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Plant-herbivore dynamics associated with an erupting ungulate population : a test of hypotheses Larter, Nicholas C. 1994

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PLANT-HERBIVORE DYNAMICS ASSOCIATED WITH AN ERUPTING UNGULATE POPULATION: A TEST OF HYPOTHESES By NICHOLAS C. LARTER B.Sc.’s University of Alaska, Fairbanks, 1984 M.Sc. University of British Columbia, 1988 A THESIS SUBMITTED IN PARTIAL FULFILMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY in THE FACULTY OF GRADUATE STUDIES (Department of Zoology)  We accept this thesis as conforming to the required standard  THE UNIVERSITY OF BRITISH COLUMBIA June 1994 © Nicholas C. Larter, 1994  In presenting this thesis  in partial fulfilment of the requirements for an advanced  degree at the University of British Columbia, I agree that the Library shall make it freely available for reference and study. I further agree that permission for extensive copying of this thesis for scholarly purposes may be granted by the head of my department  or  by  his  or  her  representatives.  It  is  understood  that  copying  or  publication of this thesis for financial gain shall not be allowed without my written permission.  (Signature)  ZbCLCC ‘  Department of  The University of British Columbia Vancouver, Canada ‘7  Dte-\  DE-6 (2/88)  /  /(/  ABSTRACT This study tests some of the predictions made by two competing hypotheses of plant community and herbivore dynamics: the equilibrium hypothesis, both the original four-stage model proposed by Riney (1964) and Caughley (1970a), and the two-stage model proposed by Sinclair (1979) and Houston (1982), and the facilitation and feedback hypothesis (McNaughton, 1979). This is one of the rare occasions where these predictions have been tested on an erupting indigenous herbivore population subjected to predation; the Mackenzie wood bison (Bison bison athabciscae) population which was reintroduced in 1963. Recently the population split into two subpopulations: the Mackenzie Bison Sanctuary (MB S) which is stabilizing, and the Mink Lake (ML) which is increasing. The aims of this study were to determine 1) if the plant community dynamics were consistent with either of the hypotheses, 2) if herbivore dynamics and demographics were consistent with either of the hypotheses, and 3) the impact of predation on the system. Net primary production of sedges and grasses in areas of willow savannas that were excluded from grazing was similar in MBS and ML, but the standing crop in areas not excluded from grazing was consistently lower in MBS than ML. This difference appears to be a direct result of different grazing pressures. Species composition of willow savannas in MBS had more unpalatable and less preferred species than savannas in ML. These results were consistent with the predictions and assumptions of both models of the equilibrium hypothesis but not the facilitation and feedback hypothesis. Forage quality was not different between MBS and ML with the possible exception of higher levels of some cations found in forages growing in MBS. The two subpopulations appear to be distinct and are at different stages of eruptive oscillation. The demographic characteristics of these subpopulations agree with the predictions of the four-stage model of the equilibrium hypothesis: the MBS subpopulation experienced a negative instantaneous growth rate (r), animals in ML had significantly (p<O.OO3) higher faecal nitrogen levels (an index of diet quality and animal condition), and animals in ML had winter diets containing superior quality items than did animals in MBS. One exception to the predictions was summer diet quality. Animals in MBS had summer diets of superior quality. I argue that this exceptional result is related to predation. The four-stage model of the equilibrium hypothesis assumes that predation will not affect the predictions, but has not previously been tested in systems with a predator. During the past 20 years bison numbers have increased, especially in MBS, while moose numbers appear to have decreased. Moose density in MBS was half that of ML. Bison and moose constituted the majority of the wolf diet, based upon wolf scat analysis. Bison made up a larger proportion of the wolf diet in MBS than in ML, however the occurrence of moose in scats was significantly greater (p<O.O0l) than expected given the availability of prey biomass in both MBS and ML. Wolf activity was greater in MBS than in ML. Given that moose make up a similar proportion of the diet in both areas, and there was a two-fold difference in moose densities between area, wolf predation may be destabilizing and exacerbating the decline in moose numbers. 11  Table of Contents Abstract Table of Contents List of Tables List of Figures Acknowledgements  II 111  .  vi vii viii  CHAPTER 1. General Introduction: the Eruption of Animal Populations  1  The Mackenzie Wood Bison Population Hypotheses on the Dynamics of Plant-herbivore Interactions Hypothesis L: Equilibrium Hypothesis 2: Facilitation and Feedback Hypothesis 3: Range Management Overgrazing Population Eruptions: A Historical Perspective The Study Population  2 3 3 5 6 7 9  CHAPTER 2. How Do Plant Communities Respond to Erupting Herbivore Populations?  15  Introduction Study Area Methods Forage Quality Snow Free Season Snow Season Fibre Analysis Lignin Analysis Nitrogen Analysis Carbon Analysis Elemental Analysis Forage Quantity Grazing Effects in Willow Savannas Forage Composition of Meadows Results Forage Quality Fibre Content Lignin Content Nitrogen Content Carbon Content Other Elements  15 19 20 20 20 21 22 23 24 24 25 26 27 27 28 28 28 30 30 33 35 ill  Forage Quantity Grazing Effects in Willow Savannas Forage Composition of Meadows Discussion Forage Quality Fibre Content Lignin Content Nitrogen Content Carbon Content Other Elements Forage Quantity Grazing Effects in Willow Savannas Forage Composition of Meadows Summary  .  35 35 39 42 42 42 43 43 44 44 45 46 47 50  CHAPTER 3. Do Grazing Mediated Changes in Forage Dynamics Affect Herbivore Demographics? 51 Introduction Methods Herbivore Demography Population Estimates Recruitment and Juvenile Survival Adult Survival Bison Diet Diet Quality Faecal Nitrogen Condition Index Dispersal Results Herbivore Demography Population Estimates Recruitment and Juvenile Survival Adult Survival Bison Diet Diet Quality Faecal Nitrogen Condition Index Dispersal Discussion Demographics Diet Quality and Composition Faecal Nitrogen Condition Index Dispersal Summary .  .  51 54 54 54 58 60 61 62 63 64 65 65 65 68 68 70 70 73 73 76 76 78 82 84 86  iv  CHAPTER 4. The Response of Predators to an Erupting Prey Base Introduction Methods Predator Diet Wolf Scat Analysis Wolf Kills Predator Abundance/Activity Prey Abundance Wood Bison Moose Results Predator Diet Wolf Scat Analysis Wolf Kills Predator Abundance/Activity Prey Abundance Wood Bison Moose Discussion Summary  87 87 90 90 90 92 92 93 93 94 95 95 95 98 98 102 102 102 102 111  CHAPTER 5. General Discussion  112  Literature Cited  119  V  List of Tables Table 2.1 Content of various elements in forages  36  Table 3.1 Ratios of calves and yearlings per 100 adult females  69  Table 3.2 Original locations of all radio collared animals  75  Table 4.1 Wolf activity in both areas based upon track counts  100  Table 4.2 Indices of wolf abundance in both areas  101  Table 4.3 Indices of moose abundance in both areas  103  vi  List of Figures Figure 1.1. The study area  10  Figure 2.1 The digestibility of forages from dry and wet meadows Figure 2.2 Lignin content of forages from dry meadows Figure 2.3 Percent nitrogen content of forages from dry and wet meadows  32  Figure 2.4 Carbon:nitrogen ratio of forages from dry and wet meadows  34  Figure 2.5 Standing crop of biomass from dry and wet meadows  37  Figure 2.6 Standing crop of biomass from grazed and exciosed plots  38  Figure 2.7 Species composition of dry meadows in both areas  40  Figure 2.8 Species composition of wet meadows in both areas  41  Figure 3.1 Population growth of the population and each subpopulation  66  Figure 3.2 The relationship between instantaneous growth rate and population density 67 Figure 3.3 Diet composition of wood bison from both subpopulations  71  Figure 3.4 Index of diet quality for both areas  72  Figure 3.5 Faecal nitrogen content for animals from both subpopulations  .  .  74  Figure 4.1 Diet composition of wolves from both areas  96  Figure 4.2 Available prey biomass in both areas  97  Figure 4.3 Proportion of prey types found at wolf kills in both areas  99  vii  Acknowledgements This thesis represents the culmination of a research program that was initiated when I first set foot in the Mackenzie Bison Sanctuary, in May 1986, to begin my Master’s research program. The initial research program subsequently expanded, almost paralleling the expansion of the wood bison population it was endeavouring to study. Consequently the number of people and agencies involved with this research have ever increased. Without the almost unlimited support of manpower and the generous support of various funding agencies, I would never have been able to complete a project such as this. I am eternally grateful to all those who have in any way shape or form helped me collect, analyze, formulate and present my findings in the following pages. The Bank Many agencies provided funding for this project. They include World Wildlife Fund-Canada, the Department of Renewable Resources Government of the Northwest Territories, the Department of Indian and Northern Affairs, Environment Canada, the Northern Heritage Society/Science Institute of the Northwest Territories, and the Natural Sciences and Engineering Research Council, in an operating grant to Tony Sinclair. To all these agencies I am most grateful. My personal funding came from the Natural Sciences and Engineering Research Council, the Northwest Territories Department of Renewable Resources, and scholarships from the Arctic Institute of North America, the Association of Canadian Universities for Northern Studies, and the Northwest Territories Wildlife Federation. The University of British Columbia provided funds in the form of teaching and research assistanceships. -  The Other Major Players The Physical Asnect I salute the international melting pot of field assistants I had the good fortune to work with, Martin Baumann, Carey Bergman, Darron Collins, Troy Ellsworth, David Hik, Lee Keary, Edward Landry, Walter Landry, Larry Penner, Benjamin Szemkus, Richard Weir, and Anita Young, for their efforts above and beyond the call of duty under extremely rigorous field conditions. Maureen Evans and Gilles Galzi of the Department of Animal Science Lab were highly cooperative in providing expert advice and with the help of lab assistants Susan Brown, and K. Sivakumaran made the lab analyses run smoothly and efficiently. Without thier help I would still be in the lab. To Air Providence and the Malewski family in Fort Providence and Renewable Resources personnel Tom Chowns, Al Helmer, Evelyn Krutko, many thanks for all the logistic support and friendly advice. Special thanks to Barb, radio operator extraordinaire, for making up believable excuses that permitted extra helicopter and other aerial support to come our way. We knew you’d never leave us stranded for too long. ...  VII’  The Mental Aspect To my mentor, supervisor, and good friend, Tony Sinclair, thank you for standing by me through thick and thin. I’m sure you got more than you ever bargained for when you took me on. Your and Anne’s never-ending encouragement kept this thesis progressing and never allowed me to get bogged down. It was a privilege to be part of the Sinclair lab. To my committee members, Corm Gates, Doiph Schiuter, David Shackleton, and Jamie Smith, the ecology group faculty and students, thank you for providing a stimulating and friendly environment conducive to research excellence. No matter how hard the University bureaucracy tried to distract our focus from the research, a few hours in our hallowed huts brought the focus back. Animated conversations with the hut-dwellers and interactions with J. Molson, C.J. Walters, D. Equis, and J. Cuervo often provided me new perspectives on my research when everything seemed to be overwhelming. 0. Harrison er al., K. Richard et at., R. Plant, and P. Floyd are graciously acknowledged for their countless inspiring words. Tinkering with AC/DC not only permitted me to power my laptop in an isolated field camp, but also to power my way through writer’s block. Even when things appeared to be in dire Straits, R. Emmet provided the emotional uplift that allowed me to Triumph and keep going. I especially want to thank my parents, Syd & Pat, my siblings and their better halves Frances & George, Jon & Kristin, and Kate & Scott, some most notable hut-dwellers Dana Atagi, Elvira Harms, David Hik, Mats Linden, Fritz Mueller, Joel Sawada, Tony Sinclair and Carl Walters, and two very special members of my Al-Anon family group Gary D. and Susan F. To you all go my heartfelt thanks for unwavering support during the personal crisis which seriously threatened the write up stages of this work. This thesis would never have been completed without your endless support and encouragement. Last, and by no means least, to two of the sweetest darling girls, Amanda and Laura, who were my best lab helpers ever!!! Thank you for your never-ending curiosity and interest in what I was doing, and for patiently giving me the time to complete my studies even when it cut into our time together I love you both your surrogate Dad. --  ...  ix  CHAPTER 1. GENERAL INTRODUCTION: THE ERUPTION OF ANIMAL POPULATIONS  Populations occasionally undergo large fluctuations that cannot be explained by annual variation in environmental conditions.  In the mammalian literature such rapid  increases have been termed eruptions (Caughley, 1970a). The increase is often, but not always, followed by a marked decline prior to the population stabilizing at a new equilibrium. The decline in the rate of increase as the population stabilizes is caused by three main density dependent factors: food availability, predation, and disease or parasites acting singly or in combination.  Food availability has been documented as a major  regulating factor by Bobek (1977), Sinclair (1977), Sinclair et a!. (1985), Skogland (1985), Clutton-Brock el a!. (1987), Choquenot (1991), and Clutton-Brock el a!. (1991); predation by Mech and Kams (1977), Messier and Crete (1985), Gasaway el a!. (1990), and Messier (1993); and disease or parasites by Christian et a!. (1960), Sinclair and Norton-Griffiths (1979), and Melton and Melton (1982). Eruptions differ from the cyclic increases shown in some small mammal and predator populations (e.g. hares, lemmings, lynx) because eruptions are unpredictable. However, the underlying factors causing the population upswing in both cyclic and eruptive situations may be similar. Most ungulate eruptions have been associated with either the introduction of exotic species into new habitats (Holloway, 1950; Caughley, 1970a; 1987) or onto isolated islands with little grazing history and no predators (Scheffer, 1951; Woodgerd, 1964; Klein, 1968; Leader-Williams, 1980).  The cause of these population  increases is simple and relatively well understood: there is a superabundance of suitable  habitat and hence abundant resources per individual. In contrast, the underlying cause of eruptions of native species in their natural habitat is far less understood. There are few well-documented cases (Caughley, 1970a; Sinclair, 1979), and most explanations remain untested hypotheses. The removal of predators is often mentioned as the cause of the population increase (Rasmussen, 1941; Leopold et al., 1947; Martin and Krefting, 1953; Bergerud, 1971; Gasaway, et at., 1990), with the subsequent overutilization of resources contributing to the population crash (Leopold el a?., 1947; Klein, 1970; Jordan eta?., 1971; Peterson, 1977); however, there has seldom been an accurate measure of available food.  The Mackenzie Wood Bison Population  The wood bison (Bison bison athabascae) is a native ungulate of northern boreal regions of Canada. The Mackenzie population was reintroduced to part of its former range northwest of Great Slave Lake in the Northwest Territories, Canada, in 1963. Since being reintroduced, the population has increased in an eruptive manner (Larter, 1988), and the increase has occurred in the presence of wolf predation. This scenario permitted me to study the plant-herbivore interactions associated with different stages of an erupting native ungulate population. My thesis examines the plant and herbivore dynamics associated with this interactive system, and tests predictions generated by various competing hypotheses.  2  Hypotheses on the Dynamics of Plant-herbivore Interactions  Monro (1967) classified plant-herbivore systems into two basic types: noninteractive and interactive. Interactive systems were defined as sytems where herbivores influence the amount of food to subsequent generations.  Caughley (1976a) further  subdivided interactive systems into interferential and laissez-faire systems.  The key  distinction between them being whether or not herbivores interfered with each other’s search for food. In laissez-faire systems herbivores do not interfere with each other’s search for food.  Non-territorial ungulates are an example of this kind of grazing behaviour  (Caughley, L976a). Noy-Meir (1975) explored the theoretical implications of laissez-faire systems when ungulates were held at a constant density, as in farmed systems. Caughley (1976b) looked at laissez-farie systems containing wild ungulates hunted by man. Three competing hypotheses have resulted that make predictions about plant and herbivore dynamics associated with an erupting herbivore population: 1) the equilibrium hypothesis (Riney, 1964; Caughley, 1970a; Sinclair, 1979; Houston, 1982), 2) the facilitation and feedback hypothesis (McNaughton, 1979), and 3) the range management overgrazing hypothesis (Dyksterhuis, 1949; Westoby et a!., 1989).  Hypothesis 1: Equilibrium  Riney (1964) hypothesized that an eruptive oscillation has four stages: 1) a progressive increase in population size in response to the disparity between the number of animals present and the carrying capacity of the environment, 2) a levelling off in animal 3  numbers in response to decreasing forage availability, 3) a decline in numbers because the population has increased beyond the carrying capacity of the environment, and 4) a phase of relative stability with population density remaining lower than peak density because the carrying capacity of the environment has been reduced by the impact of peak population density. As the ungulate population passes through the four stages, he hypothesized that food supplies are reduced and the plant community will become increasingly dominated by more unpalatable species in response to increased grazing intensity. He maintained that whether an eruptive oscillation occurred in an established or newly liberated population, the four stage sequence of events should be the same. Caughley (l 976b) claimed that this pattern is general and should occur: i) in the presence or absence of predators, ii) on continents as well as islands, iii) in areas of high or low plant diversity, and iv) in areas where herbivores are endemic or exotic. Sinclair (L979) and Houston (982) subsequently proposed that the herbivore population could increase and level off at a new higher equilibrium without undergoing a decline. They hypothesized that the population goes from stage I to 4, omitting stages 2 and 3 because the animal population does not increase beyond the carrying capacity of the environment. Both the Caughley-Riney four-stage model and the two-stage model of the equilibrium hypothesis predict the following:  1)  Peak biomass in total vegetation occurs in stage 1.  2)  Plant composition becomes increasingly dominated by unpalatable species from stage 2 onwards.  4  3)  Diet composition and diet quality will be poorer from stage 2 onwards due to an overshoot in the population and a lag in the population response to external factors.  4)  Fecundity and/or survivorship will be greater in stage 1 than in 2 or adult survivorship will be lower in stage 3 than in 4.  They differ in their predictions about ungulate population growth rate prior to stability. The Caughley-Riney four-stage predicts a negative instantaneous growth rate (r) prior to stability whereas the two stage predicts no negative r stage prior to stability in the herbivore population.  Hypothesis 2: Facilitation and Feedback  This hypothesis advocates a positive feedback between the grazer and plant community as the herbivore population increases to a new equilibrium level without undergoing a decline. Hence, there will be no negative r stage prior to stability in the herbivore population.  Grazing activity actually increases the carrying capacity of the  environment to a new level. Defoliation of plants by grazing is hypothesized to stimulate plant growth and productivity, thereby increasing the dominance of palatable species in the plant community (McNaughton, 1979). Plant productivity may further be enhanced by the addition of nutrients via urination and defecation (Kotanen and Jefferies, 1987; Day and Detling, 1990; Holland et a!., 1992). The predictions from this hypothesis are:  5  1)  Peak total biomass of vegetation occurs when the herbivore population is increasing because total biomass declines due to increasing dominance of palatable species which are constantly cropped.  2)  Plant composition becomes increasingly dominated by palatable species as the herbivore population increases. This will be reflected in improved diet composition.  3)  There will be an improvement in diet quality and composition as the herbivore population increases to its new equilibrium. Therefore animals in stage 1 (of hypothesis 1) will have the lowest quality of diet.  4)  There will be no minimum (low point) in fecundity, recruitment, and adult survival, but these levels will be highest when the herbivore population is increasing (stage 1 of hypothesis 1).  Hypothesis 3: Range Management Overgrazing  The range management overgrazing hypothesis (Dyksterhuis, 1949; Westoby el at., 1989) advocates that, in the absence of predation, there is a negative feedback between the grazer and plant community.  This is one of the theoretical outcomes of laisses-faire  systems discussed by Noy-Meir (1975). It hypothesizes that grazing activity drastically changes the plant community, eventually degrading it until only a few unpalatable species remain. Consequently, after an initial increase in, or establishment of herbivore numbers, there is a continual decrease in r of the herbivore population. Without human intervention (culling or supplemental feeding), r is hypothesized to remain negative until a depauperate 6  population remains. In the present case of the Mackenzie wood bison, predator populations are present; therefore I cannot address the predictions generated by this hypothesis, and only test the predictions of the first two hypotheses.  Population Eruptions: A Historical Perspective  Eruptions are not uncommon. They have occurred in a diversity of species and over a wide range of geographic locations. Most well known cases in large mammals have been associated with populations of ungulates. For example, Leopold et a!. (1947) estimated about 100 deer herds in the United States had entered an eruptive fluctuation between 1900 and 1945. Banfield (1949) recorded an eruption of elk (Ceri’us elaphus) in Manitoba. Erupting populations of reindeer and caribou (Rangfer tarandus) on islands have been reported by Scheffer (1951), Klein (1968), Bergerud (1971) and Leader-Williams (1980). Mech (1966) reported an eruption of a population of moose (A Ices alces) on Isle Royale. Moose populations on mainland Sweden have recently undergone an eruptive increase (R. Bergstrom, pers. comm.).  Eruptions of deer populations on the mainland have been  reported in Denmark (Andersen, 1953), New Zealand (Holloway, 1950), and the United States (Rasmussen, 1941; Mohler et a!., 1951; Martin and Krefting, 1953).  The  introduction of domestic sheep (Ovis aries) to western New South Wales in the early 1800’s resulted in a four-stage eruptive oscillation (Caughley, 1987). Muskox (Ovibos moschatus) populations across northern Canada (Le Hénaff and Crete, 1989; Heard, 1992; Nagy et a!., 1993), Alaska (Smith, 1989) and Greenland (Staaland and Olesen, 1992) are currently undergoing an eruptive increase.  Sinclair (1979) reported eruptions in populations of 7  wildebeest (Connochaetes taurinus albojzibatus) and buffalo (Syncerus caffer) in the Serengeti.  Jnawali and Wegge (1993) reported eruptions of greater one-horned rhinos  (Rhinoceros unicornis) in Nepal.  Most ungulate eruptions which followed introductions of species (reviewed by Caughley, 1970a) are consistent with the Caughley-Riney four-stage equilibrium model. However, eruptions of reindeer introduced to St. Paul and St. Matthew Islands, Alaska, were notable exceptions. Both populations crashed to extremely low numbers within years of their introductions and one eventually went extinct (Scheffer, 1951; Klein, 1968). The exaggerated decline and absence of the stability stage (stage 4) has been attributed to the slow regeneration and growth rates of lichen, the main food source, which had been overgrazed (Klein, 1968; 1987). In contrast, reindeer introduced to South Georgia between 1911 and 1925 did not show a population crash (Leader-Williams ci al.,1981). This was attributed to the temporal availability of lichen forage. Lichen, a less grazing tolerant forage, was unavailable throughout much of the winter because of snow conditions, whereas tussock grass was available year round. Tussock grasses have faster growth rates and are more resilient to grazing than lichen (Leader-Williams, 1980). In contrast to introduced ungulates, data on eruptive oscillations of native ungulate populations have been limited to descriptions of changes in population numbers and mass mortality (Rasmussen, 1941; Leopold e al., 1947; Banfield, 1949). Where demographic data are available, the populations have followed the two-stage equilibrium model. Sinclair (1979) documented eruptions of indigenous populations of wildebeest and buffalo in the presence of predators. He attributed the initial eruptions, between 1961 and 1967, to the rapid disappearance of the viral disease rinderpest. Before 1963, rinderpest was a 8  major mortality factor in calves and yearlings of both species. These eruptions were not followed by subsequent population crashes, but a levelling off at new equilibria. A second period of rapid population increase ensued between 1970 and 1978. These increases were attributed to an increase in food resources caused by a corresponding increase in dry season rainfall and hence forage production.  Once again, no subsequent population crashes  accompanied these latter increases, and it is believed that the latest equilibrium points for both populations are a result of intraspecific competition for food (Sinclair et al., 1985; Sinclair, 1989). Houston (1982), documented an eruption of an indigenous population of elk in Yellowstone National Park. In the absence of predators, the population increased from 4000 in 1968 to 12 000 in the early 1970’s and has subsequently stabilized at approximately 18 000 without subsequent population crashes (D. Houston, pers. comm.). Competition for food appears to be the regulatory factor for this population (Merrill and Boyce, 1991).  The Study Population  In 1963, eighteen wood bison (Bison bison athabascae) were released into an unpopulated area of their historic range just north of Fort Providence, Northwest Territories (Fig. Li). As predicted by Riney (1964), the newly liberated population increased in an eruptive fashion. From 1963 to 1989, the population increased exponentially at an overall rate of 20.1% per year. The 1989 estimate was 2405 bison (SE 227) (Gates and Larter, 1990).  Recently, the population split into two distinct subpopulations: the Mackenzie  (MBS), inhabiting the Sanctuary, and the Mink Lake (ML), inhabiting the Mink Lake area 9  Figure 1.1. The study area, Mackenzie Bison Sanctuary and adjacent area, Northwest Territories. The current range of wood bison is denoted by the open diamonds.  10  northwest of the Sanctuary (Fig. 1.1). Only one area of potential bison range in the overall region remains to be inhabited. This is the Mills Lake area 50 km south of Mink Lake, (Fig. 1.1). Small groups of males have occasionally been seen at Mills Lake. Evidence for the limitation of this population is fragmented. The presence of bovine brucellosis (Brucella abortus) has been associated with poor calf production of bison in Wood Buffalo National Park (Environmental Assessment Panel, 1990) and could be a limiting factor in the Mackenzie population. Bovine tuberculosis (Mycobacteriuni bovis) is also present in the WBNP bison population and has been inferred as another limiting factor (Fuller, 1951; Environmental Assessment Panel, 1990). However, neither disease has been found in the Mackenzie bison population (Tessaro, 1988; Tessaro eta!., 1993). Bison in WBNP and the adjacent Slave River lowlands have died from sporadic outbreaks of anthrax (Bacillus anthracis) since 1962 (Choquette et a!., 1972; Environmental Assessment Panel, 1990), with the most recent outbreak in summer, 1992. The first recorded outbreak of anthrax for my study population occurred in August, 1993: 172 animals died (C. Gates, pers. comm.). Larter (1988) demonstrated the potential for winter nutritional stress to limit population numbers in the Mackenzie Sanctuary, and documented that deep snow during the 1987-88 winter caused bison to feed in habitats containing less crude protein/rn . This 2 is consistent with other studies which found that food limitation in winter was the major regulating factor of erupting populations (Riney, 1964; Klein, 1968; Caughley, 1970a; Leader-Williams, 1980). However, a major difference between the Mackenzie wood bison population and the other populations is that predators were present in the former and absent in the latter. 11  Both wolves (Canis lupus) and black bears (Ursus am ericanus) inhabit the bison study area, but data on numbers of either predator are lacking. Little is known about the effect of black bear predation on bison, but wolves can and do kill both juvenile and adult bison (Oosenbrug and Carbyn, 1985; Carbyn and Trottier, 1987; 1988; pers. obs.). The instantaneous rate of population increase for the Mackenzie bison has not been as constant as previously indicated (Calef, 1984): although positive, r has been declining as the population has risen since 1975 (Gates and Larter, 1990) suggesting a density dependent response to some factor or combination of factors.  Coincidental with the  increase in population size has been an increase in bison range. Bison now inhabit an area of approximately 9 000 km . Two conventional models of dispersal could describe the 2 process involved in range increase.  One is “innate dispersal” which is considered  analogous to random diffusion of reproducing molecules (Caughley, 1970b) in response to a predetermined disposition (Howard, 1960).  The other is “pressure dispersal” which  describes the process whereby individuals disperse only when a threshold density is reached (Caughley, 1977). Innate dispersal predicts a linear increase of the range radius over time, whereas pressure dispersal predicts an eventual geometric increase of the range radius (Caughley, 1970b). Range expansion of the Mackenzie bison has not been at a constant rate, but has occurred in pulses (Gates and Larter, 1990), suggesting pressure dispersal. Dispersal has created areas where the bison population should be at different stages of eruptive oscillation. At the core of the range (MES), bison have been present since 1963 and the plant community has been exposed to grazing longest. At the leading edge of the range (ML), bison presence has been recent (since 1980) and the plant community has only recently been exposed to grazing.  I postulate that the animals inhabiting the 12  leading edge of the range will be in the initial stages of the eruptive oscillation (stage 1): the population should be increasing. Animals inhabiting the core of the range should be associated with a population that is stabilizing. This population could be stage 2 or 4 of the eruptive oscillation depending upon whether the Caughley-Riney four-stage equilibrium model or the two-stage equilibrium model holds. This situation provides an opportunity to test two competing hypotheses regarding the dynamics of a plant-herbivore community where an indigenous herbivore population is erupting in the presence of natural predators. Because I am limited to sampling from one core area (MBS), and one leading edge (ML) my results must be considered in light of the problems associated with pseudoreplication (Hurlbert, 1984).  Historical photos and  descriptions indicate similar wet sedge meadow and willow savanna communities were found throughout the study area. The main questions addressed by this study are:  1)  Are the dynamics of the plant community consistent with the predictions of one of the two competing hypotheses: 1) the equilibrium (Caughley-Riney four-stage or the two-stage), or 2) the facilitation and feedback?  2)  Are the dynamics and demography of the wood bison population consistent with the predictions of one of the two competing hypotheses?  3)  What is the impact of predators on the dynamics of this system?  In chapter two, I examine the dynamics of the plant community in response to an increasing herbivore population, and test the predictions made by the competing hypotheses in regard to plant dynamics. I compare the species composition of willow savannas and 13  wet sedge meadows between areas subjected to different grazing histories. I compare the standing crop of forages in these habitats over the course of the growing season, and determine whether differences in standing crop result from different grazing intensities. Additionally, I assess forage quality by comparing levels of nitrogen, carbon, fibre, lignin, and a variety of other elements measured during the course of the growing season in order to quantify that similar forages growing in different parts of the range are of similar quality at any point in time. In chapter three, I test the predictions generated by the competing hypotheses in regard to herbivore dynamics and demography.  I compare population trends, juvenile  survival and recruitment, diet composition, diet quality, and condition indices between areas. Chapter four treats the predator trophic level. I examine the response of predators to an erupting prey base and determine the impacts of this response on the two alternate prey populations: wood bison and moose. I compare wolf activity, wolf diet, and wolf kills over different prey densities in order to determine if wolf predation on moose is acting in an inversely density dependent fashion. In chapter five, I present a general discussion and summarize my conclusions.  14  CHAPTER 2. HOW DO PLANT COMMUNITIES RESPOND TO ERUPTING HERBIVORE POPULATIONS?  INTRODUCTION  Riney (1964) hypothesized that eruptive oscillations of herbivores go through four stages prior to reaching an equilibrium level: 1) a progressive increase in population size in response to the disparity between the number of animals present and the carrying capacity of the environment; 2) a levelling off in animal numbers in response to decreasing forage availability, 3) a decline in numbers because the population has increased beyond the carrying capacity of the environment, and 4) a phase of relative stability with population density remaining lower than peak density because the carrying capacity of the environment has been reduced by the impact of peak population density. He hypothesized that the underlying cause of this oscillation was a change in the plant community dynamics caused by grazing: food supplies would be reduced and the plant community would become increasingly dominated by more unpalatable species.  He maintained that whether an  eruptive oscillation occurred in an established or newly liberated population the four stages should be the same. Large grazing animals can affect vegetation in a variety of ways, from altering species composition to affecting productivity and nutritional quality. The degree to which changes in the plant community occur seems closely related to the intensity of grazing pressure. In an early review of the effects of grazing on rangeland communities, Ellison (1960) concluded that there was no evidence supporting the claim that grazing had anything 15  but a negative impact on vegetation. Since that review, there has been active debate in the literature as to whether or not grazing is beneficial to plants (see review by Beisky, 1986). For example, Botkin et at. (1981) found species composition and nutrient flow patterns in plant communities were strongly altered by the presence or absence of herbivores: herbivore presence enhanced nutrient cycling. Prins et at. (1980) and Cargill and Jefferies (1984) found that areas of plant communities grazed at moderate levels tended to be more productive than either ungrazed or heavily grazed areas of the community, Contrastingly, Sims and Singh (1978) and Lacey and Van Poollen (1981) found grazing decreased net above-ground primary productivity. The accumulation of dead and live plant tissue in ungrazed areas reduces productivity (McNaughton, 1983; Cargill and Jefferies, 1984). Because dead or senescent tissue is of lower food quality (Stoner er at., 1982; McNaughton, 1985), ungrazed areas may also provide lower quality forage than similar grazed areas. Two competing hypotheses have been proposed to explain the plant and herbivore dynamics associated with erupting herbivore populations that are subject to predation: the equilibrium, both the original four-stage model (Riney, 1964; Caughley, 1970a) and the subsequent two-stage model (Sinclair, 1979; Houston, 1982), and the facilitation and feedback (McNaughton, 1979).  The equilibrium hypothesis generates the following  predictions about initial plant community dynamics: 1)  peak total biomass in vegetation occurs when the herbivore population is increasing (stage 1),  2)  plant composition becomes increasingly dominated by unpalatable species from stage 1 onwards in response to increasing grazing pressure.  16  The facilitation and feedback hypothesis generates these predictions: 1)  peak total biomass in vegetation occurs when the herbivore population is increasing (stage 1),  2)  plant composition becomes increasingly dominated by palatable species from stage 1 onwards in response to increasing grazing pressure.  Although prediction 1 is the same for these hypotheses, the underlying assumptions behind this prediction differ.  The facilitation and feedback hypothesis assumes that grazing  stimulates plant growth and productivity.  Therefore, even though plant productivity  increases from stage 1 onwards, the standing crop of biomass decreases because the palatable species, which are becoming increasingly dominant, are constantly cropped. The equilibrium hypothesis assumes that grazing has a negative effect on plant growth and production.  Therefore, declining biomass from stage 1 onwards results from both  decreasing plant production and increased grazing pressure with more of the total available plant biomass being cropped. Few studies have documented the plant community and herbivore dynamics associated with an ungulate eruptive oscillation.  Most studies have focussed more on  herbivore dynamics than on plant community dynamics. It has often been implied that overgrazing is associated with the stabilization of the herbivore population, and consequently plant biomass is lower when the herbivore population stabilizes than when it was increasing (Holloway, 1950; Riney, 1964; Mech, 1966). However, available forage was never measured.  Where data on plant community dynamics were available, the  dynamics followed those predicted by the equilibrium hypothesis, with decreased total plant biomass and/or increased presence of unpalatable species. 17  Klein (1968) found a decrease in lichen biomass, and an increase in unpalatable species between 1957 and 1963 on St. Matthew Island as a reindeer population passed from stage 1 to stage 2. Caughley (1970a) described a marked decrease in forage biomass in areas where a Himalayan thar (Hem itragus fern lahicus) population was in stage 2 compared to areas where the population was in stage 1. Also, he found an increase in the occurrence of unpalatable species in areas where the thar population was in stage 2 or 3 compared to areas where the population was in stage 1. Kightley and Smith (1976) found decreased forage biomass in areas where a reindeer population was in stage 3 compared to areas where the population was in stage 1 (Leader-Williams, 1980). Jarman and Johnson (1977) documented increases in the dominance of unpalatable plant species in pastures from stage 1 onwards during eruptions of a variety of exotic mammal populations introduced to Australian pasture land. All these studies examined either populations of animals liberated onto islands with little historic grazing pressure by ungulates, or exotic species introduced into a new habitat. Data from eruptions of native species on mainlands are lacking. The Mackenzie wood bison (Bison bison athabascae) population is undergoing an eruptive increase following its liberation in 1963 (Gates and Larter, 1990). The Mink Lake (ML) subpopulation, representing the leading front of the population, is in the increasing population phase (stage 1), and the Mackenzie Bison Sanctuary (MBS) subpopulation, having not as yet undergone a marked decline, is in the stabilizing population phase (stage 2). Therefore, the plant communities located in MBS and ML should reflect the impact of the different grazing pressures exerted on them, and permit me to test the predictions of the competing hypotheses.  18  In this chapter, I compare sward species composition, grazing effects, forage quantity, and a variety of forage quality measures between plant communities in MRS and ML.  STUDY AREA  The Mackenzie wood bison herd inhabits an area exceeding 9 000 km , which 2 includes the Bison Sanctuary (MBS) and adjacent areas located on the western side of Great Slave Lake (61° 30’ N, 117° 00’ W) in the Northwest Territories (Fig. 1.1). It is located in the Upper Mackenzie Section of the Boreal Forest Region (Rowe, 1972) in the emerged bed of a once vast glacial lake. Glacial action has resulted in a nearly level but undulating topography. Post-glacial isostatic lift has lowered the water table and dried the shallow lakebeds in the area. These lakes are gradually filling in with sedges and grasses, while woody plants, notably willows (Salix spp.), are invading the lake margins. Three main open habitat types associated with the shallow lakebeds were identified: (i) wet sedge meadow, (ii) willow savanna, and (iii) sparsely vegetated rocky marl. Wet sedge meadows (WM) were found where there was year-round standing water. They contained high biomass stands of sedge, dominated by Carex atherodes and C. aquatilis.  The willow savanna (DM) habitat was located in more mesic areas often  bordering wet sedge meadows. This habitat consisted of a grass-sedge community with Calamagrostis spp., Agropyron trachycaulum, and Carex atherodes dominating the willow savanna plant community. Hordeurn jubaturn, Phalaris arundinacea, A grostis scabra, and  Carex aquatilis were also present. Willow represented only a minor portion of the ground 19  cover.  Sparsely-vegetated mans were typified by a rocky, calcareous substrate which  supported low plant densities. Grasses, sedges, forbs, and shrubs were widely scattered throughout. A fourth habitat, willow-aspen, was identified in the transition between willow savanna and forested habitats.  Willows and trembling aspen (Populus tremuloides)  dominated this habitat. Understory plant associations were similar to those found in the willow savanna. The main forest habitats identified were coniferous forest, dominated by black spruce (Picea manana), white spruce (P. glauca), jack pine (Pinus banksiana, and larch (Larix laricina); mixed deciduous-coniferous forest, dominated by white spruce, aspen, and  balsam poplar (P. balsam ifera). Two coniferous forest habitat types were recognized: dry and wet types. Dry coniferous forest was dominated by jackpine and white spruce and the wet type was dominated by larch and black spruce. Birch bog dominated by shrub birch (Betula glandulosa) was found in open, poorly drained areas of the forest.  METHODS  Forage quality  Snow Free Season Forage samples were gathered four times in both 1989 and 1990: early June, mid July, mid-August and mid-September. Larter (1988) characterized these times as the start of the growing season, the peak of the growing season, the middle of plant senescence, and peak senescence before snow cover.  Slough sedge (Carex aiherodes), water sedge (C. 20  aquatilis), grass (Calam agrostis spp., A gropyron trachycaulurn, Horde urn jubaturn, A grostis  scabra, and Elymus innovatus), and willow were the forage types sampled. One composite sample of each forage type was collected from willow savanna habitats at ML and the MBS.  Only sedges were collected from wet sedge meadow  habitats in each area because grass and willows were not present. Composite samples of sedge and grass were collected as follows. Numerous individual plants (constituting a minimum 35 g ground dry weight sample), were clipped at random from areas directly adjacent to permanently marked line transects (2 transects per habitat per area except for 3 transects in the willow savanna at MBS). Current yea?s growth was clipped at 3 cm above ground level to represent forage available to grazing bison (Larter, 1988). Composite willow samples were collected in similar fashion. Current annual growth was clipped, with accompanying leaves. Each sample was dried at approximately 60°C in a propane oven for 12 hours, and ground in a cyclone mill through a 1 mm screen. A 1 g subsample of the ground forage material described above was used for each of the forage quality analyses. Samples were analyzed for their nitrogen, carbon, element, fibre, and lignin content in order to determine whether there were differences in quality between area.  Snow Season Forage samples were collected in mid-August, November, mid-February, and early April of 1991-92.  Samples were collected as described above (snow free season).  Different abiotic conditions associated with the growing season may profoundly affect forage quality on a yearly basis (Bell, 1982; East, 1984; Boutton et at., 1988; Bø and Hjeljord, 1991). This would be reflected in the quality of forage at the start of the snow 21  season. Therefore it was necessary to include the mid-August sample as a reference point for the rest of the winter samples. Because the winter diet is almost exclusively sedge (Larter, 1988), only slough sedge (Carex aiherodes) and water sedge (C. aquatilis) were collected. Samples were prepared as above for forage quality analyses. These samples were analyzed for their fibre, nitrogen, and carbon content only in order to address area differences.  Fibre Analysis Forage fibre analyses were determined by the in vitro acid-pepsin digestibility technique following Tilley and Terry (1963) and Spalinger (1980). Larter (1992) found that this simple method gave an index of forage quality comparable to that of the more complicated acid-detergent fibre analyses (Van Soest, 1967). The acid-pepsin inoculant consisted of 2 g of pepsin (Pepsin a, 1: 10,000 Sigma Chemical Co.) dissolved in a mixture of 8.33 ml of 12N HCI and enough water to make 11 of liquid. 0.2 g of forage and 20 ml of acid-pepsin were placed in 20x150 mm test tubes. The tube contents were mixed, by carefully rotating the test tubes by hand, to saturate the forage and then placed in a 37°C water bath for 48 hours. Tube contents were mixed again at one hour, 5.5 hours, and 24 hours after the initial time of incubation. After the 48 hour incubation period the test tube contents were vacuum filtered and the remaining particulate matter was dried for 48 hours at approximately 100°C. The percent fibre digested (%DF) was determined by:  %DF  =  (1  -  (remaining particulate matter/0.2))  *  100.  22  Mean percentages digested, from four runs of each forage sample, are reported. High digestibility values indicate low fibre content and vice versa (Larter, 1992). I normalized the mean percent digested using a square root transformation and then used a five-way ANOVA to determine which factors, if any, explained significant variability in mean index of digestibility during the snow free season. The five factors were forage type, sampling time in the growing season, meadow type (willow savanna or wet sedge), meadow location (MBS or ML), and year. meadow type  (F(I  84)’  Because year (F( 1  p>O.44), and meadow location (F( 1  84)’  84)’  p>O.25) did not explain  significant variation in digestibility of forages collected during the snow free season, the data were lumped across year, meadow type, and meadow location and reanalyzed using a two-way ANOVA. I used a four-way ANOVA (no year factor) to determine which factors, if any, explained significant variability in mean index of digestibility during the snow season. Because meadow type (F( 1  23)’  1 1), and meadow location (F . 0 p> 1  23)’  p>O,26) did not  explain significant variation in digestibility, and only two forage classes were being compared, I lumped the data and used a one-way ANOVA with forage type as the factor.  Lignin Analysis Percent lignin content was determined by the acid-detergent lignin technique (Van Soest, 1963).  This analysis was conducted by the Department of Animal Science  Laboratory, University of British Columbia. The analysis was limited to C. atherodes and willow samples only.  Each sample was run through this technique once.  Duplicate  samples (n=29) were run to determine precision of the measurement which was 92.4%. 23  I used a five-way ANOVA, with the same five factors as above, to determine which factors, if any, explained significant variability in lignin content. Because meadow type ), 137 (F(  p>O.95), year (F( ), p>O.78), and meadow location (F( 137 ), p>O.31) did not explain 137  significant variation in lignin content, and only two forage classes were being compared, I lumped the data and used a one-way ANOVA with forage type as the factor.  Nitrogen Analysis Nitrogen content was determined by the micro-Kjeldahl technique (Nelson and Sommers, L973).  This analyses was conducted by the Department of Animal Science  Laboratory, University of British Columbia. Each sample was run through this technique once. Duplicate samples (n=63) were run to determine precision of the measurement which  was 98.3%. Percent crude protein was determined by the standard conversion (6.25 x Percent Nitrogen). I used a five-way ANOVA, with the same five factors as above, to determine which factors, if any, explained significant variability in summer nitrogen content, and a four-way ANOVA (no year factor) for assessing variability in winter nitrogen content.  Because  meadow location did not explain significant variability in either summer (F( 1  p>O.55),  or winter nitrogen content (F()  25)’  84)’  p>O.24), the data were lumped across meadow location  and reanalysed using four-way and three-way ANOVA’s respectively.  Carbon Analysis Carbon content was determined using an induction furnace carbon analyzer (LECO). 50 mg of sample was heated to 3000°C in an induction furnace; this liberates CO 2 from the  24  combustion. Measurement of the liberated CO, provides an estimate of percent carbon of the forage sample. Each sample was run through this technique once. Duplicate samples (n=9) were run to determine the precision of the measurement which was 98.6%. The carbon:nitrogen ratio (C:N) was determined by dividing the percent carbon by the percent nitrogen. I used a five-way ANOVA, with the same five factors as above, to determine which factors, if any, explained significant variability in carbon content and C:N.  Elemental Analysis The content of certain elements was determined using instrumental neutron activation analysis (INAA) (Bortolotti and Barlow, 1985).  Briefly, the dried, ground  samples were irradiated by a neutron flux, the gamma rays resulting from their artificially produced radio-isotopes were counted using a gamma ray spectrometer, and the gamma ray peaks converted into chemical concentrations.  This analysis was conducted using the  SLOWPOKE reactor at the University of Toronto. Forages were analyzed for the following elements: aluminum, magnesium, potassium, manganese, chloride, and calcium. Results are expressed in parts per million (ppm). I used a five-way ANOVA for each element except calcium, with the same factors as above, to determine which factors, if any, explained significant variability in elemental content. Because of analytical problems, the sample size for calcium was reduced and included samples only from limited time periods during one year. Therefore, I used a three-way ANOVA for calcium content with forage type, meadow location, and meadow type as the factors.  For potassium, magnesium, and aluminum content there were no year 25  ), 179 (F(  p>O.45 for potassium;  ), 177 F(  p>0.9l for magnesium; F( ), 179  p>O.33  for aluminum),  ), p>O.3l for potassium; F( 379 sampling time (F( ), p>O.58 for magnesium; F( 377 ), p>0.075 379 for aluminum), or meadow type (F( 1 1 magnesium; F  79)’  79)’  p> 6 . 0 2 for potassium; F(  p> 2 . 0 5 for aluminum) effects.  77).  p> 9 . 0 7 for  Therefore, I lumped the data across  forage type, meadow type, sampling time and year and used a one-way ANOVA with meadow location as the factor.  Forage Quantity  At the start of the 1990 growing season, two permanent line transects 200 m in length were set up in wet sedge meadow habitat in ML and the MBS. Concurrently, three permanent line transects were set up in willow savanna habitat at MBS (300 m, 300 m, 225 m long respectively) and two in willow savanna habitat at ML (500 m long). For each habitat and area combination, ten 0.25 m 2 plots were randomly located on these transects. All forage was clipped at a height of 3 cm aboveground to represent forage available to grazing bison (Larter, 1988). Plots were clipped at the same time as forage samples were collected: early June, mid-July, mid-August, and mid-September. Forage was dried at 60°C for 12 hours and subsequently weighed on an electronic balance. Dry weight of forage was reported. I used the Student’s t-test to compare standing crop between areas.  26  Grazing Effects in Willow Savannas  Five exclosures were erected in willow savanna habitats in each of ML and the MBS during July 1991. Each exclosure consisted of a 2.3 m x 1.7 m piece of 5 cm x 5  cm chainlink fencing. Once pegged into the ground it resulted in a domed exciosure. At the time the exciosures were erected, two 0.25 m 2 plots were clipped outside each exciosure, the position being 1 m from a randomly chosen corner.  At the end of the  growing season two 0.25 m 2 plots were clipped inside the excluded area and two 0.25 m 2 plots were clipped 1 m from the corner of the exciosure that had not been previously clipped. Clipped samples were dried at 60°C for 12 hours, and weighed on an electronic balance.  Dry weight of forage was reported.  I used the Student’s t-test to compare  standing crop between excluded and grazed areas.  Forage Composition of Meadows  The composition of both willow savannas and wet sedge meadows was determined using two methods: the point-intercept method (Kershaw, 1973), and by presence/absence records from the clip plots used to estimate forage quantity.  For the point intercept  method, I used the same permanent line transects described previously (see Forage Quantity). Ten points were randomly located on these transects. At each point a metre stick was placed perpendicularly to the transect. A nail, representing a point, was placed into the meadow substrate at 10 cm intervals along the stick. Any forage plant that was intercepted by the nail was recorded. The forages recorded were: slough sedge (Carex 27  atherodes), water sedge (C. aquatilis), grass, forb (non-woody dicotyledonous plants), and reed canarygrass (Phalaris arundinacea). Composition sampling was conducted at the same time as forage quantity and quality sampling. Composition data were lumped across the four sample times. Data were recorded as proportions: number of point intercepts/400 possible and number of plots with the forage type present/40 possible plots. I used the proportion test (Zar, 1984) to compare proportions between area.  RESULTS  Forage Quality  Fibre Content Forage type (F 3  87)’  p<O.Ol5)  and sampling time (F( 3  87)’  p<O.OOl) explained  significant amounts of variation in forage fibre content during the snow free season. The digestibility index generally declined throughout the summer for forages growing in willow savannas, indicating increasing fibre content (Fig 2.1A). There was little change in fibre content of forages growing in wet sedge meadows (Fig. 2.1B).  When the digestibility  index was lumped across sampling time there were significant differences (F( 3  90)’  p<O.0001): mean (±SD) digestibility index of willow (42.9±3.6, n=16) was higher than mean digestibility index of Carex atherodes (3 5.9±3.9, n=30), C. aquatilis (32.8±3.2, n=32) and grass (34.0±8.8, n=16). Sampling time explained significant amounts (F( ), p<O.OOl) of variation in fibre 323 28  0.5  A 0.45  0.4  0.35  0.3  a) Cl) a) c,)  0.25  Q0.2  6JN  15AG  15JL  15SE  B  0.35  0.3  0.25  6JN  I5JL  15AG  15SE  Date  Figure 2. L The general trends in forage digestibility during the snow free period: A) Carex atherodes growing in willow savannas, B) C. atherodes growing in wet sedge meadows. Mean values (n=8) with standard deviations presented. 29  content during the snow season. Fibre content increased in all forages, reaching its highest levels by mid-winter (February). By late winter (April) digestibility had risen somewhat. Sedges growing in wet sedge meadows tended to be more digestible than those growing in willow savannas, but this difference was not significant (F 1 was similar for C. atherodes and C. aquatilis (F()  23)’  p>O.O9). Fibre content  30)’  Lignin Content Forage type (F 1  37)’  p<O.0001) and sampling time (F( 3  37).  p<O.OO1) explained  significant amounts of variation in lignin content. Lignin content of both Carex atherodes and willow increased during the snow free season until August (Fig. 2.2). Absolute lignin content was consistently higher in willow than C. atherodes, regardless of sampling time, and when the data were lumped across sampling time this difference was significant (F( 1 43)’  p<O.0001).  Nitrogen Content Meadow type (F( 1  85)’  p<O.Ol5) and sampling time (F 3  85)’  p<O.0001) explained  significant variation in nitrogen content during the snow free season. Nitrogen content declined in a linear fashion for all forages during the snow free season (Fig. 2.3A), and was higher in forages in willow savannas than wet sedge meadows. Nitrogen content of all forages tended to be higher in 1990 than 1989; the difference was not significant  ), 185 (F(  p>O.O9). Nitrogen content tended to be higher in willow and grass than in sedges. Of the sedges, Carex atherodes had higher nitrogen content until the last sampling date (Fig. 2.3 A). 30  I  I  10 7.5 5 2.5  -  -‘-\ 1 JN  15AG  15JL  15SE  Date MLCAAT  UBSCAAT  MLSASP  MBSSASP  Figure 2.2. Mean (±SE) percent lignin content of Carex atherodes (CAAT) and willow (SASP) found in Mink Lake (ML) and the Mackenzie Bison Sanctuary (MBS) during the snow free season. 31  3.5  2.5  sO  1.5  0  0  0.5  15SE  15AG  15JL  1JN  a)  C.) C 0  SASP  Z1GRASS  CAAT  CAAQ  ()  2.5 C  B  C) 0  2  z  i .5 1  0.5  0  16 AG  16 FE  25 NV  I  1 APR  Date — CAAT DM  CAAQ DM  CAAT WM  CAAQ  Figure 2.3, (A) Percent nitrogen content of forage types, willow (SASP), grass, Carex atherodes (CAAT), and C. aqua/ills (CAAQ), during the snow free season. (B) Percent nitrogen content of C. atherodes (CAAT), and C. aqua/ills (CAAQ) in habitats, willow savannas (DM) and wet sedge meadows (WM), during the snow season. 16 August represents a reference point for late summer nitrogen content. 32  The linear decline in nitrogen content continued through the snow season (Fig. 2.3B).  Sampling time (F( 3  26)’  1 i) and meadow type (F( 000 P<O•  26)’  p<O.O2) explained  significant amounts of variation in winter nitrogen content. Forages in wet sedge meadows had higher nitrogen content than forages in willow savannas during early winter, but similar content by late winter. This resulted in a steeper decline of nitrogen content over the winter for forages in wet sedge meadows.  C. aquatilis had consistently higher winter  nitrogen content than C. atherodes (Fig. 2.3B), but this difference was not significant 26)’  1 (F(  ). 65 i’>°.°  Carbon Content Forage type (Ff 3  84)’  p<O.OO1) and sampling time (F 3  84)’  p<O.O25) explained  significant variation in carbon content. Carbon content decreased over time, however this decrease was less than 1.2%. Willow had higher carbon content than any other forage, however this difference was only 4 percentage points. There were no meadow type (F( 1 84)’  p>O.86), meadow location (F( 1  84)’  p>O.7l) or year (F(J  84)’  p>O.l7) effects. When all  data were lumped (n=94), the mean carbon content was 44.8% with a standard deviation of 2.15%. Because nitrogen content declined during summer, the carbon:nitrogen ratio showed significant sampling time effects (F( ), p<O.OO1), and increased as the summer progressed. 384 Additionally, there were significant meadow effects (F( 1  84),  p<O.O2), with higher  carbon:nitrogen ratios in forages from wet sedge meadows than willow savannas (Fig. 2.4). There were no species (F 3  84)’  p>O.52), meadow location (Ff 1  84)’  p>O.9), or year (F( 1  s4)  p>O.l) effects. 33  Wet Meadow  Dry Meadow  45 1  0 I.  z  T  30  15  0  1JN  15AG  15JL  15SE  Date  Figure 2.4. Changes in mean (±SE) carbon:nitrogen over time and between willow savannas (DM), n=16, and wet sedge meadows (WM), n=8. Data are lumped across forage type and location. 34  Other Elements Magnesium, potassium, and aluminum content were significantly higher in forages from MBS than ML (F( ), p<O.O1; F( 181 ), 183  p<O.Ol; F( ), p<O.O2 respectively) (Table 2.1). 183  No factors explained significant variation for chloride (F( 8 76)’  p>O.58).  76)’  p>O.38) or manganese (F( 8  All elements except chloride and manganese had significant forage type  effects.  Forage Quantity  There were significant differences in standing crop of the different meadow communities between areas (Fig. 2.5). Willow savannas in MBS had a lower standing crop at all sampling times except at the start of the growing season. These differences were significant in both mid-July (1-tail, p<O.0001, t=-7.19) and mid-August (1-tail, p<O.OO1, t 4.16) (Fig. 2.5A). Wet sedge meadows had similar standing crop at all times regardless of location except for mid-August (Fig, 2.5B), when available standing crop was significantly lower in ML (1-tail, p’<O.OOI, t=4.95).  Grazing Effects in Willow Savannas  Standing crop was not different between years for either location, but was different depending upon grazing treatment (Fig. 2.6). Exclosed areas in MBS showed a noticeable response to being released from grazing, the standing crop being about 3 times greater (1tail, p<O.OO4, t=-3.24) in areas where grazers were excluded (Fig. 2.6A).  In contrast, 35  Table 2.1 Mean concentrations (ppm) ± SE of certain elements in four types of forage found in MBS and ML. CAAT SASP  =  willow,  location, n  =  *  =  =  Carex atherodes, CAAQ  significant difference  (p<O.O2)  =  C. aquatilis, GRAS  =  grass,  in elemental concentration between  sample size except for calcium. Values in parentheses indicate sample size  for calcium.  MBS Mg*  K*  A1*  Mn  Cl  Ca  CAAT 12  126±30  3128±895  808±220  96±22  418±52  3230±377 (6)  CAAQ 16  106±2 1  4297±877  842±168  98±17  827±112  5407±1120 (9)  GRAS  6  436±23 1  12623±849 1  2236±155 1  186±107  844±355  5845±5 114 (3)  SASP  7  32±10  917±295  79±168  215±74  525±140  9519±7230 (3)  Forage  n  ML Mg*  K*  A1*  Mn  Cl  Ca  CAAT 13  210±53  3039±648  854±171  127±39  413±13  5378±2767 (8)  CAAQ 15  225±41  7019±1290  1253±241  170±39  898±147  5001±1151 (9)  GRAS  7  1955±640  19292±4484  41 14±986  455±144  393±106  20616±8648(4)  SASP  7  379±105  1573±346  4622±1193  1427±336  152±26  47568±22863(3)  Forage  n  36  Willow Savanna 35 30  **  A  25 ML 20 15 MBS  10  I  5 0  6JN  15JL  15AG  15SE  Wet Sedge Meadow  50  B 4°  30  20  ML 10  6JN  15JL  15 AG  15 SE  Date  2 dry weight) values compared between Figure 2.5. Mean (±SE) standing crop (g/m locations, Mink Lake (ML) and Mackenzie Bison Sanctuary (MBS), for (A) willow savannas and (B) wet sedge meadows. Standard error bars are presented, ** indicates significant differences (p<O.OO1) between areas. 37  a.  Excios e d  MBS  ‘91  ‘90 I.  -C  ‘91  ‘90  0)  C44  E  August  July  >1  b. Excios e d  Cl)  ‘91  E ML  ‘90  ‘90  ‘91  July  August  2 dry weight) of areas where Figure 2.6. Comparison of the mean standing crop (g/m animals were either permitted to graze or excluded from grazing. Data from willow savannas in (A) the Mackenzie Bison Sanctuary, and (B) Mink Lake (ML). Exciosures were erected in July, 1991. ** indicates significant difference between grazed and excluded areas (pcO.OO4). 38  exclosed areas in ML showed only a slight, and non-significant difference (1-tail,  p=O.26,  t=-O.66) in standing crop from areas where grazing occurred (Fig. 2.6B). Exciosed areas in MBS had an August standing crop similar to that found in both the grazed (2-tail,  p=O.57,  t=-O.58) and ungrazed (2-tail, p=O.35, t=-O.96) locations of ML. Therefore, the  exclosures demonstrated that net primary production was similar between areas and the differences in standing crop between areas was a result of different grazing offtakes.  Forage Composition of Meadows  Willow savannas in MBS had significantly more grass (Z=1.98, p<O.O25), forbs (Z=5.82, p<O.0001), and reed canarygrass (Phalaris anindinacea) (Z=2.77, p<O.OO3), and less Carex atherodes (Z2.91, p<O.OO2) and C. aquatilis (Z=2.22, p<O.O14) than willow savannas in ML (Fig. 2.7) when comparing presence/absence data. The trend was the same for the point intercept data, however not all differences were significant (Z=1.46, p<O.O75 for C. atherodes and Z=1.OO, p<O.l6 for reed canarygrass).  Wet sedge meadows had  similar composition between areas with a trend to more C. aherodes in meadows located at ML (Fig. 2.8). This difference was significant (Z=1.98, p<O.O24) for point intercept data, but not so (Z=1.60, p>O.O54) for presence/absence data. The amount of C. aquatilis did not differ between areas regardless of data set (ZO.75, p>O.22 for point intercept and Z=O.11, p>O.45 for presence/absence).  39  120  —  o  80  .  60  M66 ML  100  0 0  I  40  20  Grass  Caat  Caaq  Forb  Phar  Caat  Caaq  ForD  Phar  40  0  30  a)  C.)  a) 0 a) -  20  10  Grass  Forage  Type  Figure 2.7. The differences in species composition of willow savannas between Mink Lake (ML) and the Mackenzie Bison Sanctuary (MBS). (A) Presents data from the pointintercept method. (B) Presents data from the presence absence method, see text for description of methodologies. CAAT = Carex atherodes, CAAQ = C. aquarius, PHAR = Phoiaris arundinacea. 40  50 MBS 40  ML  freeee+e• fr4eee••  o o  30-  Cl)  t I  ta’asw.  n.ee•ee  20-  neeee.••  n.e..... neneeee nenne. nenne  neeeee  neeeeee•  e•eeee  p.&eneee  I0-  n.e.....  neeeee  n..nn.  nee.e.e  neeeeee  neeeee•  Lt•st&tt  0  Caaq  Caat  40  reteeeee  o  i..eeeeei flee •  30  e.e.ee..i eeenni e.eeeeei C. e .. ei .e.e•eei  I. .. .  C,  a, Cl) a, Ia.  i eeeeei b.eeeeeei  +•e•ee4 beeeee•e4  a,  C  ....a.,..e.t  ne•eeee4  tta%fl.%  20  ‘Ce...... iee.eeeea beeeeeea beee•ee.4 ‘Ce...... a.. •  •e.eeeei  ‘Ce...... 10  n.e..... nee•e•  •. . . . eVa  ieee..... eeeeCC4  0  t•iatSt4a,  ha a. a a a.a a  Caaq  Caat  Forage  Type  Figure 2.8. The differences in species composition of wet sedge meadows between Mink Lake (ML) and the Mackenzie Bison Sanctuary (MES). (A) Data from the point-intercept method. (B) Data from the presence absence method, see text for description of methodologies. CAAT = Carex atherodes, CAAQ = C. aqua/ills.  41  DISCUSSION  Forage Oualitv  Forage quality changed on a seasonal basis and varied between different forages, however there were no real differences in the quality of forages between the two areas. At any given sampling time the quality of one type of forage was similar between areas.  Fibre Content Bison are large bulk feeders capable of processing large amounts of low quality forage (Hanley, 1982; Hudson and Frank, 1987), and are capable of digesting low-protein, high-fibre forages more efficiently than are cattle (Peden eta!., 1974; Hawley eta!., 1981a; 1981b; Deliberto, 1993). However, during snow season, nitrogen reached its lowest level of the year (Fig. 2.3B). Crude protein levels of forage at this season were below 5.9%. Such levels are inadequate for the maintenance requirements of dry, pregnant mature beef cows (National Academy of Sciences National Research Council, 1976) and crude protein -  levels below 7% appear to depress rumen cellulolysis (Hawley, et al., 1981a).  Crude  protein levels of sedge varied from 7.8% in November to 3.2% in April. With such low crude protein in the forage, any reduction in its associated fibre content may enhance diet quality.  42  Lignin Content Larter and Gates (1991a) found significantly higher lignin content in willows than in various monocotyledonous forages found in the MBS, but no significant linear increase in lignin content of willows during the course of the growing season. An extremely low lignin content in willows found in ML during the first sampling period produced the appearance of a linear increase of lignin over time. If this outlier is removed there is no significant location or trend over time effect. Whether or not this low lignin value is an artefact of sampling or a real difference between areas remains to be resolved by future sampling. If this difference is indeed real, the quality of willows in ML would be superior to those in MBS at the start of the growing season, because lignin interferes with cellulose digestion (Van Soest, 1963). Therefore the lower lignin at ML would improve the quality of diet for animals in that area (see Chapter 3).  Nitrogen Content Nitrogen, in the form of crude protein is important for the growth and maintenance of herbivores. The nitrogen content of forages was similar over the entire study area for any given forage type at any given time. Crude protein proportions in forage were highest at the start of the growing season and declined throughout the year. The available crude protein biomass was highest in mid-July (Larter and Gates, 1991a). At this time crude protein levels were lower, but the quantity of forage was higher. Wood bison in ML may have had access to forage of similar crude protein content to forages in the MBS, but available crude protein biomass was probably much greater in ML because of the higher standing crop in willow savannas. 43  Carbon Content The carbon:nitrogen ratios were similar between area and varied from approximately 17:1 to 36:1 in willow savannas, and from approximately 23:1 to 42:1 in wet sedge meadows. Net nitrogen mineralization generally occurs in areas where carbon:nitrogen ratios of forage are below than 30:1 (Edmonds, 1980). Carbon:nitrogen ratios of forage in excess of 30:1 are generally indicative of areas experiencing net nitrogen immobilization. Nutrient availability for plants is more limited in areas experiencing net nitrogen immobilization.  Carbon:nitrogen data were not calculated for snow season samples.  However, since carbon levels remained relatively constant over time and nitrogen levels dropped during the snow season (Fig. 2.3B), one would expect that carbon:nitrogen ratios continued to decline over the course of the snow season. Because the carbon:nitrogen ratio exceeded 30:1 by September and would not have increased above this point until after snow season (eight months later), nutrient availability would be limited for most of the year. June to August would be the only time when nutrients would be readily available for plants and plant growth.  This provides a short time period for herbivores to consume  forage with high crude protein levels. Foraging on plants providing the most available crude protein during this time may be critical in order to build up sufficient body reserves to survive through the winter.  Other Elements Magnesium, potassium, and aluminum contents were higher in forages growing in the MBS than those growing in ML. The increased content of these positively charged elements in forages growing in the MBS may be a direct result of the more calcareous, 44  lacustrine substrate. Higher contents of these cations may have affected forage quality, but any noticeable shift in the use of these forages was not evident when comparing previous diet data (Larter, 1988; Larter and Gates, 1991a) with current diet data (see Chapter 3).  Forage Ouantity  Both equilibrium and facilitation and feedback hypotheses predict peak forage biomass occurs at stage 1 (increasing herbivore population). The standing crop of forages in willow savannas located at ML was significantly greater than for willow savannas at IvIBS (Fig. 2.5A). These data are consistent with the predictions of both hypotheses.  There were no significant differences in wet sedge meadow standing crop between areas except for the August sampling time (Fig. 2.5B). This difference appeared to result from unusually heavy rainfall on the Horn Plateau during early August.  Water levels  around Mink Lake rose drastically (1.5 m) during early August, but had subsided by September. Forage availability in wet sedge meadows in ML was directly affected by the change in water levels (Fig. 2.5B.), not by grazing. Wet sedge meadows were essentially ungrazed during the snow free season (Larter and Gates, 1991a; 1991b). Therefore one would not expect to see differences in standing crop between area during this season. During winter, wet sedge meadows were heavily grazed (Larter and Gates, 1991a; 1991b). Therefore one would predict differences between area in standing crop of wet sedge meadows to be similar to those found in willow savannas during the snow free season. Data to test this prediction were unavailable, however, observations during April indicated wet sedge meadows in the MBS were more heavily utilized, and had a much lower 45  standing crop of sedge material than wet sedge meadows in ML (pers. obs.).  Grazing Effects in Willow Savannas  Grazing by herbivores can have a variety of effects on the productivity of plant communities, from an increase in net above-ground primary productivity (NAPP) (Prins et aL, 1980; Cargill and Jefferies, 1984; McNaughton, 1985) to a decrease in NAPP (Sims and Singh, 1978; Lacey and Van Poollen, 1981). This study did not measure productivity, but the hypotheses make predictions about biomass which I tested in this study.  The  underlying assumptions of these hypotheses differ with respect to grazing effects. The equilibrium hypothesiss implies that grazing associated with an eruption of the herbivore population has a negative effect on NAPP. Contrastingly, the feedback and facilitation hypothesis (McNaughton, 1979) implies that grazing has a positive effect on NAPP. The impact of grazing by a high density of wood bison, the situation in the MBS, had striking effects (Fig. 2.6). Even though the exclosures were only in place for 6 weeks, and were erected after at least one month of grazing on the current year’s growth, there was a three-fold increase in standing crop inside the exclosures. The lack of any significant difference in standing crop between areas exciosed and not exclosed from grazers in ML is expected because bison density is low and grazing of the plant community has been a recent occurrence. There was still an abundance of forage in willow savannas in ML (pers. obs.), and bison densities were much lower than in the MBS (see Chapter 3).  The lack  of significant difference in standing crop between areas exciosed in MBS, and areas exclosed in ML indicates that net primary production of willow savannas is similar between 46  areas, and implies that grazing does not have a positive effect on NAPP, contrary to what is implied by the facilitation and feedback hypothesis (McNaughton, 1979).  Forage Composition of Meadows  The equilibrium hypothesis predicts species composition of meadows will become increasingly dominated by unpalatable species as the eruptive oscillation moves from stage 1 (increasing herbivore population) to a stable herbivore population. The facilitation and feedback hypothesis predicts the opposite: species composition in meadows will become increasingly dominated by palatable species.  My data indicate a higher occurrence of  unpalatable species in MBS versus ML presumably as a result of differences in grazing pressure.  These data agree with the predictions of the equilibrium hypothesis, but are  contrary to the predictions of the facilitation and feedback hypothesis. Willow savannas in MBS had significantly more reed canarygrass  (Phalaris  arundinacea), forb, and grass, and significantly less sedge (both Carex atherodes and C. aquatilis) than willow savannas in ML when using presence/absence data.  The trend  remained the same (see Figs. 2.8 and 2.9), but not all differences were significant when using point-intercept data.  This lack of statistical significance may be related to the  physical size of the sample unit. The sample unit of the point-intercept is <1 cm 2 while that of the presence/absence is 0.25 m . Therefore, even though the point-intercept method 2 provides a larger sample size than the presence/absence method, the sample unit it uses is much more susceptible to a zero recording. The resultant smaller proportions created by the point-intercept method may make it more difficult to determine statistical differences 47  between areas, even with an increased sample size, using the proportion test. Reed canarygrass is a highly unpalatable species.  It has a substantial alkaloid  content (Audette eta!., 1970; Hagman et a!., 1975), which increases with foliage removal (Woods and Clark, 1971).  Alkaloid concentrations were negatively correlated with  palatability when offered to sheep and cattle (Marten et al., 1976). Van Soest (1967) has shown P. arundinacea to be high in silicates, another potential plant defense. Silica has many detrimental effects to mammals, ranging from increased tooth wear to silica urolithiasis, which can be fatal to young animals (Baker et a!., 1959; Bailey, 1981; Van Soest, 1982). Gali-Muhtasib et a!. (1992) presented prairie voles (Microtus ochrogaster) with leaves of two grass species grown in the presence or absence of silica, and at high and low nitrogen content.  Leaves of grasses grown in the absence of silica were highly  preferred irrespective of nitrogen differences among treatments. Wood bison avoid foraging on this species (Larter, 1988). There was significantly more reed canarygrass in willow savannas in MBS than ML. Forbs could be a valuable food source because of their generally high digestibility and nitrogen content (Belovsky, 1986; Renecker and Hudson, 1988), but they are rarely grazed by wood bison (Larter, 1988; Larter and Gates, 1991a; 1991b). Smith (1990) found that the majority of forb species present were of the prostrate growth form.  This was  attributed to wood bison grazing which had removed the competitively superior erect graminoids and hence freed resources for prostrate forbs which could escape grazing below 3 cm height (pers. obs.). Forbs were consumed if they reached a sufficient height (Smith, 1990; pers. obs.), but patches of willow savannas with higher densities of forb growth were grazed less frequently than those with lower forb densities (Smith, 1990). Willow savannas 48  in MBS had significantly higher forb densities than those in ML. Grass is generally of equal or superior quality to sedge (see Forage Quality and Larter and Gates, 1991a). However, grass tussocks, unlike sedge plants, come in a mix of live green and previous dead material, which bison with their large mouths cannot sort (Hanley, 1982). Because of this unsorted mix of live and dead material, the grass that is cut and eaten is of lower quality than similar forage of sedge. Willow savannas in MBS have a higher grass component than those in ML. Sedges, in particular C. atherodes, appear to be a preferred and important forage for bison throughout the year, (Reynolds el a!., 1978; Hawley, et a!., 1981a; Larter, 1988; Larter and Gates, 1991a; 1991b).  Willow savannas in ML had a higher occurrence of  sedge, both C. atherodes and C. aquarius, than those in MBS. Although there was a higher occurrence of the preferred sedge (C. atherodes) in wet sedge meadows in ML than MBS, this difference was not significant for both data sets. Wet sedge meadows are primary winter foraging areas, rarely used for summer foraging (Larter, 1988; Larter and Gates 199a; 1991b). Because foraging does not occur during the growing season, removal of one species (C. atherodes) is less likely to provide a competitive advantage to the other species (C. aquarius) during winter. There is no analogy to the competitive advantage gained by forbs from the constant removal of graminoids during the growing season. Possibly winter grazing has a positive effect by removing the dead leaf matter from the previous summer’s growth, so that with the onset of spring there is more light and possibly heat available for the newly growing material. Any changes in sedge species composition would be much more gradual. The carbon:nitrogen ratio of forages in wet sedge meadows is higher than for forages in dry meadows, indicating that 49  wet sedge meadows are more nutrient limited, further restricting any rate of change in sedge species composition of wet sedge meadows.  Therefore, the differences in the  occurrence of C. atherodes’ may just be approaching levels that are statistically detectable.  SUMMARY  L)  The standing crop of sedges and grasses in willow savanna habitats located in the Mackenzie Bison Sanctuary (MBS) was consistently lower than that of willow savannas located in the Mink Lake (ML) area. This result is consistent with the predictions of both the equilibrium and the facilitation and feedback hypotheses.  2)  Net primary production was similar between MBS and ML.  The difference in  standing crop between areas appears to be a direct result of different grazing pressures between the two areas. This implies grazing has a negative effect on NAPP which is consistent with the assumptions of the equilibrium hypothesis, but contrary to the assumptions of the facilitation and feedback hypothesis. 3)  The species composition of willow savannas differed between areas, with savannas in the MBS showing an increase in more unpalatable or less preferred species in relation to willow savannas in ML. This is consistent with the predictions of the equilibrium hypothesis, but contrary to the predictions of the facilitation and feedback hypothesis.  4)  There was no significant difference, between area, in any of the different quality measures examined.  Forages during the snow free season were of comparable  quality in MBS and ML. 50  CHAPTER 3. DO GRAZING MEDIATED CHANGES IN FORAGE DYNAMICS AFFECT HERBIVORE DEMOGRAPHICS?  INTRODUCTION  Two competing hypotheses have been proposed to explain the dynamics and demographics of herbivore populations which are undergoing an eruption in the presence of predation: the equilibrium (Riney, 1964; Caughley, 1970a; Sinclair, 1979; Houston, 1982), and the facilitation and feedback (McNaughton, 1979). Riney (1964) hypothesized that an eruptive oscillation would have four stages prior to reaching an equilibrium level: 1) a progressive increase in population size in response to the disparity between the number of animals present and the carrying capacity of the environment, 2) a levelling off in animal numbers in response to decreasing forage availability, 3) a decline in numbers because the population has increased beyond the carrying capacity of the environment, and 4) a phase of relative stability with population density remaining lower than peak density because the carrying capacity of the environment has been reduced by the impact of peak population density.  Caughley (1970a)  hypothesized that the declining rate of increase from stage 1 to stage 3 and the rising rate of population increase from stage 3 to stage 4 was a result of either differences in fecundity and/or differences in mortality, especially the rate of mortality over the first year of life, in relation to reduced food supplies. The predictions generated from this original four-stage model of the equilibrium hypothesis are:  51  1)  a negative instantaneous growth rate (r) prior to stability in the herbivore population,  2)  better diet quality and hence superior diet composition of animals in stage I than 2.  Sinclair (1979) and Houston (1982) subsequently proposed that the new higher equilibrium density would be reached by a gradual levelling of the population and without a decline phase. In this two-stage model, they hypothesized that the population goes from stage 1 to 4, omitting stages 2 and 3 because the animal population does not increase beyond the carrying capacity of the environment. The declining rate of population increase is associated with decreasing recruitment and increasing mortality.  Therefore this  modification of the original equilibrium hypothesis makes the same predictions except that there will be no negative r stage prior to stability in the herbivore population. The facilitation and feedback hypothesis advocates that as the herbivore population increases to a new equilibrium level, without undergoing a decline, a positive feedback is established between the grazer and plant community. Thus grazing activity modifies the carrying capacity of the environment by increasing it to a new level. Plant defoliation is hypothesized to stimulate plant growth and productivity thereby increasing the dominance of palatable species (McNaughton, 1979).  The predictions from this hypothesis are:  1)  no negative r stage prior to stability in the herbivore population.  2)  an improvement in diet quality and composition as the herbivore population increases to its new equilibrium, therefore animals in stage 1 (of hypothesis I) will have the lowest quality of diet.  Few studies have assessed the direct influence of food shortage on population 52  parameters (Bobek, 1977; Skogland, 1985; Bayliss, 1985; Clutton-Brock et al., 1987, Fryxell, 1987; Clutton-Brock  el  a!., 1991; Choquenot, 1991). Even fewer studies have  investigated the demographic characteristics of an ungulate eruptive oscillation. Klein (1968) found a pronounced decrease in the ratio of both calves and yearlings to adult female reindeer (Ran gifer tarandus) in a five year span before peak population density was reached. A negative  r  followed the peak density, but the population did not  stabilize; it crashed to virtual extinction. Caughley (1970a) studied four subpopulations of introduced Himalayan thar (Hemitragusjemlahicus) and found birth rate, especially those of younger animals, and juvenile survival was higher in increasing (stage 1) than stabilizing (stage 2) subpopulations.  Juvenile survival was the major component dampening  population increase, and a reduction of available food was the ultimate factor affecting rates of increase from stage 1 to stage 3.  Leader-Williams (1980) studied three herds of  introduced reindeer on South Georgia. Each herd was at a different stage of eruption. Calf mortality was higher and sex ratios were biased toward females in the population at stage 2 when compared with the population at stage 1. Demographic data on eruptions of native species on mainlands has generally been limited to descriptions of changes in population numbers and mass die-offs (Rasmussen, 1941; Leopold  et  al., 1947; Banfield, 1949). Where demographic data are available, the  populations have followed the two-stage equilibrium model.  Both wildebeest  (Connochaetes taurinus aibojubatus) and buffalo (Syncenis caffer) populations appear to have gone directly from an increasing to a stabilizing phase without showing a marked decline and consequently experiencing a negative instantaneous growth rate (Sinclair, 1979).  Increased juvenile survival appeared to be the major factor involved in the 53  population upswing (Talbot and Talbot, 1963), with stable equilibrium points being reached as a consequence of decreased adult and juvenile survival through food limitation (Sinclair, 1979; Sinclair et al., 1985). The Mackenzie bison population is undergoing an eruptive increase following its liberation.  The Mink Lake (ML) subpopulation, representing the leading front of the  population, is in the increasing population phase (stage 1), and the Mackenzie Bison Sanctuary (MBS) subpopulation, having not as yet undergone a marked decline, is in the stabilizing population phase (stage 2).  Therefore, these subpopulations should exhibit  different demographic characteristics (Caughley, 1970a; Klein, 1970), and permit testing of some predictions generated by the two competing hypotheses. In this chapter I determine population estimates, fecundity, survival, juvenile recruitment, diet quality and composition, and an index of physical condition in order to address the various predictions.  METHODS  Herbivore Demography  Population Estimates Censuses have been carried out at approximately two year intervals since 1964. All estimates prior to 1989 were based upon total count aerial censuses. Prior to 1987, aerial reconnaissance was conducted over all open habitats associated with the major lakebeds 54  (see Fig. 1.1) and the travel corridors in and around these lakebeds. Total counts were conducted in March or April, at a time when bison are congregated in the open habitats and when fresh tracks in the snow are easy to distinguish from the air. Various staff of the Northwest Territories Department of Renewable Resources conducted these censuses. The 1987 total count was conducted over the entire study area. Open habitats were censused as above. Parallel transects 3 km apart were flown over all forested habitats in order to cover the entire area. All fresh tracks were followed until animals were sighted. The plane then returned to the transect. The population estimate for 1989 was carried out using a stratified sampling technique. The area was divided into high, medium, and low strata based upon population density estimates from previous aerial reconnaissance flights. In the high density stratum (all open habitat and major travel corridors) a total count was employed, with photographs used to verify counts.  Two observers, one on either side of the Cessna 185, were  responsible for counting groups, while one person recorded the information directly onto a 1:250,000 map of the area.  Strut markers were positioned on the aircraft wings that  permitted observers to monitor a transect width of 500 m on either side of the aircraft (following Norton-Griffiths, 1978). Parallel transects 1 km apart were flown at a constant altitude of 500 m above ground level (agi) so that the whole stratum was surveyed. In medium and low density strata (forest), systematically spaced parallel transects were flown. Sampling intensity was allocated optimally based upon available air time and expected density; more time was devoted to medium than low density strata. The distance between transects was 5 km and 7.5 km for medium and low density strata respectively. Transects ran across the shorter dimension of each stratum to increase sampling units and 55  reduce sampling error. Stratum area was determined using a polar planimeter. Estimated population size (Y) and sample variance of the estimate were calculated for each stratum following Jolly’s (L969) method for unequal sized sampling units (Norton-Griffiths, 1978).  Y  =  Z*(y/z)  =Z*R Var. (Y) where:  =  2 N[(Nn)/n]*(Sy  -  2R*Szy + 2 *Sz R )  y  =  individual transect total of animals  z  =  individual transect areas  Z  =  total animals counted (y)/totaI area searched (z)  R N n  total census area  =  =  total possible number of transects actual number of transects flown  2 Sy  =  variance of transect totals of animals  2 Sz  =  variance of transect areas  Szy  covariance between animals counted and transect area  The population estimates and sample variances for each individual stratum were summed to obtain the total population estimate. I tested for differences in estimates between years using the following formula (Norton-Griffiths, 1978 adapted from Cochran, 1954):  56  1 (Y  -  ) 2 Y  d=  (Var (Y ) 1  where: 2 Y  +  Var (Y 5 ) 2 )°  =  population estimate time 1  =  population estimate time 2  Var (Y ) 1  =  sample variance of estimate Y  Var (Y ) 2  =  sample variance of estimate Y 2  I determined a sightability correction factor (following Steinhorst and Samuel, 1989) for animals in forested habitats by comparing the number of radio collared females in forests actually counted by the observers in forests with the number of radio collared females known from radio locations to have been in that habitat during the counts. This correction factor was calculated during the 1987 census as a rough approximation. One of three radio collared animals in the forest was missed giving a correction factor of 33%. This may be a conservative estimate because visibility of gregarious animals in forested habitats ranges from 50-70% (Samuel and Pollock, 1981; Samuel et a!,, 1987). However, in Wood Buffalo National Park 60-70% of radio collared bison located in forested areas were actually seen and counted (E. Wilson, pers. comm.). The 1992 census was designed to be a total count, but logistic problems combined with a change in animal distribution prevented a proper total count of the study area. Pre census reconnaissance of the area indicated animals were dispersed throughout more of the forested habitats than during other years. This was attributed to higher water levels than  57  the previous summer and fall which had flooded prime winter foraging habitat (C. Gates, pers. comm.). Consequently, all transects flown over forested habitats had to be corrected for sightability bias. For forested areas I used the sightability correction factor determined during the 1987 census, and assumed that all animals in the open habitats were seen and so represented a total count. All animals counted on transects in forested habitats represent a sample of the total number of animals located in forested habitats.  The population  estimate was determined as described for the 1989 census. The instantaneous rate of growth (r) was determined for the entire period and for each period between censuses using the following formula:  (lnN(÷fl)  -  lnN())  r=  n  where:  t  =  n N  =  initial time (year) number of years from the initial time population estimate  I calculated population density by dividing the estimated population number by the area ) censused. 2 (km  Recruitment and Juvenile Survival Aerial sex/age class counts were conducted in MBS once during summer from 1984 to 1988. From 1989 to 1991, aerial sex/age class counts were conducted in early (mid-late 58  June) summer and late (mid-late July) summer in both MBS and ML. At least 80% of all calves are born by 1 July (Komers, 1992). Observers were positioned near groups of bison by helicopter, and the bison were encouraged to parade past the observers so that all animals could be classified into different sex/age categories. Animals were classified into the following categories: calves, yearlings, adult females, immature males, and two classes of adult males (old and young mature). Body size, pelage, horn shape/wear and size were the characters used to separate animals into the 6 categories.  Because one group of  observers consistently confused immature males with yearlings during 1990, I corrected for this by calculating the ratio of immature males to adult females from 1989 surveys and removed this proportion from the adult female to yearling ratio.  This represents a  conservative correction. Aerial sex/age counts were conducted in one day so that the probability of recounting individuals in the different areas was negligible. Ground-based sex/age counts were also conducted between 1986 and 1991. Observers used all-terrain vehicles to position themselves near groups.  Ground-based  counts were conducted daily throughout the summer and winter field trips. The probability of recounting individuals increases with the length of time over which ground counts are conducted.  Additionally, combining data over extended periods of time masks actual  changes in adult female:calf caused by changes in neonatal mortality. Therefore I lumped aerial and ground count data only when ground count data were collected within a 10 day period encompassing the aerial Count.  Attempts to conduct counts in early winter  (November) and late winter (March-April) were hampered by logistic problems and small sample sizes; therefore these samples were lumped over the entire winter period in order to provide some measure of winter ratios. 59  To calculate juvenile recruitment and survival of calves, I used the ratios of calves: 100 adult females and yearlings: 100 adult females. The standard error of the ratios was computed following Cochran (1963) and Sinclair and Grimsdell (1978).  Ratios were  compared between area and between different sampling times using the proportions test (Zar, 1984), and over all sampling times using the sign test (Conover, 1980).  Adult Survival Adult survival was estimated from radio telemetry data. Between 1986 and 1991, 53 adults (17 females and 36 males) were equipped with radio collars. I calculated an  estimate of adult survival (S) following White and Garrott (1990):  S  =  ni/no  Var (S)  =  S(1  -  S)  no  where:  1 n no  number of animals alive at end of year =  number of animals alive at start of the year  Because the number of animals with radio collars varied greatly between years and data were lost by limiting it to year long periods I calculated a second estimate.  I calculated  the number of radio collar-months by summing the number of months an individual animal had a functioning radio collar over all radio collared individuals during the 5 year study period. I determined the number of radio-collared adults that died during the study period and used the two figures to estimate adult survival.  60  Bison Diet  Fresh faecal samples of approximately 30 g (wet weight) were collected during summer (June to August), 1989 and 1990, and in winter (March) 1990. Samples were grouped by month, with a minimum number of 18 samples per month (range 18 to 62). Samples were kept frozen until they were oven dried at 60°C for 48 hours. Diet composition was determined by analyzing faecal plant fragments (FPFA) (Sparkes and Malechek 1968).  Samples were ground in a Brinkman centrifugal mill  through a 1 mm wire mesh screen, and washed with Hertwig’s clearing solution before mounting on microscope slides for identification of cellular fragments. The analysis was conducted by C. Melton following the method outlined in Hansen et a!. (1976). Although the reliability of this method has been questioned, I deemed it suitable for this study because: i) differing proportions of forage classes, not changes in individual species composition were of importance, and ii) previous studies indicated that the diet was dominated by monocots (Reynolds et a!. 1978; Larter, 1988; Larter and Gates 1991a). Differential digestibility of forages, especially in diets dominated by forbs, has been cited as the main cause of inaccurate diet determination from FPFA (Pulliam and Nelson, 1979), especially when compared with analyzing rumen plant fragments (RPFA) (Voth and Black 1973; Smith and Shandruk 1979; Leslie et a!. 1983; Putman 1984; Barker 1986). Differential digestibility was not a problem when the diet was dominated by monocots (Hansen eta!. 1973; Todd and Hansen 1973), nor by browse (Holechek eta!. 1982) as both FPFA and RPFA gave similar results. Johnson et a!. (1983) determined that digestion did reduce the mean weight, but did not eliminate the particle, and they found the same 61  proportion of discernable plant fragments before as after digestion. The majority of dietary components were only identified to genus. groups: grass, sedge, shrub, forb, and lichen.  They were placed into five major forage Shrub material was represented almost  exclusively by willow. Trace quantities of Equiseizim spp., moss, pine, and spruce needles were discarded from the analyses. Because of the lack of independence between forage class proportions, parametric tests were only conducted on one forage class: sedge, the most important bison forage (Larter, 1988; Larter and Gates, 1991a; 1991b). transformed.  Sedge proportions were arcsine  I used a three-way ANOVA on the transformed summer dietary sedge  proportions to determine whether there were year, month, or area (MBS and ML) effects. Because there were no year effects (F()  498).  p>O.2), I pooled the data across years and used  a two-way ANOVA with month and area as the factors. I used a oneway ANOVA to test for differences between area in the winter dietary sedge proportions (Zar, 1984).  Diet Ouality  I calculated an index of diet quality by combining different forage quality measures (see Chapter 2), with diet composition data (see above) (Larter, 1988). The index (1) was determined for each month and area combination by the following formula:  I  =  (CP  *  DG()  *  DC())  62  where:  CP  the mean percent crude protein (see Chapter 2)  FT  the forage type (sedge, grass, willow)  DO  =  the mean percent digested (see Chapter 2)  DC  =  the percent composition in the diet  Because sedge, grass, and willow made up the majority of the summer diet, the index is based upon only their forage quality measures. I used the forage quality measures of Carex atherodes collected from willow savannas to represent the summer dietary sedge  component. The winter (March) index is based solely upon the dietary sedge component. I used forage quality measures of C. atherodes collected from both willow savannas and wet sedge meadows during mid-February and early April in combination with the March dietary sedge component. Because there were no significant year effects for the different forage quality measures (see Chapter 2), I calculated the index for all time and area combinations and then  pooled the data across years.  I compared relative differences  between areas using log (1).  Faecal Nitrogen Condition Index  Comparative faecal samples were collected from ML and the MBS during summer, early winter (November), mid-winter (February), and late-winter (March-April) as a second measure of diet quality. 30 ml (wet weight) scoops from a minimum of five separate faecal pats were lumped together to form one sample. There were between 7 and 14 samples collected from each area during each time period. Samples were stored frozen, dried for 63  48 hours at 60°C, and ground in a centrifugal grinder through a 1.0 mm screen. I used the macroKjeldahl procedure (Parkinson and Allen, 1975) to determine faecal nitrogen content. Briefly, 1 g of dried sample was placed onto a ashless filter paper. The filter paper was folded to enclose the sample in a pouch and then placed into a 300 ml digestion test tube.  Samples were predigested by adding 15 ml of a sulphuric acid-  ) solution, and approximately 2 ml of hydrogen peroxide, 0 2 selenium-hydrogen peroxide (H and gently stirring the solution until a boiling reaction occurred.  Samples were then  digested by placing the test tubes into a block digester, heated to 410°C for 50 minutes, cooled for LU minutes before adding approximately 2 ml of 11202, and returned to the block digester for an additional 10 minutes. The remaining solution was cooled and then made up to 100 ml with distilled water. A subsample of this solution was autoanalysed at the Animal Science Department, University of British Columbia, to determine the nitrogen content (ppm). The data were corrected for dry matter and are expressed as total percent nitrogen.  Animals with high faecal nitrogen content have a better quality of diet and  therefore are in better condition than animals with low faecal nitrogen content. I used a two-way ANOVA to determine if sampling time and/or location explained significant variation in faecal nitrogen content. I used the Student’s t-test to compare means between area for a given sampling time.  Dispersal  Between 1986 and 1991, 93 animals were marked, either with brands, radio collars, or both. The original location of the animal at time of marking was noted. The 64  majority (85) of these animals were initially marked in MBS. Movements were monitored by relocation of marked individuals. Relocation was either by active monitoring of collared animals, or by passive observation of marked animals. Additional aerial reconnaissance of Mills Lake, at the southwest edge of the study area, was conducted in order to determine if animals had moved into the last major area of potential wood bison habitat. Animals that moved from MBS to ML, ML to MBS, or into the Mills Lake area, and remained in that new area for successive relocations were considered to have dispersed.  RESULTS  Herbivore Demography  Population Estimates The total bison population (MBS and ML subpopulations combined) has increased sigmoidally since 1964 (Fig. 3.la). Population estimates of the two separate subpopulations have only been conducted since 1987. Although there appears to be an increasing trend for ML from 1987 to 1992, and not for MBS during the same period (Fig, 3.lb), the 1989 subpopulation estimates for both MBS and ML are not significantly  (p>O.O5) different from  the 1992 estimates. When total population density is plotted against the instantaneous rate of growth a cyclical density dependent pattern appears (Fig. 3.2). Both population growth rate and density increase until population density exceeds approximately 0.55 animals/km . Once 2 this point is reached both rate of growth and density decrease and the cycle begins again. 65  3000  a  II  2500  C 2000  0  -  1500  Q. 0  a.  500  0 64  66  68  70  72  74  76  78  80  82  84  86  88  90  92  Year (19__)  2000  b C  0  MBS  1600  -I  0.  1200  0  Q. 800  C,) 400  ‘ML’ 0 1987  1989  1992  Year  Figure 3.1. (a) Growth of the Mackenzie wood bison population since its introduction. (b) Growth of the MBS and ML subpopulations since [987. Bars represent 95% confidence interval. 66  0.5  0.4 I  0.3  cc  0.2  -  .4-’  0  0.1  I  0  0  -0.1  -  0.15  0.25  0.35  0.45  0.55  0.65  0.75  Density  Figure 3.2. The relationship between instantaneous growth rate (r) and total population density over time. Each point represents the population density of a given year, starting with 1964, and the corresponding r of the population from that given year to the next year the population was estimated.  67  The decrease in density occurs even though the population is increasing because animals have dispersed into previously unoccupied areas. Hence, animals are distributed over an enlarged area and population density declines. The cycle has been completed twice since the population was introduced with the two dispersal episodes occurring in 1969 and 1980, the years following peak population density.  There is preliminary evidence that adult  female and juvenile bison occupied the Mills Lake area for the first time during winter 1992-93 (T. Chowns, pers. comm.).  Recruitment and Juvenile Survival Both ratios of calves:100 adult females and yearlings:lO0 adult females showed considerable yearly variation (Table 3.1). Both ratios were higher in the ML than the MBS subpopulation, except in June 1990 (Table 3.1). The ratios were not significantly higher in ML at any one time (Z scores <1.15, p values >0.12, ratio test), because of the high SE associated with the ML data. However, between June 1989 and July 1991 both adult female:calf and adult female:yearling ratios were significantly higher in ML (p<0.036, n=8 for both ratios, sign test). The mean (±SE) probability of survival from calf to yearling was higher in the ML subpopulation (0.865±0.195) than in the MBS subpopulation (0.711±0.070).  Adult Survival Eight adults (4 females and 4 males) died during 1362 radio collar-months of observation, 7 of natural causes including wolf predation and one male was killed illegally. The estimates of survival (S) (following White and Garrott, 1990) ranged from: 75 (Var. 68  Table 3.1 The ratios of the number of calves (Ca) and yearlings (Yr) per hundred adult females (Cow) from 1984-1991. Ratios (±SE) are presented for each subpopulation and for multiple times per year where available.  MBS  Ca/iQO Cow  ML  Yr/].0O Cow  Ca/100 Cow  Yr/100 Cow  1984  July  48.0  1985  July  53.0  1986  July  45.0  1987  July  41.0  1988  July  32.0  1989  June  46.4 ± 12.5  26.1 ± 11.0  48.0 ± 25.0  32.0 ± 23.3  July  43.1 ± 9.5  22.5 ± 8.0  56.3 ± 18.7  35.3 ± 18.1  August  37.6 ± 10.6  23.4 ± 9.2  61.9 ± 21.7  28.6 ± 20.2  Winter  35.9 ± 11.6  15.7 ± 8.8  45.8 ± 28.8  18.8 ± 22.5  June  32.3 ± 5.8  37.3 ± 6.2  11.4 ± 10.].  36.8 ± 15.3  July  33.7 ± 6.5  30.6 ± 6.9  34.2 ± 11.5  34.9 ± 11.6  August  34.9 ± 6.3  26.3 ± 6.6  June  47.6 ± 6.7  20.8 ± 5.4  55.6 ± 24.8  25.4 ± 21.8  July  39.2 ± 10.2  24.1 ± 8.9  46.7 ± 35.3  33.3 ± 33.3  August  36.8± 11.7  25.7±10.6  1990  1991  —  —  —  —  —  —  —  —  —  —  22.0  —  —  —  —  —  —  —  —  —  —  38.0  —  —  —  —  —  —  —  —  —  —  42.0  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  —  14.0 23.0  —  —  ——  —  —  —  —  —  —  ——  —  —  ——  69  4.7%) to 100% (Var. 0%) for adult females, 67 (Var. 3.7%) to 100% (Var. 0%) for adult males, and 75 (Var. 4.7%) to 96% (Var. 0.1%) for all adults combined. Using radio collar information lumped over the entire study period, the probability of an adult bison dying during one year was 7.1%. Adult females had similar survival rates to males: 91,5% a year versus 94.0% for males; this difference was not significant.  Bison Diet  There were significant area (F( 1  499).  p<O.OOI) and month (F( ), p<O.OO1) effects 2499  on the proportion of sedge in the summer (June to August) diet. Bison inhabiting the MBS had a higher dietary sedge component (Fig. 3.3), and the proportion of dietary sedge was highest in June. During late winter (March) bison inhabiting ML had a significantly (F( 1 48)’  p<O.OO1) higher proportion of sedge in the diet (Fig. 3.3). The proportion of willow in  the diet was greater for animals from ML.  Diet Ouality  Bison inhabiting MBS had a consistently higher index of diet quality than bison at ML during summer, regardless of year (Fig. 3.4). This was reversed in winter; bison inhabiting ML had a higher index of diet quality. Although the difference in winter diet quality index is not great in absolute terms, in relative terms the index measure is 1.47 times higher for animals inhabiting ML. The relative differences during summer were much less ranging from 1.06 to 1.17 (August and July respectively). Therefore during late 70  MBS 100%  75%  50%  P. .4  25%  C 0 4-’  N N N NI  -  0%  0 0  E  0  ML 100%  C-) •1-’  a)  Jun  Sedge  Jul  Grass  Aug  Shrub  Mar  Forb  Lichen  Figure 3.3. Diet composition of animals inhabiting the MBS and ML during summer (June to August) and winter (March). Summer data are lumped over two years (1989 and 1990).  71  RMBS VML  x C >%  C  II June  Aug  July  Mar  Month  Figure 3.4. Index of diet quality between area. See text for calculation of index. 72  winter bison in ML have a relatively far superior diet to bison in MBS.  Faecal Nitrogen Condition Index  There were significant location (F  83)’  p<O.OO1) and sampling time (F( 4  p<O.0001) effects on nitrogen content of faecal material.  83)’  Faecal nitrogen content was  higher in summer than winter, and declined throughout the winter (Fig. 3,5). Regardless of sampling time, faecal nitrogen content was significantly higher in animals from ML, indicating a better quality of diet for ML animals (range  p<O.OO3, t3.59, n=8  in February  to p<O.0003, t5.7, n8 in November).  Dispersal  The majority of animals remained in the area where they were originally collared. There were more radio collared animals at MBS and also more males (Table 3.2). No animals moved from either MBS or ML to the Mills Lake area (Fig. 1.1). Small numbers of adult males were the only residents of Mills Lake. They spent fall and winter at Mills Lake, but moved into both ML and MBS during summer and throughout the rut (mid-July to late August). Only two of the five males collared in Mills Lake did not lose their collars in fights during the first rut after they were radio collared. Both of these animals returned to Mills Lake for the subsequent winter. The remaining male with a functioning radio collar in 1992 returned for a fourth consecutive winter to Mills Lake. Two adult females moved from MBS to ML and remained there for as long as their 73  3.5 ML 3  I  MBS  2.5 2 1.5  -  10.5  -  0 July  August  November February  April  Figure 3.5. Mean (±SD) faecal nitrogen content (%) of bison from ML and MBS at different times of the year. In all cases nitrogent content is significantly (p<O.OO3) higher for animals from ML. 74  Table 3.2. The number of radio collared male and female bison and their original location when radio collared.  MBS  ML  Mills  Males  31  0  5  Females  14  3  0  Total  45  3  5  75  radio collars functioned: one female moved 8 months after being radio collared in 1986 and remained at ML for at least 46 months. No adult females collared in ML moved to MBS and remained there.  One adult female spent three weeks of one summer in MBS, but  returned to ML. One adult female and one adult male from MBS did the opposite, briefly moving to ML from MBS. Again this occurred during a few weeks in mid-summer.  DISCUSSION  Demographics  The MBS subpopulation appears to be stabilizing, whether or not it has experienced a period of negative instantaneous growth (r) is debatable. Future subpopulation estimates are required in order to determine if the calculated negative r is real or an artefact of sampling. If future estimates confirm negative r, this finding will be most consistent with the original four-stage model of the equilibrium hypothesis. Juvenile survival and recruitment were higher in the ML than the MBS subpopulation as predicted by all hypotheses. Although differences in juvenile survival and recruitment were not statistically significant when compared at any one time, they were when compared over the length of the study. The lack of statistical difference, caused by high SE’s, between ratios compared over the short term may hide real biological differences.  Small populations of gregarious animals, like the ML subpopulation, are  inherently susceptible to high SE’s because the number of groups observed comprises the ratio estimate and has a major effect on the SE.  The biological difference of these ratios 76  is reflected in subpopulation growth rates. High recruitment provides the potential for continued population increase at the dispersing front of an erupting population (Riney, 1964; Caughley, 1970a; Leader-Williams, 1980) and that has been evident in the ML subpopulation. A more direct comparison of calf production between areas would overcome the statistical shortcomings of the adult female:calf ratio. Pregnancy status would provide such a measure, however determining this requires capture and immobilization. Capture and immobilization of animals has been expensive, stressful to the animal, and hampered by small sample sizes (Greer and Hawkins, 1967; Valkenburg eta!., 1983; Wood eta!., 1986). Recent research has focussed on determining the pregnancy status of various ungulates by measuring the level of hormones in faecal material (Desaulniers ci a!., 1989; Kirkpatrick et at., 1990; Messier et al., 1990; Larter et a!., 1993a; 1994). These results have been promising, as faecal hormone concentration levels associated with non-pregnant and pregnant animals have been documented over a wide range of extraction procedures (Larter et at., 1993b). Soon it may be possible to determine accurately the pregnancy status of a large segment of the population, based upon faecal samples, and hence obtain a better measure of potential calf production. Decreased adult survival may not only depress overall population size, but may also affect the adult female:calf and adult female:yearling ratio estimates of juvenile survival and recruitment (Caughley, 1974).  Ratios can change by variation in any one or  combination of the following: juvenile survival, adult fecundity, adult survival. Therefore, higher calves:100 adult females and yearlings:100 adult females ratios in ML could be a function of consistently lower adult female survival.  I do not have comparative adult 77  female survival rates between area. However, pooled over the past six years the average adult female survival rate was 91.5%. The 75% estimate, following White and Garrott (1990), resulted from a sample size of four.  The average of the four other one year  estimates based on sample sizes of 6, 10, 10, and 12, was 93%. In large mammals, once animals have been recruited into the population adult survival is generally consistently high (Caughley, 1977). It is unlikely that adult female survival would be consistently lower in ML than MBS. Furthermore, movement of adult females from one area to another would tend to increase adult female:calf and adult female:yearling in the area from which animals were leaving. This appears to be happening to some extent in this system with animals moving from IVIBS to ML. Therefore the difference in ratios and consequently juvenile and recruitment between areas is conservative.  Higher calves:100 adult females and  yearlings:100 adult females in ML are most probably a result of higher adult fecundity and juvenile survival, not of lower adult female survival. One additional prediction that I was unable to address adequately was that younger females in the ML subpopulation should have higher calving success. I could not classify two-year old females in the aerial sex/age class counts and determine whether or not they were accompanied by calves.  Diet Quality and Composition  The equilibrium hypothesis and optimal foraging theory (Pyke eta!, 1977; Stephens and Krebs, 1986) predict that animals in ML should have a superior quality diet than animals in the MRS. whereas the facilitation and feedback hypothesis predicts that animals 78  in MBS should have the superior quality diet. Animals in ML had superior quality diets in winter, while animals in MBS had superior quality diets in summer. Given access to greater amounts of available high quality sedge forage, why did ML animals consume such a large amount of lower quality willow forage during late summer? Summer leaves and stems of willow contain tannins and lignin.  Tannins can reduce digestibility (Zucker,  1983), and lignin is essentially indigestible, interfering with cellulose digestion (Van Soest, 1963). Lignin levels in willows found in ML were lower than those from MBS during early summer, but increased to comparable levels by late summer (August) when the dietary willow component of ML animals was greatest. One reason for ML animals’ feeding behaviour may be related to the distribution and availability of willows. Willow savanna (WS) habitat in ML had a greater abundance of sedge than WS in MBS, but data on available biomass of willows are lacking. Available willow biomass may be greater in ML than MBS. In MBS, WS habitats are associated with shallow lakebeds (Fig I 1) which are distributed patchily amongst forested habitats and represent large discrete foraging patches. These patches are connected by traditional travel corridors of forest. In ML, WS habitats are scattered as smaller, less discrete patches, and travel corridors include more riparian areas with high willow densities. Hence, although the proportion of WS in both areas is similar, its distribution is different, which may affect how willow forage is encountered. Animals that forage while moving between WS patches are more likely to encounter willow as a potential forage in ML than MBS. Also, the ratio of available sedge to willow may be lower in the smaller WS habitat patches of ML. Consequently animals may be more likely to use willow forage before abandoning the smaller habitat patches. Either 79  scenario would lead to a larger willow component in the diet of ML animals during summer. Another possible explanation for a higher sedge component in the summer diet of MBS animals is related to annual energetic costs. Summer is the only time when there is  an abundance of forage high in crude protein. During winter most wild ungulates face undernutrition due to the seasonality and unreliability of food supplies (White et al., 1981; Tyler, 1987a). Stored fat reserves do not fully substitute for winter forage, but do play an important energetic role in winter (Tyler, 1987b; Tyler and Blix, 1990).  The bison  digestive system has adapted to make it possible for them to maintain themselves on coarse low quality forage (Janis, 1976; Hofmann, 1989). Bison can extract more nutrition from poor quality forage than can domestic cattle or yaks (Richmond, et cii., 1977; Hawley eta!., 198 Ia; Deliberto, 1993). However, it may still be necessary for certain threshold levels of body fat to be reached during summer in order for the animal to maintain itself through the winter. I hypothesize that MBS animals have higher annual energetic costs, due to predation, than do ML animals and therefore must reach higher threshold levels of body condition during summer. In order to reach these levels it is essential that animals in MBS forage in an energy maximizing fashion during summer. This would result in a diet with a high sedge component. I base my hypothesis on differences in predator pressure between subpopulations. Firstly, wolf activity in winter is greater in MBS than ML (see Chapter 4). Secondly, MBS animals have been subjected to human hunting pressure (in winter) since 1988, while ML animals have not been hunted. When bison are harassed by either predator their response is to flee (pers. obs.).  Bison are capable of running at speeds 80  greater than 50 km/h. The costs expended while fleeing, especially in deep snow and when temperatures are often well below 0°C, cannot be negligible. Flight response is not a single event, but happens regularly during winter. The hunting zone includes Falaise Lake and three of the other six major lakebeds in the MBS (Fig 1.1). Hunting efforts are concentrated on lakebeds. Hunting has displaced bison to lakebeds out of the hunting zone in ever increasing numbers (N. Larter, unpubi. data), which effectively decreases the relative availability of sedge forage for animals in the MBS. Decreases in sedge availability may have been further exacerbated by recent increases in water levels. This reduction in available sedge has been reflected in both the winter diet composition and diet quality data. There has been a drastic change from a winter diet of almost 100% sedge (for animals in the MBS) in years immediately prior to active hunting (Larter, 1988; Larter and Gates, 1991a; 1991b), to the current level of 40% sedge. Bison in the hunting zone have modified their behaviour in response to increased predator harassment. They showed an immediate flight response when researchers came within 1 km of them, whether on snowmachine or on foot (pers. obs). Animals found out of the hunting zone or in ML did not flee when approached, were exceedingly tolerant of researchers even on snowmachines, and would flee only when approached to <25 m. Hunting appears to have had a major impact on bison behaviour and distribution. Actual forage availability has been reduced by the combination of increased hunting pressure and high water levels. I therefore hypothesize that MBS animals incur much higher energy costs than do ML animals, and consequently must forage in an energy maximizing fashion during summer in order to survive in winter, whereas ML animals lack the necessity to 81  forage in an energy maximizing fashion. These different foraging strategies are reflected in the higher sedge component of the summer diet of MBS than ML animals.  Faecal Nitrogen Condition Index  The equilibrium hypothesis predicted that animals inhabiting ML would have higher faecal nitrogen (FN) levels than animals inhabiting MBS. The feedback and facilitation hypothesis predicted that MBS animals would have higher FN levels. Animals inhabiting ML had significantly higher FN levels. This finding is in agreement with the predictions of both models of the equilibrium hypotheses, but is contrary to the predictions of the facilitation and feedback hypothesis. Strong correlations between faecal nitrogen (FN) content and dietary nitrogen (DN) have been demonstrated in a variety of wild herbivores: deer, Odocoileus hemionus (Leslie and Starkey, 195; Mubanga et al., 1985), elk, Cervus elaphus (Mould and Robbins, 1981), snowshoe hare, Lepus americanus (Sinclair et al., 1982), pocket gopher, Thornornys bottae (Loeb and Schwab, 1989), and several wild African ruminants (Erasmus et a?., 1978). However, the use of FN as a predictor of diet quality has been criticized (Hobbs, 1987; Robbins et a?., 1987).  Summer leaves and stems of deciduous forage contain tannins.  Tannins can reduce digestibility (Zucker, 1983) by precipitating proteins into essentially indigestible complexes which are excreted in the faeces.  Consequently, increased FN  values could lead to an overestimate of DN, and therefore diet quality, because the diet contains forage with high concentrations of protein-binding phenolics (Mould and Robbins, 1981). However, Caughley and Sinclair (1994) noted that in studies where animals were 82  allowed to forage on normal feeds, not forced to feed on unnatural feeds, the relationship between faecal and dietary nitrogen is well documented for a variety of animals. Bison feed predominantly on sedges and grasses and generally do not ingest large quantities of browse (Larter, 1988; Larter and Gates, 1991a). Therefore, I had no a priori reason to dismiss FN as a potential index of bison diet quality and condition. My diet quality data indicate that in winter ML animals had a better quality of diet than did MBS animals. These data are based upon quantitative measures of nitrogen and fibre content of the forages found in the winter diet (Chapter 2).  The FN results  corroborate this: ML animals had significantly higher FN values than did MBS animals in winter. In contrast, while the summer data showed FN levels were higher for ML than MBS animals, diet quality measures indicated MBS animals had a higher quality of summer diet than did ML animals. There are two possible explanations for these inconsistent results concerning summer and winter diets. First, the negative effect of the willow proportion of the diet may be underestimated by my diet quality index measure. Acid-pepsin digestibility and fibre content are not as highly correlated in willow as they are in grass and sedge (Larter, 1992). ML animals have a higher willow component in their summer diet. Second, species differ in their ability to tolerate dietary tannin (Robbins et a!., 1991), and those most tolerable of dietary tannin are capable of neutralizing or metabolizing tannins in order to minimize FN loss. Grazers (cattle and sheep) are much less tolerant of dietary tannin than are browsers (mule deer) (Robbins el a!., 1991), therefore, elevated FN levels in ML animals may be related to the higher willow component of their diet.  However, the  relationship between faecal nitrogen and dietary nitrogen has not been determined for 83  bison naturally foraging on willow. Whether or not ML animals have an inferior summer diet is unclear. However, the quality of winter diet must offset any summer inadequacies because an inferior annual diet is not reflected by demography. Adult fecundity is higher in ML than MBS and there are no outward signs of malnutrition (pers. obs.).  Dispersal  Bison did not appear to disperse from areas of superior habitat to areas of inferior habitat: the classic source-sink dispersal theory where the disperser’s only chance of survival is to leave (Lidicker, 1985). Evaluation of habitats in both areas indicates that ML habitats currently provide more available high quality forage than do MBS habitats. This is related in part to the different grazing intensities each area has been subjected to (see Chapter 2). Dispersal did not appear to be a slow diffusion process (Caughley, 1970b), but rather it occurred in pulses possibly related to some density dependent effect, such as food supply, on the population (Gates and Larter, 1990). Males appeared to initiate the dispersal in a fashion similar to that documented in most mammals (Greenwood, 1980). Possibly the two subpopulations are not distinct units and the radio relocation data are equivocal. However, data on the calving period from class counts of 1990 demonstrate that they are two separate subpopulations. Females inhabiting ML calved later than those inhabiting MBS.  Random mixing of females between the two subpopulations would  preclude such a distinct difference in the timing of calving between areas. In conclusion, the two subpopulations appear to be distinct and are at different 84  stages of eruptive oscillation. The demographic characteristics of these subpopulations are more consistent with the predictions of the equilibrium hypothesis, with one exception: summer diet quality.  Animals inhabiting ML were predicted to have diets of superior  quality to that of animals inhabiting MBS. I attribute this exceptional result to the presence of predation in this system. While Caughley (1976b) argued that the presence of a predator does not affect the predictions of the original four-stage model of the equilibrium hypothesis, my findings indicate predation may affect the predictions. Since predation is by both wolves and man, these predator effects are confounded at present. Future studies where this hypothesis can be tested in systems with predators are required to better assess the effects of predation on the robustness of the four-stage model.  85  SUMMARY  1)  The MBS subpopulation shows signs of stabilizing or possibly declining. Whether it is stabilizing at a significantly lower level than its peak level remains to be determined by future censuses.  2)  The ML subpopulation shows an increasing trend but this increase is not statistically significant.  3)  Animals inhabiting ML had poorer summer diets than animals inhabiting IvIES. This trend was reversed in winter when animals in ML had superior diets. I suggest that these results stem from differences in winter energy costs associated with different predation pressures between areas.  4)  The faecal nitrogen condition index indicated that animals inhabiting ML had a better quality of diet and hence were in better condition than animals inhabiting MRS during both summer and winter. I suggest that the different results between  the two measures of diet quality is a function of the unknown relationship between dietary nitrogen and faecal nitrogen for bison when willows make up a significant proportion of the diet. 5)  Juvenile survival and recruitment were higher for animals inhabiting ML versus MRS.  6)  Dispersal has occurred in pulses once a certain density threshold has been passed. Movement between the two subpopulations appears minimal.  86  CHAPTER 4. THE RESPONSE OF PREDATORS TO AN ERUPTING POPULATION OF BISON  INTRODUCTION  My data (see Chapter 3) suggest that the current eruption of wood bison is following the dynamics predicted by the four-stage equilibrium model (Riney, 1964; Caughley, 1970a) in the presence of wolf predation. Beyond documenting that the model’s predictions appear to hold for an indigenous ungulate population in the presence of predators, further important questions arise regarding the response of the wolf population to an increasing prey base.  Do wolf numbers increase in response to the increasing  abundance of wood bison? If wolf numbers increase, what impact does an increased wolf population have, not only on the wood bison population, but also on moose (A ices aices), the only other ungulate prey population in this study area? In systems where there is an alternate prey source which the predator can utilize, it is theoretically possible for predation rate to be either directly density dependent or inversely density dependent over all densities (Ricklefs, 1979). Inverse density dependent predation rates occur when per capita risk of predation increases while prey density decreases, a situation that could be destabilizing and potentially drive a prey population to extinction. I hypothesize that predators in this system prey upon both moose and bison, and that wolf numbers have increased in response to increasing wood bison numbers. This hypothesis predicts that over a range of moose densities wolves will continue to utilize moose to a greater extent than expected given moose and bison availabilities, thus 87  providing the potential for local extinctions of moose populations. Moose are an important northern food resource for native peoples. Areas that have been proposed for future wood bison introductions have low densities of moose and wolves, and a subsistence moose harvest. There is the potential for local extinctions of moose as a result of the bison-wolf-moose interaction. Predation limits population growth by increasing mortality, but the magnitude of its impact, and whether or not it is a regulating factor remains unresolved (Bergerud and Snider, 1988; Sinclair, 1989; Messier, 1993). According to one school of thought, there is increasing evidence that wolves (Canis lupus) can regulate their prey populations. Various authors (Bergerud el a!., 1983; Messier and Crete, 1985; Van Ballenberghe, 1987; Bergerud and Snider, 1988; Ballard et a!., 1990; Gasaway et a!., 1990; Messier, 1993) believe that where moose (A ices aices) are the primary prey and live at low densities, wolf predation is the primary factor limiting moose densities. However, others argue that the supporting evidence is weak (Sinclair, 1989; Skogland, 1991; Boutin, 1992). Nevertheless, there is increasing evidence that predators can regulate small mammal populations (Pearson, 1964; Erlinge et a!., 1983; Kid and Lewis, 1987; Korpimaki and Norrdahl, 1989; Sinclair et at., 1990; Pech, et al., 1992).  Jarman and Johnson (1977) suggested that the decline in introduced hare (Lepus europaeus) and native rat kangaroo (A epyprymnus spp.) populations in Australia may have been a result of introduced fox (Vuipes vuipes) populations being sustained by rabbits (Oryctolagus cuniculus).  Consequently foxes may have been able to eliminate a less  numerous and more catchable prey.  Reynolds (1990; pers. comm.) describes a similar  situation in England where fox numbers are being maintained by rabbits at numbers high 88  enough to significantly reduce breeding densities of gray partridge (Perdix perdix). Rockwallaby (Petrogale lateralis) populations in Western Australia are being decimated by fox predation (Kinnear et at., 1988), because fox numbers are being kept high by a substantial rabbit population. Elk (Cervus elaphus) densities in the Bow Valley (Banff National Park) increased from an all time recorded low of <1 elk/km 2 in the mid 1970’s to 3 elk/km 2 in the mid 1980’s (Woods, 1991). Wolves became established in the Bow Valley during the mid 1980’s. Moose numbers declined concurrently with the appearance of the wolf population, and now there are few moose in the Bow Valley (J. Woods, pers. comm.). However, disease may have played a confounding role in this example. Liver fluke was prevalent in the moose population and was a definite source of mortality.  Wolf predation was  believed to be additive to disease mortality, however the potential synergistic relationship between parasite and predator cannot be ruled out. The vulnerability of the primary prey relative to an alternate prey determines the impact of predation on the abundance of primary prey (Messier, 1993). In areas of western North America, substantial moose populations are the main prey for wolf populations which can remain at levels high enough to deplete caribou (Rangifer tarandus) populations. This situation has been used to explain declines in the Nelchina caribou population in Alaska (Bergerud and Ballard, 1988), in woodland caribou populations in the Wilimore Wilderness Park, Alberta (Edmonds, 1988), and in various woodland caribou populations in northern British Columbia (Bergerud and Elliot, 1986). Seip (1992) documented that a woodland caribou population lived apart from wolves and moose during summer, and was therefore less vulnerable to wolf predation. It also had a lower adult mortality rate and higher calf 89  survival than a similar woodland caribou population that lived with wolves and moose. He believed that because the wolf population was sustained primarily by moose, the latter caribou population could potentially be extirpated. Gates and Larter (1990) proposed that an increasing wood bison population in the Northwest Territories may not be diverting wolf predation from moose and caribou by providing a substantial alternate prey source; rather it could be increasing it by maintaining high wolf populations capable of depleting the more vulnerable prey populations. In this chapter I compare wolf diet, wolf abundance, relative prey abundance and available prey biomass between areas.  I compare wolf-bison-moose dynamics to test  whether wolf predation on moose, in the presence of bison, can act in a destabilizing and inversely density dependent fashion.  METHODS  Predator Diet  Wolf Scat Analysis Wolf seats were collected opportunistically throughout the course of the study. Seats found near wolf kills were not collected, because they could bias the data set. Seats collected in winter were kept frozen, while seats collected in summer were air dried prior to preparation for laboratory analysis. Scat preparation and analysis followed Kennedy and Carbyn (1981).  Seats were autoelaved at 120°C to 130°C for 20 minutes, washed to  remove dirt and other debris, and then dried at 70°C for 24 hours. Following the key found 90  in Kennedy and Carbyn (1981), bone fragments, hair, and feather characteristics were used to determine the prey item(s) present in seats. I compared the presence of prey items in wolf scats between study areas during winter and between seasons, for the Mackenzie Bison Sanctuary (MBS) area only, using the proportion test (Zar, 1984).  Small sample size precluded comparing seasonal  differences for the Mink Lake (ML) area. I calculated available prey biomass (bison and moose) for each area (see Prey Abundance this chapter). I used the log-likelihood ratio (0test, Zar, 1984) to compare the frequency of seats containing the different prey items with that available (live biomass) in each area. I estimated relative amounts of bison and moose consumed using the relationship described by Floyd et a!. (1978).  y  where:  0.38  +  0.02x  kg of prey/collectible scat  y x  =  =  the average weight of individuals of each prey type.  I estimated x for each prey in each area by determining the proportion of juveniles and adults killed based on wolf kill data, and multiplying that proportion by the average weight of juveniles and adults. I used the following average winter weight estimates: juvenile bison 280 kg, adult bison 720 kg, juvenile moose 240 kg, and adult moose 400 kg. Juvenile weights are an average of calf and yearling weights and adult weights are an average of male and female weights (see Prey Abundance for detailed description of 91  weights). I calculated y for each prey, and multiplied y by the number of seats in which the prey items occurred. I multiplied by frequency of seats because the majority of seats (77%) contained only one major prey item (either bison or moose). This was calculated for each area and the relative amounts of each prey species is expressed as a ratio of kg of bison consumed per kg of moose.  Wolf Kills Wolf kills were located opportunistically by ground based and aerial reconnaissance. Whenever possible, kills located from the air were verified by ground observations. Evidence that an animal was killed by wolves included obvious chase sequences in the snow and/or signs of struggle before death (i.e. extensive blood covered vegetation or snow, and broken vegetation). The sex/age class and species of prey were determined based upon visual analysis of the remains. Limb bones (femur, humerus, tibia) were collected when available.  Percent fat content of the bone marrow was determined following Neiland  (1970) and Hout (1988).  Predator Abundance/Activity  I calculated indices of wolf abundance in both study areas using two techniques: track counts and visual observations. Track count lines were established during winter 1989-90 and consisted of three, 1 km transects along seismographic survey lines in each area. The start and end points of each transect were flagged for identification and track counts were made on them throughout the winter when conditions permitted. The number 92  of track lines was increased to five and the transect length increased to 3 km during winter 1991-92 because of the number of zero counts during 1989-90. No data were collected during winter 1990-91. Data were collected during three winter periods: early (November), mid- (February), and late- (March-April). A track-day was defined as a 24 hour period when all new tracks along the trail were made within the preceding 24 hours. Because there were different numbers of successful track-days between area, all data were expressed as number of tracks counted per track-day.  I used a square root transformation to  normalize the data and the Student’s t-test to determine if there were differences in the number of tracks/track-day between areas. Wolves were observed during aerial reconnaissance, and ground-based travel throughout the study area. Abundance indices for both study areas were calculated as the number of wolves seen per day of field work, or the number of wolves seen per flight-hour.  Prey Abundance  Wood Bison Population estimates are given in Chapter 3. I used the 1989 census as the best estimate of bison numbers, and took the sex/age composition of each subpopulation determined from the 1989 census and composition surveys (Gates et a!., 1991) to estimate numbers of the different sex and age classes. I multiplied the numbers of each sex age class by a winter weight estimate to determine available prey biomass. Winter weight estimates were: calves 200 kg, yearlings 360 kg, adult females 590 kg, and adult (old and young mature) males 850 kg (Larter and Gates, 1990; Gates, unpubi. data). 93  Moose Crude aerial transect surveys were conducted in MBS in February L965 and December 1971, by members of the Department of Renewable Resources, Government of the Northwest Territories. The number of moose observed per kilometre of transect flown was determined. In April 1987, I calculated the number of moose observed per kilometre of transect flown during the bison census of the entire study area. A more sophisticated survey was conducted in November 1991 (Shank, 1992), using the Gasaway et a!. (1986) method to estimate moose densities in both ML and MBS. Additional indices of abundance were collected from aerial and ground based reconnaissance of the study area by personnel during 1986 to 1992. The ground based index was calculated as the number of moose seen per day of field work. For convenience, this was then converted into moose seen per week of field work. The aerial based index was calculated as the number of moose seen per flight-hour. Available biomass of moose was estimated as for bison. I used moose age/sex ratios and population estimates collected during the 1991 census (Shank, 1992). Weight estimates were: calves 160 kg yearlings 240 kg, adult females 350 kg, and adult males 453 kg (Blood et at., 1967; Ritcey, 1974; Banfield, 1977; Saether, 1985).  94  RESULTS  Predator Diet  Wolf Scat Analysis Although bison and moose made up a large proportion of the diet, small mammals including mice (Cricetidae), snowshoe hare  (Lepus  americanus),  and birds (both  Gallinaceous and waterfowl) were also consumed. The number of scats containing bison and moose was equal in ML during winter. In MBS, bison was present in more scats than moose (Fig. 4.1). The proportion of scats from MBS containing moose and bison was similar  (p>O.2) between seasons  (Z=O.93 for moose, Z=O.09 for bison). During winter the  proportion of scats containing bison was greater in MBS than that at ML (p’<O.O5, Z2.91); there was no difference in the proportion of scats containing moose (p>O.3, Z=-O.58). Insufficient sample size prevented the seasonal comparison of seats found in ML. The number of seats containing moose was significantly greater than that expected given the available biomass of moose and bison in both ML and MBS (G=340.O ML, G=1624.3 MBS, p<O.OO1) (Fig. 4.2). Wolves consumed 1.62 kg of bison per 1 kg of moose in MBS, compared to 1.17 kg of bison per 1 kg of moose in ML.  95  14  MBS Summer  80  n17 40  40  3 20  a)  Iloii  hlOcs•  Hri  I  hiousi  C)  a) C) 0  0  45  MBS Winter  80  “.54  > 0 40  U. a) LL.  20  3  2 0  Hagi  2  Mouse  Bite  a) C) a) 0  ML Winter so 18  p.33  18  60  40  20  2  2  0 Ham  Prey  -  --  Mouse  -  BItd  Type  Figure 4.1. Percent frequency occurrence of various prey items found in wolf scats during different seasons and in different areas. Values above histograms are the number of occurrences. n = total number of scats. 96  Bison  A  900 0) a) C C  600  Moose  -  300  0  MBS  ML  n.54  Bison  a) 0 C 40 a)  B  Moose  n.33 0 0 30  0  0  20  C  a) 10  a) I LI 0  Figure 4.2. (A) Available bison and moose biomass in MBS and ML, (B) frequency of occurrence of bison and moose in wolf seats (during winter) in MBS and ML. n = total number of seats. 97  Wolf Kills Forty-one of the 46 documented wolf kills were bison. Of the 41 bison kills, 31 were juveniles: 26 calves and 5 yearlings (Fig. 4.3).  The greater proportion of bison,  especially juveniles, in wolf kills was more pronounced in MBS. Moose made up a larger proportion of kills at ML than at MBS (Fig. 4.3). Data on percent fat content of limb bone marrow was limited to 7 animals. All 7 animals were calf bison; six of these animals were killed in late winter (March-April). Percent fat content was highly variable (range 8 to 86%). The calf killed in January had an 83% fat content.  Predator Abundance/Activity  In general, wolf activity was greater in MBS than ML during all three periods of the winter (Table 4.1). These differences between area were not significant when compared at any one sampling period (p=O.O63, t=1.58, early winter;  p=O.65,  t=-O.40, mid-winter;  p=O.3’7, t=O.33, late winter). However, when the data were lumped across the entire winter the difference became significant: 0.43 and 0.22 mean tracks/track-day in MBS and ML l, t2.07). 2 respectively (pO.O Ground based and aerial observations showed a similar pattern to the track count data, but with greater disparity between area. The ground based index of abundance was four times higher in MBS than ML while the aerial based index was twice as high for MBS than ML (Table 4.2).  98  n = 36  n=1O  Other Adu1t Moose iAdu1t Bison •Juv. Bison  0 0 0 I 0 > 0  MES  ML  Figure 4.3. Proportion of prey type found at wolf kills in MBS and ML. n kills.  =  number of  99  Table 4.1.  Comparative wolf activity between the MBS and ML during three winter  periods: early (November), mid (February), and late (March-April). Values are the mean number of tracks/track-day and are lumped over years. The number in parentheses is the number of track days used to calculate tracks/track-day. areas at  p<O.O5,  *  **  indicates significance between  indicates significance at p<O.l.  Winter Area  Early  Mid  Late  Lumped  MBS  0.65 (17)*  0.09 (1 1)  0.43 (37)  0.43 (65)**  ML  0.23 (26)*  0.14 (30)  0.27 (22)  0.21 (78)**  100  Table 4.2.  Indices of wolf abundance between MBS and ML: ground (wolves  observed/field day), aerial (wolves observed/flight hour).  The values in parentheses  represent the number of field days or the number of flight hours the index is based upon. See text for more detailed explanation of indices.  Area  Ground  MBS  0.312 (571)  2.18 (16.5)  ML  0.068 (133)  1.18 (8.5)  Aerial  101  Prey Abundance  Wood Bison There are about 3 times as many bison in MBS as at ML (see Chapter 2 for population estimates). The estimated available biomass of bison was H06.2 tonnes for  MBS (4563 km 2 area) and 343.3 tonnes for ML (3500 km 2 area) (Fig. 4.2A).  Moose The [965 and 1971 transect surveys of the MBS gave similar results: 87 moose along 1834 km of transect, and 82 moose along 1496 km of transect, respectively. In 1987, one moose was observed in ML on 1075 km of transect covering both the MBS and ML. The 1992 moose survey gave densities of 0.12 moose/km 2 in MBS and 0.25 moose/km 2 in ML (Shank, 1992). Abundance indices based upon ground and aerial reconnaissance data from 1986-1992 also indicate lower moose numbers in MBS relative to ML (Table 4.3). The estimated available moose biomass was 46.4 tonnes for MBS (4563 km 2 area) and 90.7 km area) (Fig. 4.2A). tonnes for ML (3500 2  DISCUSSION  Whether moose numbers have declined over the entire study area from 1971 to the present is debatable. The crude line transect surveys do not provide estimates of moose density, and any density estimates derived from them would be underestimates (Gasaway  et a!., 1986). Local hunters maintain that moose numbers were historically much higher 102  Table 4.3.  Indices of moose abundance between areas: ground (moose observedlfield  week)’, aerial (moose observed/flight hour)’, survey (moose/km . ) 2  The values in  parentheses represent the number of field weeks or the number of flight hours the index is based upon. See text for more detailed explanation of indices.  Ground  Aerial  Survey  MBS  0.214 (46.7)  0.91 (30.9)  0.12  ML  0.630 (14.3)  1.57 (11.5)  0.25  2  Data collected between 1986 and 1992 2  Data from Shank (1992).  103  in the MBS in the early 1970’s, and the actual numbers seen during line transect surveys in the MBS show a marked decline. Currently moose numbers are lower in MBS than ML, and likely have been since 1986. Wood bison numbers have been increasing since the 1963 introduction. The increase has most likely had a greater potential impact on moose in MBS than ML based upon total numbers of resident bison and the length of time bison have been resident in the two areas. Decreased moose densities could have resulted from competition between moose and bison for food resources, changes in habitat, or wolf predation. Competition for food resources is unlikely.  Moose are concentrate selectors  (Belovsky, 1978; Hofmann, 1989) with diets dominated by browse species like willow (Salix spp.), aspen (Populus tremuloides), and paper birch (Berulapapynjera) (Peek, 1974; Risenhooven, 1989). In contrast, bison are classic grass/roughage eaters (Hofmann, 1989) with diets dominated by graminoids (Reynolds e. a!., 1978; Larter and Gates, 1991a; 1991b).  In exceptionally dry summers, when grass and sedge productivity is reduced  relative to more wet summers, bison have foraged extensively on willows (Larter and Gates, 1991a), but they rarely foraged on willow during winter. Consequently, any diet overlap was limited in duration and occurred during the growing season when forage quality and quantity are still highest. Habitat changes resulting in decreased willow cover could cause declines in moose numbers. Aerial photos indicate that over the past 30 years, willows have invaded dry meadow communities and willow cover has actually increased, although some of the willow cover may have grown Out of reach for moose browsing. Fires have been infrequent over the past 30 years, but have occurred more recently in ML than MBS. This could have 104  provided more early successional willow growth in ML. If moose numbers have been declining, it is unlikely that habitat changes causing a reduction in browse are responsible. Predation remains an alternate possibility. Wolves have been present in the area for many years. Local residents believe that wolf numbers are on the rise, but there are no historical data on wolf numbers. Messier and Crete (1985) have provided the best evidence that wolves regulate their prey. They documented wolf predation over a range of moose densities, and concluded that wolf predation acted in a density dependent fashion. They found at low densities of 0.22 and 0.17 moose/km , wolves were capable of regulating moose populations. 2  However, at  densities around 0.40 moose/km , moose numbers appear to be regulated by food. On Isle 2 Royale moose numbers were regulated by food even in the presence of wolves, and moose densities have fluctuated between 1.6 to 2.8 moose/km 2 (from Peterson, 1977; 1992). Messier (1991) estimated that competition for food had a regulatory effect on moose density on Isle Royale when densities were between 1 and 2 moose/km . Crete (1989) 2 estimated that densities greater than 2.0 moose/km 2 were required if carrying capacity (K) (Macnab, 1985) was to be reached in eastern Quebec. Messier (1984) suggested that a density of 0,2 moose/km 2 approximated a threshold below which wolf packs cannot subsist without an alternate ungulate prey species. Current moose densities in the MBS are 0.12 2 (Shank, 1992), i.e. below Messier’s (1984; 1993) threshold. moose/km Wolf predation on moose can potentially exacerbate a moose decline because of its antiregulatory effect (Messier, 1991). Additionally, wolf predation has a greater impact on moose populations when the moose population is declining (Gasaway el a!., 1983). My data support the contention that wolf predation on moose is destabilizing, because in areas 105  of both low and high moose density, predation on moose is greater than expected given the available prey biomass. Moose remains were found in significantly more wolf seats than would be expected given the availability of moose and bison biomass in the study area (Fig. 4.2). This was not the case for wolf kills. The proportions of bison to moose kills were similar to that expected (Fig. 4.2A) given prey availability (Fig. 4.3). The difference between data sets may be an artefact of sampling methodology. Wolf kills were generally found by groundbased travel and were much easier to spot in more open habitats. Sightability of animals and kills in forested habitats is restricted (Larter, 1988).  The majority of travel was  through open habitat patches and along travel corridors that connect these open habitat patches.  Consequently, travel was concentrated in bison habitat as opposed to moose  habitat. Wolf seats were found opportunistically, were not collected from recent kill sites and likely represent a less biased sample of wolf diet composition than kill data. Estimated relative amounts of each prey item consumed were 1.62 kg bison per 1 kg moose in MBS and 1.17 kg bison per 1 kg moose in ML. Given that the available biomass of bison is 23.8 times that of moose in MBS and 3.8 times that in ML, these data further indicate moose as the preferred prey item. Bison and moose are the two largest species of North American ungulates. The typical attack success of wolves ranges from 5% on larger prey (Haber, 1977) to 40% on smaller ungulates (Kolenosky, 1972).  Given a choice, wolves exhibit a preference for  moderate size prey species (Murie, 1944; Mech and Frenzel, 1971; Van Ballenberghe eta!., 1975; Carbyn, 1983). Only juvenile moose and bison are small, and the smaller size of 106  moose compared to bison may affect prey preference. The spatial distribution of prey can alter the effective search of wolves and regardless of prey preference may affect prey selectivity (Huggard, 1991; 1993). Clumping of prey means a lower effective search because the group, not the individual, represents a search item (Taylor, 1979; Huggard, 1991). Bison are gregarious, spending much of the year in large groups. These large groups are dominated by females, calves, and immature animals (Larter, 1988).  Moose are solitary animals (Geist, 1963; Banfield, 1977).  Consequently even though available prey biomass is skewed highly toward bison, the number of search items may be closer to parity. When bison and moose census data were converted from individuals into the number of groups (or search items) counted, there were 138 groups of bison and 79 groups of moose in MBS and 48 groups of bison and 190 groups of moose in ML.  However, there is some evidence that hunting success is  decreased for predators that prey on groups rather than solitary prey (Van Orsdol, 1984). Moose tend to be well dispersed, spatially predictable, and historically always part of a year-long prey base (Messier, 1993). Bison tend to a more clumped distribution, and being recenlty introduced may be less a vulnerable prey than moose regardless of the number of search items. The selectivity of the predator, especially in regard to the sex and age classes of the different species which they prefer to attack, can shape the prey populations in different ways (Huggard, 1991; Mills and Shenk, 1992). Consequently predation may affect bison and moose populations quite differently. My data are limited, but the majority of bison killed by wolves were juveniles, whereas the majority of moose killed were adults. If adults are taken in a greater proportion than juveniles, predation will have a greater impact 107  on the prey population because it directly influences fecundity and mortality. Wolves do not remain in areas of high prey density, but travel frequently throughout their territory (Carbyn, 1983; Huggard, 1991).  Wolves preying upon bison in Wood  Buffalo National Park (WBNP) have a long average distance between kills suggesting that they must travel frequently between herds for kills. Although previously attacked bison herds are more alert, wolves occasionally follow these herds (Oosenbrug and Carbyn, 1985). Wolves may be more successful when they encounter large herds because of the higher probability of finding weak or vulnerable individuals, however, vulnerability does not necessarily imply weaker or sick individuals. Carbyn (1983) found that very few of the animals killed by wolves had low marrow fat contents. In this study, bison killed in late winter generally had lower marrow fat contents than those killed in early winter, but none had gelatinous marrow indicative of weakened condition. Data restricted to calves may be biased towards lower marrow fat content. Average marrow fat content for different sex/age classes and for each subpopulation were unavailable. Wolves attack bison herds with calves preferentially over herds without calves (Carbyn and Trottier, 1987). This preference for juveniles and especially calves would be expected because calves are generally slower, less dangerous and less experienced with predators. Calves are easier to kill than adults (Carbyn and Trottier, 1988), however, adult bison are still taken. Van Camp (1987) found that in WBNP five packs of wolves killed and consumed 15 adults (13 females and 2 males) and four calves during a nine week period in late winter. Calf remains in scats are indistinguishable from other age classes, once calves are 3 months old and have lost their red pelage. Therefore scats collected in winter provide 108  no information upon whether the prey was juvenile or adult. However, 26 of 41 bison killed year round by wolves were calves.  Radio collar information and sequential  calves:100 adult females ratios from my study area (see Chapter 3) indicate that bison calves are preyed upon year round, but predation pressure may increase during the summer on newborn individuals. Sequential calves:100 adult females ratios during summer 1991 in the MBS show a decline between June and August: 46.7±6.7 to 36.8±11.7. During summer, radio-collared calves were lost to wolf predation at a rate of one calf every 152.7 collar-days, but during winter the rate decreased to one calf every 354.4 collar days (unpubl. data). The possible preference for more vulnerable newborn bison calves as prey during summer may have important consequences for juvenile wolf survival. It is during summer that pups are being reared and foraging radius is reduced. Wolves often den in the rasied historic shorelines of the lakebeds in the study area adjacent to the open habitats most commonly used by foraging bison in summer (C. Gates, unpubi. data, pers. obs.). Hunting success of wolves generally increases in deep snow (Kolenosky, 1972; Haber, 1977) because deep snow hinders ungulate movements. However, deep (Ca. 1 m) soft snow can be of greater hinderance to predators than their prey (D. Shackleton, pers. comm.). Snow depths of 50-5 5 cm hinder bison calves (Van Camp, 1975), whereas depths of 80-84 cm are required to hinder moose calves (Kelsall, 1969). Adult bison can still successfully forage at snow depths of 75-85 cm and snow densities of up to 0.2 g/cm 3 (Van Camp, 1975), while adult moose movements become hindered at 98-105 cm (Kelsall, 1969). Snow levels in the study area commonly reach 50 cm by mid-winter (Larter, 1988; Larter and Gates, 1991a), but rarely reach 80 cm. The greatest depth recorded was 82 cm during late winter 1991. Therefore, bison calves should be the most susceptible prey item 109  during winter. Bison calves made up a larger proportion of wolf kills than any other prey class (species, sex, and age), especially in MBS. If the wolf population has been increasing in response to an increasing alternate prey base, then the probability of a random encounter with moose may also increase. Travel between open habitat patches (bison habitat) necessitates travel through forested moose habitat. If a pack of wolves encountered moose while actively searching for bison it is unlikely that they would pass up the opportunity to make a kill, especially when the likelihood of success is high. In order to critically evaluate whether wolf predation on moose is acting in an inverse density dependent manner it is necessary to have a measure of wolf numbers (i.e. the numerical response), and a measure of killing rate per predator (i.e. the functional response). My data do not provide these definitive measurements; however they do provide a comparison of wolf activity, prey abundance, and wolf diet composition between two areas.  The index of wolf abundance (wolf activity) is different between areas and  consequently can be considered as an estimate of the numerical response of wolves to an increasing bison population over time. The wolf diet composition in conjunction with prey (both moose and bison) abundance provides a crude estimate of the functional response. Since the frequency of moose remains occurring in wolf scats is similar between areas, whereas the availability of moose is lower in the high wolf density area, wolf predation on moose is potentially increasing as the moose population declines.  This situation could  produce an inverse density dependent relationship.  110  SUMMARY  1)  During the past 20 years bison numbers have increased, especially in MBS, while moose numbers may decreased. Currently moose density in MBS is half that of ML.  2)  Wolf activity indices are higher in MBS than in ML.  3)  Bison and moose constitute the majority of the wolf diet, based upon carcass kills and wolf scat analysis, during both summer and winter.  4)  Bison make up a larger proportion of the wolf diet in MBS than in ML, however the occurrence of moose in scat is significantly greater than expected given the availability of prey biomass in both MBS and ML.  5)  Based upon scat analyses, moose make up a similar proportion of the diet in both areas. Given there is a two-fold difference in moose densities between area, wolf predation may be destabilizing and exacerbating the decline in moose numbers.  111  CHAPTER 5. GENERAL DISCUSSION  There are many examples of eruptions of ungulate populations (Leopold etal., 1947; Martin and Krefting, 1953; Caughley, 1970a). However, investigations of the mechanisms underlying plant-herbivore dynamics during the eruptions are rare. Most studies have been descriptive. Few studies have measured available food, food quality, or animal condition. The earlier hypotheses regarding plant and herbivore dynamics associated with erupting ungulate populations were formulated from perturbed systems.  Thus, the  eradication of natural predators leading to changes in herbivore numbers gave rise to the range management/overgrazing hypothesis (Dyksterhuis, 1949; Westoby eta!., 1989). This suggests a negative feedback between the grazer and plant community that leads to a declining instantaneous growth rate of the ungulate population. The introduction of exotics gave rise to the equilibrium hypothesis, which advocates that ungulate numbers reach equilibrium once the population has passed through increasing, levelling, and declining phases (Riney, 1964; Caughley, 1970a). In both scenarios, ungulate populations start from low numbers in the midst of a superabundance of suitable habitat and hence abundant resources per individual. In the few cases which have been studied in detail, mainly exotic introductions, the four-stage equilibrium model appears to hold (Caughley, 1970a). Eruptions of native ungulate populations in less perturbed systems have not followed the dynamics proposed by the four-stage model of the equilibrium hypothesis, but have resulted in an alternate two-stage model which advocates the new equilibrium level will be reached without the ungulate population going through a decline phase (Sinclair, 1979; Houston, 1982). A further hypothesis arising from the dynamics of more natural systems 112  is the facilitation and feedback hypothesis (McNaughton, 1979). This hypothesis advocates a positive feedback between the grazer and the plant community whereby grazing increases plant productivity. In my study system, a population of wood bison (Bison bison athabascae), an indigenous species, has been erupting since its reintroduction into its natural historic range. Natural predators are relatively abundant and are a key part of the system. Dispersal by bison has provided the opportunity to examine in detail the dynamics of both the plant and herbivore communities. The bison subpopulation at the leading front of the range (ML) continues to show an increasing growth rate, while the subpopulation in the core area (MBS) is stabilizing and may be starting to decrease. This provided me the opportunity to test the predictions of both equilibrium hypotheses and the facilitation and feedback hypothesis.  Although not all the predictions for one single hypothesis were met, the  predictions of the equilibrium hypothesis, in particular the Caughley-Riney four-stage model were most consistent with my data. Plant communities in willow savannas in both the MBS and ML had similar net primary production. However standing crop in willow savannas was consistently lower in MBS than ML as a direct consequence of heavier grazing pressure in MBS than ML. This finding was consistent with the predictions of both hypotheses, but not the assumption of the facilitation and feedback hypothesis that grazing had a positive effect on production. The higher proportion of unpalatable species found in willow savannas in MBS is consistent with both equilibrium hypotheses but contrary to the predictions of the facilitation and feedback hypothesis. There has been no establishment of grazing lawns (cf. McNaughton, 1984) in the MBS unlike a variety of other grazing systems (for example 113  Prins et a!., 1980; McNaughton, 1986; Tomlinson, 1986; Hik, et a?., 1992).  Because  productivity was not measured, I could not detect whether grazing stimulated productivity especially in the short term (C. Skarpe, pers. comm.). Preliminary estimates of productivity in willow savannas were obtained for the MBS in 1986-87, however the results were equivocal (Smith, 1990). There was some indication of short term increases in productivity in a few sites, but the majority of sites showed no increase (Smith, 1990). Short term increases in productivity may be of little consequence for this system, but increases in the quality of grazed forage may be important. Preferred plant species are heavily grazed, but they are not becoming a larger component of willow savannas.  If productivity were increasing I would expect to see  differences in nitrogen content of similar plants between MBS and ML because of more rapid cycling of nitrogen and increased nitrogen uptake by plants (Thaine, 1954; Reuss and McNaughton, 1984; McNaughton and Chapin, 1985; McNaughton e. al., 1988). My data do not support this. The lack of differences in the quality of available forage between areas would indicate that differences in the demographics of the two subpopulations of bison are not caused by variation in forage quality. Juvenile survival and recruitment and physiological condition indices (faecal nitrogen) were higher in ML than MBS. These findings are all consistent with the predictions of the equilibrium hypothesis but contrary to the predictions of the facilitation and feedback hypothesis. The MBS subpopulation appears to have peaked in 1989, and exhibited a negative instantaneous growth rate (r) between 1989 and 1992. This finding is consistent only with the predictions of the four-stage model of the equilibrium hypothesis. It could be argued 114  that the negative growth rate is not associated with a major decline in subpopulation numbers, but represents a random fluctuation around a new higher equilibrium, therefore supporting the two-stage equilibrium model.  Continued monitoring of subpopulation  numbers will be required in order to see if the decline continues; if it does this would refute the argument. Additionally, since r includes emigration and immigration, emigration of individuals from the MBS to the ML subpopulation may have contributed to the negative growth rate.  The impact of emigration on the growth rate is important because the  mechanism of population decline is different from that associated with the various models I tested in this study. Continued monitoring of radio collared individuals, and increasing the number of marked individuals will be necessary in order to determine the effect of emigration on subpopulation growth rates and should be incorporated in future study. The facilitation and feedback hypothesis predicts that animals in the MES subpopulation have a consistently higher quality of diet whereas both equilibrium hypotheses predict a consistently lower quality diet. During summer, animals in the MBS did indeed have a better diet (in agreement with the facilitation and feedback hypothesis) even though the availability of preferred forages was lower (contrary to the facilitation and feedback hypothesis). I may have used an improper measure of forage preference, and the availability of preferred forages may have been higher. This is unlikely because previous studies have indicated a preference for monocots over browse, and within the monocots, for sedges over grasses (Larter, 1988; Smith, 1990; Larter and Gates, 1991a; 1991b). Bison switched from summer diets high in sedge to diets high in grass and browse only when availability of sedge was significantly reduced (Larter and Gates, 1991a). During winter, animals in IvIL had better diets and the availability of the same preferred forages was 115  higher in ML than in MBS. It seems unlikely that forage preference changes from summer to winter for just ML animals. Given the 8-month winter, diet quality of ML animals is superior to that of MBS animals on an annual basis. Integrating the predator trophic level into the hypotheses I have tested has yet to be achieved. It has been assumed that these models and their predictions can be generalized in the presence or absence of predators.  These hypotheses employ the underlying  assumption that primary productivity, in particular productivity of palatable forage, is the main regulatory mechanism. In other words, the ecosystem is thought of as bottom-up driven. An alternate top-down view, where predation plays the key role in regulating the system, was originally proposed by Hairston  el  a?. (1960).  Predation is undeniably a  limiting factor; whether it is a regulating factor is still a matter of debate (Sinclair, 1989; Messier, 1993). This study does not address whether predation is a regulating factor for the bison subpopulations because data on wolf densities are lacking, and the issue is left for future work. However, predation may be a factor in the MBS subpopulation decline. Schmitz (1992) has predicted that food chain dynamics in low to moderately productive systems would generally be controlled by top-down rather than bottom-up forces.  If top-down processes regulate this system, could this explain some of the  contradictory results? In particular, could this explain why bison inhabiting areas where high quality, preferred forage was more abundant, consistently ate poor quality forage? This finding not only runs contrary to the predictions of all three hypotheses, but also contradicts classic foraging theory (Stephens and Krebs, 1986). In a hypothetical top-down controlled system where the herbivore population exhibits logistic growth, removal of the predator should result in a trophic cascade (Schmitz, 1992; Strong, 1992), i.e. there should 116  be an increase in the herbivore abundance with a resulting decrease in plant abundance. Also, without natural predators, herbivores will not need to be vigilant when foraging thus providing the opportunity to forage more selectively and/or more efficiently (Schmitz, 1992). Bison inhabiting ML have lower predation pressure than bison inhabiting MBS. If this is a top-down controlled system, there has been a removal of predation pressure at ML relative to MBS, and there should be evidence of a trophic cascade in ML. However, bison inhabiting ML have not shown any increase in selectivity for preferred summer food items, they have shown the opposite, even with higher availability of preferred forages. These results are confounded because the MBS subpopulation is stabilizing while the ML subpopulation is still increasing and direct comparisons cannot be made until both subpopulations are stabilizing.  Top-down regulation of this system cannot explain the  contradictory results vis-a-vis foraging selectivity. There is a developing consensus that bottom-up and top-down forces act simultaneously on populations and communities, and that the dichotomy between the two regulatory mechanisms is artificial (Oksanen el al., 1981; Hunter and Price, 1992). The extent to which the relative strengths of these forces act on different systems, and how these forces change in relation to differing population densities presents interesting avenues of future research. As I described in Chapter 3, there was a strong link between plant and herbivore dynamics in both study sites of this system. However, as I alluded to in chapter 4, the impact of predation and its effect on the system dynamics may vary between study sites. This variance may be directly responsible for the counterintuitive results regarding diet quality. 117  The potential for an increasing bison population to modify the ungulate prey base available to predators has important implications where there are few prey species for predators. Increasing bison numbers could either decrease predation on an alternate prey population (here moose) by diluting the functional response, or increase predation by creating a numerical response. Because of the relative vulnerabilities of bison and moose, it appears that the latter may be occurring in this system. Thus, increasing bison numbers may exacerbate the decline of moose numbers.  Reintroductions of wood bison have been  proposed for northern Canada and Alaska, and there is an abundance of suitable habitat (C. Gates and B. Stephenson, pers. comm.). 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