineral form. The forage consumed by elk is converted to the animal’s excreta, which is more rapidly mineralized than the unconsumed duff and thus more quickly made available in forms once again usable by the plants. But the recycling does not in and of itself increase or decrease the northern range N content. That is determined by a complex of input and output processes—fixation, denitrification, ammonia volatilization, lightning and aerosol transport, erosion, seasonal movement of elk body tissues and fluids—which are largely unmeasured. Until they are, it will not be possible to reconstruct the N budget of the northern range and the influence of elk on that budget. One preliminary indication, at least since the large exclosures were built in 1957 and 1962, would be a comparison of soil N content inside and outside of them. But the evidence is mixed. Augustine and Frank (2001) measured higher soil N content outside three of the large northern range exclosures than inside. But Frank and Groffman (1998) did not find significant inside-outside differences in their measurements of soil % N and C/N ratios (their table 1). If the northern range N budget is still largely undetermined, the biogeochemistry of other important nutrients is totally unexplored. With different complexes of internal, input, and output processes, the entire subject is an almost completely unknown world.

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Synthesis

15 Consilience: The explanations of different phenomena most likely to survive are those that can be connected and proved consistent with one another. —Edward O. Wilson

TOWARD AN ECOSYSTEM PERSPECTIVE The central purpose of this book is to synthesize the available evidence toward generalizing the effects of elk on structure and function of the northern range ecosystem and by implication to test the validity of the natural-regulation ecological hypothesis. The amount of evidence is immense. But contrary to statements in park publications, the research has not had an ecosystem orientation. There has been no conceptual model of the system that would guide systematic selection of projects addressing its structural and functional composition and would allow their integration into inferences about system behavior. Rather, the research has largely proceeded by periodic emphases in response to the public relations or policy sensitivities of the moment: natural regulation, “is the northern range overgrazed?,” the 1988 fires, wolf introduction. Emphasis has largely been placed on a few components of the system, primarily charismatic animal and plant species: elk, bison, aspen, willows, and currently wolves. There has been little or no research on most of the fauna—small mammals, birds, invertebrates—or on soil microfauna and flora and on several aspects of the vegetation. Except for ungulate demography, herbaceous and sagebrush production, erosion, and some aspects of nitrogen cycling, there has been little attention to system processes. Most projects have spanned only 1 to 2 years; the ungulate 280

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censuses and periodic measurements inside and outside the large exclosures are exceptions. The temporal dimension is important for two reasons. First, the several changes in elk numbers over the past 132 years (refer to figure 5.1) constitute a quasi-experimental manipulation, with two replications each of low and high numbers, of the independent variable. In an ideal world, continuous measurements of key system variables through this period would have given a clear, quantitative picture of elk effects; it would have given a basis for predicting the quantitative consequences of management alternatives. A second reason for the importance of continuing measurements over time is in synthesizing the character of the northern range ecosystem. O’Neill et al. (1986) and Allen and Hoekstra (1992) conceptualize ecosystems as dual hierarchies of components and processes, each structured by systems of constraints. They argue that one cannot reconstruct system behavior knowing only component behavior. Without temporal measurements of processes, it has been necessary to improvise with component information not originally obtained for that purpose but that can be used to reconstruct at least qualitative trends, and in some cases semi-quantitative ones. These include comparison of early anecdotal accounts with contemporary observations; early photographs matched with contemporary ones of the same locations (Kay 1990; Meagher and Houston 1998); dendrochronology of aspen and sagebrush (Romme et al. 1995; Ripple and Larsen 2000; Wambolt and Hoffman 2001); Keigley’s plant architectural analyses; and inside-outside measurements of the large exclosures over a 33-year period. I infer system function in these cases over the intervals between two or more measurements or descriptions of state variables. I have correlated or related these functions, based on what I have discerned in the preceding nine chapters about changes in northern range state variables, with the four elk population phases over the past ~132 years. This covers the period from prehistory—some unspecified period prior to park establishment in 1872—to the mid- or latter 1990s. The newer annual winter movement from the park of, on average, a third of the herd may be creating a new set of conditions that have not yet been adequately researched. However, the effects of twothirds of an equilibrium population of 16,000 censused animals remaining in the park may be no different from the effects of 10,000 censused animals during the 1930s–1950s (figure 5.1). Thus the question remains open. A second, new variable is what the wolf effects will be—impossible to say at this point in time. Hence the reasons for not extending inferences beyond ca. 1999, except for the calculations of elk population equilibrium in chapter 2, and speculations on other major environmental changes under way. I have also grouped elk influences into first- and higher-order effects on major northern range system components and processes for which there is some evidence. This is a first step in conceptualizing the system in the O’Neill et al. (1986) and Allen and Hoeckstra (1992) hierarchical framework and arraying the constraints in this structure.

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For some readers, my evidence sources may not justify this attempt at synthesis. They do indeed vary in quantity and quality. But they converge on a paradigm in which inferences about the constituent parts and functions are internally consistent. And the inferences are consistent with what is known about the biology of those parts and contemporary ecological theory. Like all science, this paradigm is a pro tem judgment based on the prevailing evidence to date and remains the most probable unless and until overwhelming new evidence points to an alternative model. Thus my purpose in this chapter is to review the major conclusions of the preceding chapters and synthesize them into a conceptual model for the northern range ecosystem.

FOUR LEVELS OF ELK ABUNDANCE As discussed in chapters 2–5, the weight of the evidence points to a northern herd that occurred at four levels of abundance over the period from before park establishment to the beginning of the new millennium (figure 5.2). The northern herd, fluctuating in the short term as well as changing through these four periods of abundance, is that group of animals summering in the higher elevations of what is now the park. Prior to 1872 it moved north of present park boundaries to winter at lower elevations. From the latter 1800s to 1988, it wintered inside the park in an 83,000 ha area and an adjacent 17,000–ha area outside, collectively considered the northern range. Since the winter of 1988–89, on average a third of the herd has wintered in an additional 52,663 ha outside the park adjacent to the original 100,000. For the first period of abundance (pre-1872 to ~1884) I offer a speculative estimate of ~5,000–6,000 animals on average (chapter 3). A reasonable hypothesis is that numbers were held at this level by some combination of aboriginal hunting and predation by large mammalian carnivores. Variations in winter severity probably induced some fluctuations in the population, allowing increases in mild winters, reducing numbers in severe ones. Between approximately 1884 and 1962, the second population period, the herd exceeded censuses of 6,000, increasing rapidly from 1884 to a reported 20,000–35,000 (chapter 4) by the second decade of the 1900s. These may be conservative numbers, and the herd may well have exceeded them by some margin. Numbers declined after 1917, initially as a result of a series of severe winters but subsequently in response to increasing hunting removals outside the park, increasing herd reductions by the park, and I suspect increasing nutritional deprivation owing to forage depletion by the large herd. This original increase and the continuing high numbers, even if declining, were facilitated by removal of the prepark constraints: ending Native American hunting of the animals, prohibition against Euro-American hunting of animals in the park, and reduction of predator numbers. They were facilitated by blockage of what had been the annual exodus out of the park area to lower-elevation winter range.

Synthesis

In the third period, the herd was reduced to ~ 5,000 censused animals in the winter of 1961–62 by park control efforts and outside hunting. It was reduced further through 1967–68, the last year of park control, to a censused 3,172. With cessation of park control, the herd immediately began increasing, rising to 9,981 censused animals by 1972–73. Thus the censuses counted fewer than 6,000 animals from 1959 to 1970, an 11-year period. The herd grew continuously after 1970, with occasional severe winter setbacks, through the 1970s, 1980s, and early 1990s to censuses of 16,019 by 1981–82 and 19,045 in 1993–94. It declined somewhat in the latter 1990s and early 2000s, but continued at numbers above 10,000. As discussed in chapter 2, the calculated equilibrium number for 1988–89 to 2002–2003 is 16,800 censused animals. The increase since 1968 has occurred in the absence of any park control efforts. But it has proceeded in the face of increasing outside hunting kills and increasing density-dependent pressures imposed by growing nutritional deprivation. These two constraints together have set the number at which the herd, on average, has achieved equilibrium. I have stressed that these are “censused” numbers to standardize the seasons of the counts (winter), taken after the fall hunting and park removals. Fall numbers have thus exceeded the censused numbers. And I have specified the census numbers because of the several studies that have shown the censuses to undercount actual population size by approximately 25% (chapter 2). The latter could be obtained by dividing the census numbers by 0.75. Thus the herd has varied through alternating periods of low and high abundance, here placed at numbers below and above 6,000 censused animals. The dates of these four periods are slightly different from, but their population levels have resulted from, the four policy phases outlined in chapter 1. Human actions have been major determinants of all the levels.

EFFECTS ON SYSTEM COMPONENTS First-Order Effects on the Vegetation I estimated in chapter 6 that aspen woodland at the time of park establishment occupied ~ 8,000–12,000 ha and 10–15% of what is now the park portion of the northern range. By the 1960s and 1970s, at the end of more than 75 years with high elk numbers, Houston and Barmore estimated the area of aspen at 1,418 ha, and 2% of the park portion of the northern range. Thus the type may have been reduced by 80–85%. There is fragmentary evidence of aspen response during the 1962–73 elk-population low. But the Romme et al. (1995) and Ripple and Larsen (2000) dendrochronological analyses showed no tree recruitment during the postreduction herd recovery. Aspen stand structure is now comprised of low densities of superannuated trees at or near the limits of their longevity and with understories largely of grasses,

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especially exotic species. This contrasts with stands in early photographs, in contemporary exclosures, and outside the park that had/have several strata composed of multiple aspen age classes and diverse understories of shrubs and forbs. With continuation of present trends, aspen as a species and as an ecological subsystem could largely disappear from the northern range in coming decades. At 43,900 ha, the sagebrush-steppe is the most extensive of the major vegetation types and provides most of the forage for the several ungulate species. With no early measurements, and sagebrush too short-lived to provide dendrochronological data, the photographic record provides the only evidence of early sagebrush abundance. That record (see figure 7.3) shows it to have been a substantial component of the northern range sagebrush-steppe. By the 1920s, park biologists were expressing concern over heavy browsing impacts; the first measurements of its abundance in the vicinity of eight new exclosures in 1958 and 1962 showed it to have half or less the canopy cover that it would develop under subsequent protection (figure 7.4). The photographic record similarly showed a decline (figure 7.3) between 1920 and 1960. Periodic measurements at the higher-elevation exclosures showed shrub canopy-cover increases, starting during the elk population low, and continuing to 1990 to levels ~2 to 3 times those of 1967 (figure 7.4). But the 1990 levels inside the exclosures were ~twice those on the outside. Photographic evidence suggested the same increase (figure 7.3). Cover measurements inside the Gardiner exclosures also showed increases to 1990, but continuing decline on the outside to virtual elimination of the species in the area (figures 7.4, 7.5). In 1994, Wambolt and Sherwood (1999) measured sagebrush density, production per plant, and production per unit area inside the higher-elevation exclosures at 1.61, 1.88, and 3.03 times those on the outside, respectively. Thus sagebrush abundance has waxed and waned over time with changes in elk numbers. It is significantly more abundant inside exclosures than outside. There are no sources of evidence that allow inferences on the condition of northern range herbaceous vegetation prior to park establishment when elk numbers were low. It cannot be judged from photographs. The first accounts of herbaceous vegetation condition were expressions of alarm by park investigators in the latter 1920s, following more than 40 years of high elk numbers. Houston (1982) attributed these conditions to the drought of the 1930s, but the early accounts preceded the drought years. By 1958 and 1962 at the end of ~75 years of high elk populations, two sets of herbaceous vegetation measurements inside and outside the new exclosures produced low values (figure 7.8). Parker transects measurements both inside and outside the exclosures increased about twofold between 1967 and 1981, the increase starting during the elk population low (see figures 7.8). Cover measurements were not repeated until 1986, by which time they had also increased to more than twice the 1967 measurements (figure 7.8). The Parker measurements declined between 1981 and 1990, the cover measurements between 1986 and 1990 (figure 7.8). Inside-outside measurements have

Synthesis

not differed significantly at all exclosures, although measurements at the Gardiner exclosures have consistently been lower than those at the higher elevations. Park investigators have generally inferred that the evidence indicates no significant effects of elk use on the herbaceous vegetation, generally accepting Houston’s discount of the early reports. Indeed comparisons of vegetation composition inside and outside the exclosures have shown little difference. Yet the early accounts, the low 1958–62 exclosure measurements, and the subsequent increases cannot be dismissed out of hand, especially because they coincide with major differences in elk numbers. I hypothesize the following scenario for the effects of elk on the northern range herbaceous vegetation. Elk have a negative direct (first-order) effect on the herbaceous vegetation by grazing it, and a positive indirect (second order) effect by suppressing sagebrush competition. After ~75 years of grazing and browsing by a massive elk herd, both sagebrush and herbaceous vegetation had been driven to low levels by 1958–62 (figures 7.4, 7.8). With reduction of the herd and construction of the exclosures, sagebrush increased from 1967–90 (figure 7.4), herbaceous vegetation from 1967–81 (figure 7.8), both inside and outside the exclosures. But by 1981, both grazing pressure and sagebrush competition had increased outside the exclosures to the point of reducing the herbaceous between 1981 and 1990 even though the sagebrush was only about half as abundant as that on the inside. Herbaceous abundance on the inside, although freed of elk herbivory, was forced to coexist with twice the sagebrush competition on the outside and could increase no more than that on the outside. Despite comparable herbaceous abundance inside and outside the exclosures, the lower sagebrush density on the outside effects a more open plant community and allows invasion of nonnatives, something prevented by the dense vegetation on the inside. Thus the evidence implies that the abundance, production, composition, and physical structure of the northern range sagebrush-steppe have varied during park history in response to alternating levels of elk use. These changes include the invasion of exotic plant species facilitated by that use. Whether the herbaceous decline of the 1980s (figure 7.8) would continue, and whether a sagebrush decline would set in with continuation of elk numbers of the 1990s, thus moving the vegetation back toward the conditions of 1958–62, is not known. Although conifers occupy some 41% of the park portion of the northern range, they do not contribute a significant amount of forage for browsing ungulates as shown by food habits studies and implied by the virtually universal highlines that remove the foliage from browsing reach. But they provide resources for other components of the biota, and the highlining chronology gives another indication of elk impacts on the northern range ecosystem. The photographic record does not show significant highlining before 1900. Reports of park personnel and photographs indicate its appearance in the early 1900s, some 2 to 4 decades into the first period of high elk numbers. Demographic research is needed to ascertain whether there are young replacement trees in the mature conifer stands.

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There are no data on the early abundance of the 20+ species of deciduous shrubs and small trees that are minor components of the northern range vegetation. At least eight of these produce berries edible for humans and wildlife. Early explorers commented on their prevalence, but early park investigators 30–50 years later commented on heavy browsing impacts on these largely palatable species. Kay (1990, 1995) measured extreme inside-outside differences in these at the exclosures and near absence of berry production on the outside plants. All observers agree that the photographic record and reports of early observers indicate robust willow growth in moist areas throughout the northern range prior to and in the early years after park establishment, and that the area occupied by the species has declined sharply during park history. Early photographs and observers’ comments indicate full-statured willow growth in suitable habitats throughout the northern range. This growth endured in a period when, as the evidence shows, elk were present in low numbers and migrated out of the northern range in winter. The first reports of browsing impacts appeared in 1915 and the 1920s. Kittams commented on and photographed willow hedging and disappearance in the 1940s. Following construction of the large exclosures in 1957 and 1962 at the beginning of the elk reductions, Barmore (1980) measured immediate growth response of willows inside, but continued suppression outside. But he, O’Gara (personal communication, 1996), and Patten (1968) all inferred some regrowth during the population reduction. Willow suppression continued through the latter 1900s during the elk population recovery (figure 10.4). Kay (1990) shows a 1965 park photograph with willow growth that was no longer evident in his 1988 retake of the site. He also comments on growth of willow inside the Tower Junction exclosure constructed in 1957 but its disappearance by 1973 following dismantlement in 1971. Thus riparian willows are another vegetation component of the northern range that has waxed and waned with the changing phases of elk abundance: full-statured and widely distributed during the elk population low at the time of park establishment; heavily impacted during the first population high at the end of the nineteenth and first half of the twentieth century; slight response during the short period of reduced numbers; then continued suppression during the postreduction population recovery. Park publications have stated that willow has declined 50% during park history. Kay (1990) has claimed 95% reduction. Unfortunately these numbers fail to distinguish between areas occupied by full-statured shrubs or to areas where the species has been eliminated completely (figure 10.1). It is unfortunate that there has not been a systematic study of this question by comparing early photographs with field inspection as Kay and Wagner (1996) conducted on aspen clones. Early park investigators commented on browsing impacts on cottonwoods by the 1920s. Kittams (1948) compared early photographs of cottonwood stands with his 1940s retakes and found that they had “virtually disappeared.” Dendrochronological analyses of three samples of narrowleaf cottonwoods by Keigley

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(1997b, 1998) showed active tree establishment between 1840 and 1894, then only one established between 1894 and 1934, and none between 1952 and 1962. He aged 17 trees that grew between 1963 and 1974, the approximate period of low elk numbers, and then none established between 1974 and 1992 when he conducted his study. More recently, Beschta (2003) examined all cottonwood trees in a 9.5–km2 area of the Lamar River valley in 2001 that were >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.

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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.

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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.

Synthesis

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.

Synthesis

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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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

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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 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

Why the Science Missed the Mark

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

Why the Science Missed the Mark

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

Why the Science Missed the Mark

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

Why the Science Missed the Mark

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

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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

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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.

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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).

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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.

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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

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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

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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 Science–Policy Interface

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