Abstract

Climate change will likely alter the distribution and abundance of northern mammals through a combination of direct, abiotic effects (e.g., changes in temperature and precipitation) and indirect, biotic effects (e.g., changes in the abundance of resources, competitors, and predators). Bioenergetic approaches are ideally suited to predicting the impacts of climate change because individual energy budgets integrate biotic and abiotic influences, and translate individual function into population and community outcomes. In this review, we illustrate how bioenergetics can be used to predict the regional biodiversity, species range limits, and community trophic organization of mammals under future climate scenarios. Although reliable prediction of climate change impacts for particular species requires better data and theory on the physiological ecology of northern mammals, two robust hypotheses emerge from the bioenergetic approaches presented here. First, the impacts of climate change in northern regions will be shaped by the appearance of new species at least as much as by the disappearance of current species. Second, seasonally inactive mammal species (e.g., hibernators), which are largely absent from the Canadian arctic at present, should undergo substantial increases in abundance and distribution in response to climate change, probably at the expense of continuously active mammals already present in the arctic.

INTRODUCTION

Bioenergetics describes the process by which food resources are acquired, assimilated, and allocated among maintenance, growth, and reproduction (McNab, 2001). Accordingly, energetics serves as a primary linkage between physiological processes of individuals and the ecological patterns of populations (Odum, 1963; Yodzis and Innes, 1992). By definition, energetics focuses on the acquisition and allocation of energy, and indeed, energy is widely accepted as the primary limiting resource in most terrestrial systems (McNab, 2001). Additionally, many of the general principles employed in energetics (e.g., mass-balance equations, trophic linkages) are equally applicable to interactions limited by other resources.

The earth is in the midst of a pronounced warming trend, and more substantial changes in temperature, precipitation, ocean circulation, and ice dynamics are expected during the next century (Houghton et al., 2001). As evidence for widespread biological impacts of recent climate change accumulates (Parmesan and Yohe, 2003; Root et al., 2003), there is growing demand to anticipate the future responses of plant and animal populations to ongoing environmental change (Ludwig et al., 2001). The ability of biologists to anticipate these biotic responses is limited to some degree by lingering uncertainty about the underlying causes of climate change and how regional climates will be affected by the complex, interactive effects of global changes in temperature, precipitation, and circulation patterns (Houghton et al., 2001). Nevertheless, arguably the greater uncertainty is how biological communities will respond to even the most robust and widely accepted components of projected climate change. Besides, as we hope to illustrate in this review, temperature and seasonality are so fundamental to our understanding of the organization of biological communities, that further investigation of their effects can provide biological insights quite independent of issues related to climate change.

The impacts of climate change on animal populations are likely to include both direct and indirect effects. For the purposes of this review, we consider direct effects of climate change to be those resulting from changes to a population's abiotic environment. Hence, direct effects of climate change on animal populations might include altered mortality or reproductive rates resulting from changes in ambient temperature, the frequency and severity of extreme weather events, the depth and duration of snow and ice cover, and the availability of surface water. In contrast, we consider indirect effects to include those resulting from changes in a population's biotic environment, specifically the distribution and abundance of their resources, predators, and competitors. Because patterns of individual energy intake and expenditure represent the point of integration for abiotic and biotic influences, energetic approaches are uniquely suited to prediction of both direct and indirect effects of climate change.

In the current review, we illustrate three ways in which energetic approaches can be used to predict the impacts of climate change on mammal populations in northern environments. The first of these relies on biogeographical patterns of environmental energy availability and biodiversity to infer how climate change will impact regional species diversity. The second focuses on species that routinely experience seasonal periods of negative energy balance, and quantifies direct effects of climate on the duration and severity of these seasonal energetic bottlenecks to predict how climate change will modify species range limits. The third uses trophic energy flow models to predict the combined impacts of direct and indirect effects of climate change on simplified northern food webs. These same approaches could be applied to terrestrial vertebrates other than mammals, but we do not do so here because amphibians and reptiles are generally absent or present at low densities in subarctic and arctic regions, whereas birds are abundant, but their local dynamics are complicated by migration and allochthonous energy inputs (Jefferies and Rockwell, 2004; Gauthier and Bêty, 2004). Furthermore, by focussing on energetics in this review, we do not mean to imply that non-energetic factors, such as disease, stress, or abiotic habitat modification, will be unimportant in shaping wildlife responses to climate change. Instead, we view energetics as simply a uniquely appropriate starting point for predicting climate change impacts, establishing an “all-else-being-equal” platform of expected change that studies of other non-energetic factors will surely modify.

I. SPECIES-ENERGY PATTERNS AND REGIONAL DIVERSITY

Regional biodiversity varies dramatically with latitude (MacArthur, 1965; Rosenzweig, 1995). Animal and plant diversity tends to increase from the equator to the tropics, then decline precipitously from tropical to polar latitudes (Rosenzweig, 1995). According to the species-energy hypothesis (Currie, 1991), variation in radiant energy flux per unit area is the primary determinant of the tropical-polar diversity gradient. High potential energy fluxes characteristic of warm, tropical regions are speculated to translate into high primary productivity, which in turn supports complex, species-rich food webs. In contrast, low potential energy fluxes characteristic of cool, polar regions are suggested to translate into low primary productivity, which in turn supports simple, species-poor food webs. Consistent with this hypothesis, general climatic variables reflective of environmental energy availability, such as temperature and annual potential evapotranspiration, account for a large proportion (i.e., typically >75%) of regional variation in species richness (Currie, 1991; Badgley and Fox, 2000). Recent literature emphasises the extent to which these general ecological patterns may correspond to basic temperature-dependent biochemical and metabolic processes (Gaston, 2000; Allen et al., 2002).

The well-studied biogeography of North American mammals conforms to general species-energy patterns (Simpson, 1964; Currie, 1991; Badgley and Fox, 2000). The species density of North American mammals in 58,275 km2 quadrats varies from greater than 180 in southern Mexico to less than 20 in the most northerly continental regions of Canada (Fig. 1; Badgley and Fox, 2000). Five environmental variables, representing seasonal extremes of temperature, annual energy and moisture, and elevation, account for 88% of this continental variation in species density (Badgley and Fox, 2000). Within Canada, mammal species density varies from greater than 80 in southern British Columbia to less than 10 in the arctic archipelago, and 74% of this variation is explained by mean annual temperature alone (Fig. 2; Kerr and Packer, 1998).

The species-energy hypothesis provides an empirical basis for predicting the future impacts of climate change on regional biodiversity (Kerr and Packer, 1998; Currie, 2001). Kerr and Packer (1998) used the empirical relationship between annual average temperature and mammal species density across Canada (Fig. 2) to predict how projected climate change might affect regional mammal diversity. Projected temperature changes over a 75 yr time frame were derived from the Goddard Institute of Space Science (GISS) global circulation model (GCM; Russell et al., 1995), assuming a 1% annual CO2 increase. According to this and most other GCM's (e.g.,Johns et al., 1997; Boer et al., 2000; Dai et al., 2001), annual average temperature is expected to increase over the next 75 yrs in most continental regions of Canada. Based on current species-energy patterns, Kerr and Packer (1998) predict a corresponding increase in mammal diversity across Canada, with the largest proportional increases occurring in far northerly regions where current mammal diversity is lowest and projected temperature increases are greatest (Fig. 3).

Predicted impacts of climate change on biodiversity derived from current species-energy patterns must be interpreted with caution. The approach relies on the assumption that the largely unknown mechanisms presently linking radiant energy input and species richness will not change in concert with climate (Currie, 2001). Furthermore, the predictions relate only to a given environment's potential to support high levels of diversity, not how realized species richness will vary over a fixed timeframe. The latter depends critically on the time lag between environmental change and biotic responses to that environmental change (Davis, 1986; Malcolm et al., 2002). Finally, the species-energy approach provides no indication of whether potential increases in boreal and arctic diversity will be realized by the addition of new species to existing communities or by wholesale community replacement. Community replacements in proximity to geographical barriers (e.g., oceans) may leave many northern-adapted species with “nowhere to go” (p. 268; Kerr and Packer, 1998; Hansell et al., 1998).

Despite the limitations of the species-energy approach in predicting the realized impacts of climate change on regional biodiversity, it provides an important and robust null expectation. Boreal forests and arctic tundra are among the coldest and least productive terrestrial environments on earth (Roy et al., 2001). Climate change projections generally concur that these regions will undergo unprecedented warming during the next century (Houghton et al., 2001). Global patterns of environmental energy availability, productivity, and biodiversity indicate that this warming will enhance the potential for increased diversity in the region (Currie, 1991; Gaston, 2000). As a result, the environmental impacts of climate change in temperate and polar regions are likely to be shaped by the appearance of new species at least as much as by the disappearance of current species.

By what mechanisms will the expected increases in boreal and arctic mammal diversity occur? Over the long-term, appearance of new species may result from speciation and adaptive radiation. But in the short-term, poleward range expansions of temperate-zone species will provide the major mechanism by which boreal and arctic diversity might increase. In the next section, we describe how energetics can be used to predict the extent of poleward range expansions for some mammal species.

II. SEASONAL ENERGETIC BOTTLENECKS AND RANGE LIMITS

Many mammals routinely experience prolonged, seasonal periods of negative energy balance when energy requirements exceed energy intake (Robbins, 1993). To survive, individuals must accumulate energy reserves in advance of the period that are large enough to support the deficits incurred. Thus, the length and severity of such seasonal energetic bottlenecks may serve as a direct constraint on what environments mammal populations can and cannot occupy. If this is the case, the considerable complexity normally involved in linking individual energetics to population-level processes (e.g.,Moen et al., 1997) becomes reduced to three simple factors; the size of the energy reserves accumulated prior to the bottleneck, the rate at which the reserves are depleted during the bottleneck, and the length of the bottleneck. If the size is less than the rate times the length, individuals cannot survive the bottleneck and populations will be unable to persist.

For mammals in boreal and arctic regions, seasonal energetic bottlenecks typically occur in winter, when thermoregulatory requirements are elevated and food resources are scarce. Winter survival requires that sufficient energy can be stored in summer and autumn to support the energetic deficit accumulated through to spring. If the energy reserve accumulated prior to the bottleneck is smaller than the deficit incurred during the bottleneck, the individual will run out of energy and die before spring. Some mammals accumulate a food hoard to support requirements during energetic bottlenecks (Vander Wall, 1990; Humphries et al., 2002a), but the majority of mammals rely solely on body energy reserves. We focus on mammals in the latter category for the remainder of this section.

Fat is the predominant body energy reserve used by mammals to fuel metabolism during seasonal bottlenecks. Because lipids yield much more energy per gram stored than do proteins, mammals rarely maintain non-essential protein reserves (Speakman, 2000). Thus, the size of fat stores alone provides a good indication of the energy reserves available to an individual. Due either to morphological constraints or costs associated with fat storage, the maximum size of fat stores is consistently 40%–50% of total body mass in both small and large mammals (Millar and Hickling, 1990; Humphries et al., 2003). Given that lipids generate 37.3 kJ energy per gram metabolized, as a rule of thumb, the maximum amount of energy available in fat (measured in kJ) at the onset of a seasonal energetic bottleneck will equal 37.3 × 50% of total body mass (measured in grams), or approximately 19 × total body mass. Detailed field studies of pre-bottleneck fat accumulation allow more precise estimates of maximum reserve size for particular species (e.g.,Kunz et al., 1998).

The rate at which fat stores are depleted during energetic bottlenecks depends on resting rates of metabolism, additional costs of thermoregulation, and the capacity for metabolic depression. Resting metabolism scales approximately to mass0.75 in mammals, such that large mammals expend less energy per unit mass than small mammals. Because fat storage capacity increases proportionately with body mass (mass1), the fasting endurance of euthermic mammals generally increases according to body mass0.25 (Lindstedt and Boyce, 1985; Millar and Hickling, 1990). Thus, while large mammals can potentially support resting metabolic requirements throughout winter using only fat reserves, small mammals can support resting requirements with body fat for less than a month. Furthermore, maintaining a constant, elevated body temperature in winter requires additional thermoregulatory expenditure beyond resting requirements for most mammal species. Thermoregulatory requirements are especially high for small mammals because of their higher lower critical temperatures and higher thermal conductance, relative to larger mammals (McNab, 2001). Once thermoregulatory requirements are factored in, the winter fasting endurance of most small mammals reduces to only a few days, while that of larger mammals continues to extend several months. Thus, it is not surprising that many small mammals (and only a few large mammals) have evolved the capacity to reversibly enter a state of metabolic depression that can extend their fasting endurance beyond the length of the seasonal energetic bottlenecks they experience.

Mammal hibernation provides an especially clear instance where winter survival depends critically on sufficient pre-winter reserve accumulation and where species range limits might be imposed by seasonal energetic bottlenecks. Following termination of activity in late summer or autumn, most hibernators rely completely on stored fat until spring emergence (Lyman et al., 1982; French, 1992; Humphries et al., 2003). Fasting endurance is extended by allowing body temperature to drop to near ambient levels, which reduces metabolic rate to less than 10% of euthermic levels (Geiser, 1988; Heldmaier and Ruf, 1992). However, even when in hibernation, mammals remain constantly in a metabolically and thermally regulated state, alternating between brief bouts of euthermy, when body temperatures are regulated at normal, elevated levels, and prolonged bouts of torpor, when body temperatures are regulated at a much lower torpor set-point (Lyman et al., 1982; Nedergaard and Cannon, 1990). Energy requirements when euthermic and torpid, as well as the frequency of arousals, vary strongly with ambient temperature.

Recent research on the hibernation energetics and biogeography of the little brown bat (Myotis lucifugus) provides an example of how seasonal energetic bottlenecks can limit range distributions and facilitate prediction of range expansions due to climate change (Humphries et al., 2002b). M. lucifugus is a small (body mass 6–10 g), insectivorous bat that hibernates in caves and abandoned mines, relying entirely on stored body fat during the prolonged winter period when flying insects are inactive. Integration of the well-quantified hibernation energetics of this species reveals that total winter energy requirements vary substantially with winter length and hibernaculum temperature (Humphries et al., 2002b; Fig. 4). Because relatively simple thermal relationships determine the length of the hibernation period and the maximum cave temperatures available for hibernation, both variables can be readily estimated from regional climate data. Comparison of maximum fat storage capacity and the projected total hibernation energy requirements of M. lucifugus in different regions of Canada indicates that this species is excluded from regions where winters are too long and too cold to permit successful hibernation with maximum fat reserves (Humphries et al., 2002b; Fig. 5). To evaluate how this energetically-imposed range limit would be altered by future climate change, temperature projections from a global climate model (Hadley Centre Coupled Model 2 with aerosol simulation mean HadCM2—GAX; Johns et al., 1997) were used to forecast the energetic costs of bat hibernation in 2080. In most regions, the winter range limit of M. lucifugus is expected to move northward by about 6 km per year over the next eighty years (Humphries et al., 2002b; Fig. 5).

The energetic bottleneck approach used to predict the impacts of climate change on the northern range limits of hibernating bats is equally applicable to the many other northern mammals that experience seasonal energetic bottlenecks. Although parameterization of the model for M. lucifugus was facilitated by the relatively simple thermal relations that define the duration of their hibernation period and the hibernaculum temperatures to which they are exposed, the basic logic and structure of the model applies equally to arctic ground squirrels (Spermophilus parryii) hibernating on Alaska's north slope (Buck and Barnes, 1999a), caribou (Rangifer tarandus) relying on a combination of foraging and fat reserve depletion to survive on wintering grounds (Russell et al., unpublished), and land-locked polar bears (Ursus maritimus) fasting on the shores of Hudson Bay during ice-free periods (Derocher et al., 2004). In the final section, we evaluate the implications of these and other energetic adaptations in a broader, food web context, seeking to predict the combined outcome of direct and indirect effects of climate change on energetically differentiated mammal populations.

III. TROPHIC ENERGY FLOWS

Animal populations are frequently excluded from potentially suitable habitats by a scarcity of resources or an abundance of competitors and predators. In these situations, the impacts of climate change on biogeographic distributions will involve a combination of direct and indirect effects. Trophic energy flow models provide a useful conceptual approach to predict how direct effects of climate on one component of a food web will translate into indirect effects on other components. Because most northern mammals are either herbivores that feed on plants, or carnivores that feed on herbivores, their community interactions are reasonably represented by three-link trophic food webs. We first consider how climate change might alter the total biomass occupying these three levels via trophic cascades, before evaluating what functional groups within a given trophic level may show opposing responses to climate.

IIIa. Trophic cascades and the ecosystem exploitation hypothesis

According to the hypothesis of exploitation ecosystems (Oksanen et al., 1981; Oksanen and Oksanen, 2000), a combination of low primary productivity and high endotherm energy requirements severely limits the number of sustained trophic levels in northern food webs. At the low end of northern productivity gradients, plant growth occurs to such a limited extent that no viable herbivore population can be supported (Fig. 6, Zone I). At slightly higher levels of productivity, herbivores are able to persist but carnivores are not, and thus herbivores impose strong grazing pressure on vegetation (Fig. 6, Zone II). At still higher levels of productivity, carnivores begin to persist and strongly regulate herbivore densities, thereby releasing vegetation from intense grazing pressure (Fig. 6, Zone III). Because terrestrial systems generally do not exceed three trophic levels (terrestrial carnivores specializing on other terrestrial carnivores are rare), the bracketing of herbivore dynamics between vegetation and carnivore regulation may account for the predominance of plants in terrestrial systems (green world hypothesis, Hairston et al., 1960). Although there is considerable disagreement about the actual productivity levels and geographic localities that conform to the different zones of Oksanen's ecosystem exploitation hypothesis (Oksanen and Oksanen, 2000; Gauthier and Bêty, 2004), the basic hypothesis that primary productivity plays a major role in trophic organization is widely accepted.

There is general agreement that summers will become longer, warmer, and wetter throughout most of northern Canada over the next century (Johns et al., 1997; Boer et al., 2000; Dai et al., 2001; Houghton et al., 2001) and that above-ground primary productivity will increase in accordance with these long-term trends (Hansell et al., 1998; Aber et al., 2001; Currie, 2001; Henry, unpublished). In the absence of strongly negative, direct effects of climate change on mammalian herbivores and carnivores, the population responses of mammals should be determined primarily by trophic position as predicted by the ecosystem exploitation hypothesis. Specifically, in response to increased primary productivity associated with climate change, herbivores should 1) begin to occupy the most productive zone I regions from which they are currently excluded, 2) increase in abundance in zone II regions where they currently occur, and 3) maintain similar levels of abundance in zone III regions where carnivores are present (Fig. 6). Carnivores, on the other hand, should begin to occupy the most productive zone II regions from which they are currently excluded (and, as a result, regulate the herbivore biomass in these regions) and increase in abundance in zone III regions (Fig. 6). Strongly negative, or highly divergent, direct effects of climate change on herbivores and carnivores could of course impede these anticipated responses. Thus, a key avenue for future research involves evaluating the direction and magnitude of climate change's direct effects on not only individual populations, but also on trophic levels in aggregate.

At a finer level of resolution, embedded within aggregate trophic-level responses, component populations are likely to respond differentially to climate change. The response of a given population will depend on the direction and strength of both the direct and indirect effects influencing its dynamics. For example, caribou (Rangifer tarandus) populations may decline as a result of climate change, if benefits resulting from increased productivity of plant resources are overwhelmed by reductions in the accessibility of resources due to increased depth and hardness of winter snow cover (Caughley and Gunn, 1993). However, given that direct effects are likely to differ among populations occupying the same trophic level, the decline of one population may be compensated by population increases of others. Thus, although caribou may be negatively impacted by increased snow cover, lemmings (Lemmus, Dicrostonyx spp.) and other small mammals are likely to be positively impacted (Reid and Krebs, 1996). The latter populations would therefore experience a triple advantage from a climate change scenario of increased resource abundance and increased snow cover—more resources, better access to resources, and reduced competition from populations negatively impacted by snow cover—and can be expected to increase in abundance as a result. Thus, while it is clearly critical to consider the possibility of direct, population-specific effects when making community-level predictions, it is equally important that the ramifications of population-specific effects be considered within their broader trophic context. Frequently what is bad for one population will tend to be good for another. Thus, a second key avenue for future research involves identification of different functional groups within trophic categories that will exhibit opposing population responses to climate change. We identify one such functional group classification in the next section.

IIIb. Apparent competition and coexistence along a continuum of seasonal strategies

Boreal and arctic ecosystems are characterized by short summer pulses of primary productivity, interspersed by prolonged winter periods of cold temperatures and vegetation senescence or growth arrest. Northern winters are a major energetic challenge for endotherms because snow cover and low ambient temperatures simultaneously reduce resource availability and elevate resource requirements (King and Murphy, 1985; Davenport, 1992; Robbins, 1993; Speakman, 2000). Endotherm adaptations to the resulting seasonal energetic bottleneck include anticipatory energy storage (Humphries et al., 2003), increased insulation (Scholander et al., 1950; McNab, 2001), refuge occupation (Huey, 1991), torpor expression (Lyman et al., 1982) and migration (Dingle, 1996).

Differential expression of seasonal energetic adaptations generates a continuum in the trophic status of endotherms during winter. At one extreme are species that avoid the bottleneck via hibernation in refuges (temporal avoidance) or migration out of the region (spatial avoidance), and thus completely decouple from the food web during winter. Because terrestrial mammals are either non-migratory or do not migrate sufficient distances to truly decouple from seasonal environments (Dingle, 1996), and because migration introduces the added complexity of allochthonous energy inputs into local systems (Jefferies and Rockwell, 2004), we limit our consideration here to mammals expressing in situ winter inactivity. We refer to these species as seasonal. At the other extreme of the seasonality continuum, are species that simply tolerate the bottleneck, and thus exhibit unchanged or even increased resource consumption and predator susceptibility during winter. We refer to these species as continuous. Populations occupying different positions along this continuum should be differentially affected by the moderated seasonality that climate change will introduce to northern food webs.

Because winters become increasingly severe at higher latitudes, seasonal strategies might be expected to predominate among the mammals present in the far north. To test this prediction, we considered all herbivorous mammal species occupying the Deciduan, Coniferan, and Tundran subregions of North America as delineated by Hagmeier (1966). Using Banfield (1974) and Wilson and Ruff (1999), we qualitatively ranked the extent to which each of these species decouple from resources and predators during winter. The lowest seasonality rank (SR = 1) was assigned to species characterized by no significant declines in resource consumption and predator susceptibility during winter (e.g., hares and ungulates). The highest rank (SR = 4) was reserved for species with substantial pre-winter energy storage, greatly reduced predator susceptibility during winter, and a capacity to substantially depress metabolism via torpor expression (e.g., ground squirrels and chipmunks). Intermediate ranks were applied to species with small to moderate reductions in either resource consumption or predator susceptibility during winter (SR = 2; e.g., lemmings and red squirrels) or with small to moderate reductions in both (SR = 3; e.g., beaver and pika).

In contrast to the prediction that seasonal energetic strategies will predominate where winter conditions are harshest, seasonal mammalian herbivores are more common in southern and central North America than in northern North America. Highly seasonal species (SR = 3 or 4) account for approximately one third of Deciduan and Coniferan mammals but less than 10% of Tundran mammals (Table 1). The only highly seasonal mammal occupying tundra communities is the Arctic ground squirrel (Spermophilus parryii; SR = 4), and even this species occurs only in the southern portion of the Tundran subregion. Accordingly, highly seasonal mammals form a major component of mammalian herbivore communities in boreal forests, but are entirely absent from high arctic communities (Table 2). Migratory birds and diapausing insects are present throughout the arctic (Gauthier and Bêty, 2004; Danks, 2004), and thus overall herbivory in arctic systems may remain highly seasonal. However, for mammals, which are unable to migrate as far as birds or develop as rapidly as insects, high arctic summers are apparently too short and winters too long for a seasonal strategy to be viable.

Seasonal mammals may be absent from the far north because of purely abiotic limitations or because of competitive exclusion by continuous species. In Section II, we identify a situation where the northern range limit of a seasonal species (Myotis lucifugus) appears to be imposed by abiotic constraints involving an interaction between maximum autumn energy reserves and minimum winter energy requirements. Permafrost barriers to subterranean thermal refugia may represent an additional constraint on subterranean hibernators in the arctic (Buck and Barnes, 1999b). Conversely, seasonal species may be excluded from otherwise suitable habitats by the presence of continuous competitors that suppress their resources and support populations of shared predators (Holt and Lawton, 1994). Competition between seasonal and continuous species is a key consideration even in the case of abiotic exclusion of seasonal species, because if climate change were to eliminate or weaken the abiotic constraint, then competition outcomes with continuous herbivores would become the key determinant of the invasibility of seasonal populations.

For seasonal and continuous herbivores to coexist under resource competition and shared predation from carnivores, the relative competitive advantage of the two strategies must reverse with season (Namba, 1984; Chesson, 2000). This trade-off could be structured in one of two ways. Relative to continuous herbivores, seasonal herbivores could experience greater biomass increases during summer (because of higher maximum consumption rates, higher reproductive output, and/or lower predation rates) and greater losses during winter (because of fasting and/or higher winter mortality). Conversely, seasonal herbivores could experience lesser biomass increases during summer (because of lower maximum consumption rates, lower reproductive output, or higher predation rates) and lesser losses during winter (because of lower energetic costs or lower predation rates). Whichever population experiences a relative competitive advantage during summer (i.e., growth maximizers) should be characterized by greater seasonal fluctuations than the population garnering a competitive advantage during winter (i.e., loss minimizers).

Although either form of this seasonality trade-off is sufficient to maintain long-term coexistence in a seasonal environment, the distinction is critical because it will determine how continuous and seasonal species will respond to climate change. Whichever strategy currently garners a competitive advantage during summer (i.e., is characterized by a proportionately higher spring to autumn biomass increase) will benefit from the longer and warmer summers projected to occur in the future, at the expense of the strategy currently garnering a competitive advantage during winter (i.e., characterized by proportionately lower autumn to spring biomass decreases). Thus, if seasonal species currently undergo more pronounced seasonal fluctuations in biomass than continuous species do, then seasonal populations should increase in response to climate change while continuous populations decline. This prediction should hold even if continuous and seasonal herbivores are characterized by less than complete dietary overlap, so long as climate change effects on the duration of seasons are not perfectly confounded by differential climate change responses of their respective plant resources (Grover, 1995).

Current latitudinal distributions of seasonal and continuous mammals suggest that seasonal species will benefit most from climate warming. Given that seasonal mammals currently occur everywhere in northern North America except for the regions characterized by the longest winters and shortest summers, seasonal species appear either to be abiotically excluded from northern environments or to garner a competitive advantage relative to continuous species as a result of longer summers and shorter winters. Thus, in the future, as northern Coniferan and southern Tundran summers lengthen and winters shorten, we can expect a shift in the relative competitive advantage of the two strategies, with seasonal herbivores increasing in abundance in regions where they currently occur (Fig. 7, Zone IIIb) and expanding their ranges into regions where they are currently excluded (Fig. 7, Zone IIIa).

Modifying Oksanen's productivity zones to incorporate the functional distinction between seasonal and continuous species, continuous herbivores are the first mammals to appear along a north-to-south (low-to-high) productivity gradient, followed by continuous carnivores, followed by seasonal herbivores (Fig. 7). Zone IIIa of Figure 7 is thus representative of vast regions of the Canadian arctic where resident carnivores (e.g., arctic foxes and wolves) have an important influence on continuous herbivore dynamics, and where seasonal mammal herbivores are absent (Gauthier and Bêty, 2004; Krebs et al., 2003). Zone IIIb of Figure 7 is representative of many other regions of northern North America where continuous and seasonal herbivores coexist under shared predation from continuous carnivores.

Concomitant with the predicted northward spread of seasonal herbivores due to climate change, continuous herbivores already present in the Canadian arctic are likely to experience population declines for several reasons. First, increased summer length will reduce the proportion of the annual cycle during which continuous species experience a relative competitive advantage over seasonal species. Second, increased primary productivity will be monopolized by seasonal species that are at a competitive advantage in summer when vegetation growth occurs. Third, increased carnivore biomass, resulting from productivity cascades, will have disproportionate top-down control on continuous herbivores because their seasonal competitors become unavailable to predators during winter. This latter outcome reflects an important cryptic benefit of a seasonal energetic strategy. Although evaluation of the advantages of hibernation are often limited to energetic considerations (Humphries et al., 2003), the capacity for a population to avoid supporting its predators for prolonged seasonal periods and to transfer this burden to its competitors, may represent a major ecological benefit of hibernation.

Before concluding, we briefly consider the seasonality of mammalian carnivores along the North-South productivity gradient. The few mammalian carnivores present in northern systems (e.g., wolf, foxes, lynx, wolverine, weasels) tend to be continuously active, whereas several carnivores limited to more southerly regions are seasonally inactive (e.g., badger, skunks, racoons). Canids and, to a lesser extent, felids are present in the Tundran, Coniferan, and Deciduan subregions always as continuous carnivores. All Tundran and northern Coniferan mustelids are active throughout winter (e.g., wolverine, Gulo gulo; fisher and marten, Martes spp.; weasels and mink, Mustela spp.), whereas several mustelids present in the southern Coniferan and throughout the Deciduan are characterized by varying extents of winter dormancy (badger, Taxidea taxus; skunks, Mephitis spp., Conepatus leuconotus, Spilogale putorius). The occurrence of hibernating grizzly bears (Ursus arctos) in southern Tundran regions is a partial exception to this carnivore trend, but even within North American bears the general pattern holds, with the most northerly species (polar bear, Ursus maritimus) characterized by less seasonality than its southern congenerics (U. arctos, U. americanus). Thus, the northern range limits of seasonal carnivores may result from similar abiotic or biotic constraints affecting seasonal herbivores, albeit at different thresholds of seasonality and productivity.

SUMMARY

The causes and projected consequences of global climate change are not yet fully resolved, and it remains unclear how animal populations will be affected by even the most robust and widely accepted components of projected climate change. Although mammals are characterized by extraordinarily wide thermal tolerance and possess highly specialized adaptations for dealing with seasonal temperature variation, they, like all animals, remain fundamentally affected by temperature. Given that climate change will substantially alter ambient temperature patterns in northern regions, as well as change many other components of the abiotic (e.g., precipitation, snow and ice cover) and biotic environment (e.g., abundance and distribution of resources and predators), the mammal fauna of northern Canada is likely to undergo major changes during the next century.

Additional data and better theory are needed to reliably predict the direction and magnitude of climate change impacts on specific mammal species. In this review, we have therefore attempted only to illustrate how bioenergetic approaches relate to climate change impacts at various levels of resolution (e.g., regional diversity, species range extensions, trophic interactions) and to generate hypotheses that might motivate additional research. Further progress in predicting the impacts of climate change on northern mammals awaits integration of the conceptual approaches outlined in this review with detailed field studies of specific northern communities. Thus, rather than concluding with speculative predictions about possible impacts of climate change on particular northern mammals, we conclude with four general points that are central to, and have emerged from, the bioenergetic approaches we employed:

  1. Current global patterns of energy availability, productivity, and biodiversity suggest that the impacts of climate change in temperate and polar regions will be shaped by the appearance of new species at least as much as by the disappearance of current species.

  2. Because climate change will produce change at multiple levels of biological organization, prediction of its impacts on mammals need to be extended beyond consideration of individual populations or species, to the functional groups and trophic categories that comprise biological communities.

  3. Bioenergetic approaches are ideally suited to predicting the impacts of climate change at multiple levels of organization because energy budgets integrate biotic and abiotic influences and translate individual animal function into population and community outcomes.

  4. Seasonality is a fundamental property of northern systems with major effects on all biotic components. A key consideration in the prediction of climate change impacts on northern mammals thus involves how mammal energetics and trophic dynamics vary with season and how seasonality will be altered by climate change.

Table 1. Species diversity of continuous and seasonal mammal herbivores in three subregions of North America

Table 2. Herbivore composition of northern mammal communities

Fig. 1. Contour map of mammalian species density (number of species per 58,275 km2 quadrat) in North America (from Badgley and Fox, 2000)

Fig. 1. Contour map of mammalian species density (number of species per 58,275 km2 quadrat) in North America (from Badgley and Fox, 2000)

Fig. 2. The relationship between contemporary mammal species richness patterns in Canada and average annual temperature (from Kerr and Packer, 1998)

Fig. 2. The relationship between contemporary mammal species richness patterns in Canada and average annual temperature (from Kerr and Packer, 1998)

Fig. 3. Contour map showing percent increases in mammal diversity predicted to occur by 2070 using a species-energy model (from Kerr and Packer, 1998)

Fig. 3. Contour map showing percent increases in mammal diversity predicted to occur by 2070 using a species-energy model (from Kerr and Packer, 1998)

Fig. 4. The effect of hibernaculum temperature on total winter energy requirements of Myotis lucifugus (dashed line), based on a winter length of 193 days. The horizontal lines indicate approximate maximum (solid line) and minimum (dotted line) hibernation fat stores (from Humphries et al., 2002b)

Fig. 4. The effect of hibernaculum temperature on total winter energy requirements of Myotis lucifugus (dashed line), based on a winter length of 193 days. The horizontal lines indicate approximate maximum (solid line) and minimum (dotted line) hibernation fat stores (from Humphries et al., 2002b)

Fig. 5. Contemporary and projected M. lucifugus range distributions in northern North America. Confirmed northern hibernacula locations are indicated with black stars, the contemporary, model-predicted distributional limit of hibernacula is indicated by the solid grey area, and the 2080 model-predicted distributional limit is indicated by the black and white banded area. The lower margin of grey and banded zones indicate the limit beyond which to the south almost all bats could successfully hibernate, and the upper margin indicates the limit beyond which to the north almost no bats could successfully hibernate (redrawn from Humphries et al., 2002b)

Fig. 5. Contemporary and projected M. lucifugus range distributions in northern North America. Confirmed northern hibernacula locations are indicated with black stars, the contemporary, model-predicted distributional limit of hibernacula is indicated by the solid grey area, and the 2080 model-predicted distributional limit is indicated by the black and white banded area. The lower margin of grey and banded zones indicate the limit beyond which to the south almost all bats could successfully hibernate, and the upper margin indicates the limit beyond which to the north almost no bats could successfully hibernate (redrawn from Humphries et al., 2002b)

Fig. 6. Biomass of vegetation, herbivores, and carnivores predicted along a gradient of environmental productivity by Oksanen's ecosystem exploitation hypothesis (Oksanen et al., 1981; Oksanen and Oksanen, 2000). Zones I, II, III represent regions expected to be characterized by one-link, two-link, and three-link trophic dynamics, respectively. If direct effects of climate on vegetation supersede direct effects on herbivores and carnivores, then responses of most northern communities to climate change should involve a shift to the right along the productivity axis

Fig. 6. Biomass of vegetation, herbivores, and carnivores predicted along a gradient of environmental productivity by Oksanen's ecosystem exploitation hypothesis (Oksanen et al., 1981; Oksanen and Oksanen, 2000). Zones I, II, III represent regions expected to be characterized by one-link, two-link, and three-link trophic dynamics, respectively. If direct effects of climate on vegetation supersede direct effects on herbivores and carnivores, then responses of most northern communities to climate change should involve a shift to the right along the productivity axis

Fig. 7. Biomass of vegetation, carnivores, continuous herbivores, and seasonal herbivores predicted along a gradient of environmental productivity. Predictions are derived from Oksanen's ecosystem exploitation hypothesis (Oksanen et al., 1981; Oksanen and Oksanen, 2000) and contemporary distributions of seasonal and continuous herbivores (Table 1). Zones I, II, III represent regions expected to be characterized by one-link, two-link, and three-link trophic dynamics, respectively. Zone III is further subdivided into regions without (IIIa) and with (IIIb) seasonal herbivores. If direct effects of climate on vegetation supersede direct effects on herbivores and carnivores, then responses of most northern communities to climate change should involve a shift to the right along the productivity axis

Fig. 7. Biomass of vegetation, carnivores, continuous herbivores, and seasonal herbivores predicted along a gradient of environmental productivity. Predictions are derived from Oksanen's ecosystem exploitation hypothesis (Oksanen et al., 1981; Oksanen and Oksanen, 2000) and contemporary distributions of seasonal and continuous herbivores (Table 1). Zones I, II, III represent regions expected to be characterized by one-link, two-link, and three-link trophic dynamics, respectively. Zone III is further subdivided into regions without (IIIa) and with (IIIb) seasonal herbivores. If direct effects of climate on vegetation supersede direct effects on herbivores and carnivores, then responses of most northern communities to climate change should involve a shift to the right along the productivity axis

1

From the Symposium Biology of the Canadian Arctic: A Crucible for Change in the 21st Century presented at the Annual Meeting of the Society for Integrative and Comparative Biology, 4–8 January 2003, at Toronto, Canada.

References

Aber
,
J.
, R. P. Neilson, S. McNulty, J. M. Lenihan, D. Bechelet, and R. J. Drapek.
2001
. Forest processes and environmental change: predicting the effects of individual and multiple stressors.
BioScience
 ,
51
735
-751.
Allen
,
A. P.
, J. H. Brown, and J. F. Gillooly.
2002
. Global biodiversity, biochemical kinetics, and the energetic-equivalence rule.
Science
 ,
297
1545
-1548.
Badgley
,
C.
, and D. L. Fox.
2000
. Ecological biogeography of North American mammals: Species density and ecological structure in relation to environmental gradients.
J. Biogeogr
 ,
27
1437
-1467.
Banfield
,
A. W. F.
1974
. The mammals of Canada. University of Toronto Press, Toronto.
Boer
,
G. J.
, G. M. Flato, and D. Ramsden.
2000
. A transient climate change simulation with greenhouse gas and aerosol forcing: Projected climate change to the 21st century.
Climate Dyn
 ,
16
427
-450.
Buck
,
C. L.
, and B. M. Barnes.
1999
. Annual cycle of body composition and hibernation in free-living arctic ground squirrels.
J. Mamm
 ,
80
430
-442.
Buck
,
C. L.
, and B. M. Barnes.
1999
. Temperatures of hibernacula and changes in body composition of arctic ground squirrels over winter.
J. Mamm
 ,
80
1264
-1276.
Caughley
,
G.
, and A. Gunn.
1993
. Dynamics of large herbivores in deserts: Kangaroos and caribou.
Oikos
 ,
67
47
-55.
Chesson
,
P.
2000
. Mechanisms of maintenance of species diversity.
Annu. Rev. Ecol. Syst
 ,
31
343
-366.
Currie
,
D. J.
1991
. Energy and large-scale patterns of animal- and plant-species richness.
Am. Nat
 ,
137
27
-49.
Currie
,
D. J.
2001
. Projected effects of climate change on patterns of vertebrate and tree species richness in the conterminous United States.
Ecosystems
 ,
4
216
-225.
Dai
,
A.
, T. M. L. Wigley, B. A. Boville, J. T. Kiehl, and L. E. Buja.
2001
. Climates of the twentieth and twenty-first centuries simulated by the NCAR climate system model.
J. Climate
 ,
14
485
-519.
Danks
,
H. V.
2004
. Seasonal adaptations in arctic insects.
Integr. Comp. Biol
 ,
44
85
-94.
Davenport
,
J.
1992
. Animal life at low temperature. Chapman and Hall, London.
Davis
,
M. B.
1986
. Climatic instability, time lags, and community disequilibrium. In J. Diamond and T. J. Case (eds.), Community ecology, pp. 269–284. Harper & Row, New York.
Derocher
,
A. E.
, N. J. Lunn, and I. Stirling.
2004
. Polar bears in a warming climate.
Integr. Comp. Biol
 ,
44
163
-176.
Dingle
,
H.
1996
. Migration: The biology of life on the move. Oxford University Press, New York.
French
,
A. R.
1992
. Mammalian dormancy. In T. E. Tomasi and T. H. Horton (eds.), Mammalian energetics, pp. 105–131. Comstock Publishing, Ithaca, N.Y.
Gaston
,
K. J.
2000
. Global patterns in biodiversity.
Nature
 ,
405
220
-227.
Gauthier
,
G.
, J. Bêty, J.-F. Giroux, and L. Rochefort.
2004
. Trophic interactions in a high Arctic snow goose colony.
Integr. Comp. Biol
 ,
44
119
-129.
Geiser
,
F.
1988
. Reduction of metabolism during hibernation and daily torpor in birds and mammals: Temperature effect or physiological inhibition?
J. Comp. Physiol
 ,
158
25
-38.
Grover
,
J. P.
1995
. Competition, herbivory, and enrichment: Nutrient-based models for edible and inedible plants.
Am. Nat
 ,
145
746
-774.
Hagmeier
,
E. M.
1966
. A numerical analysis of the distributional patterns of North American mammals. II. Re-evaluation of the provinces.
Syst. Zool
 ,
15
279
-299.
Hairston
,
N. G.
, F. E. Smith, and L. B. Slobodkin.
1960
. Community structure, population control, and competition.
Am. Nat
 ,
94
421
-425.
Hansell
,
R. I. C.
, J. R. Malcolm, H. Welch, R. L. Jefferies, and P. A. Scott.
1998
. Atmospheric change and biodiversity in the Arctic.
Environ. Monit. Assess
 ,
49
303
-325.
Heldmaier
,
G.
, and T. Ruf.
1992
. Body temperature and metabolic rate during natural hypothermia in endotherms.
J. Comp. Physiol. B
 ,
158
696
-706.
Holt
,
R. D.
, and J. H. Lawton.
1994
. The ecological consequences of shared natural enemies.
Annu. Rev. Ecol. Syst
 ,
25
495
-520.
Houghton
,
J. T.
, Y. Ding, D. J. Griggs, M. Noguer, P. J. van der Linden, X. Dai, K. Maskell, and C. A. Johnson.(eds.)
2001
. Climate change 2001. The scientific basis. Contribution of Working Group I to the Third Assessment Report of the Intergovernmental Panel on Climate Change. Cambridge University Press, Cambridge.
Huey
,
R. B.
1991
. Physiological consequences of habitat selection.
Am. Nat
 ,
137
S91
-S115.
Humphries
,
M. M.
, D. W. Thomas, C. L. Hall, J. R. Speakman, and D. L. Kramer.
2002
. The energetics of autumn mast hoarding in eastern chipmunks.
Oecologia
 ,
133
30
-37.
Humphries
,
M. M.
, D. W. Thomas, and J. R. Speakman.
2002
. Climate-mediated energetic constraints on the distribution of hibernating mammals.
Nature
 ,
418
313
-316.
Humphries
,
M. M.
, D. W. Thomas, and D. L. Kramer.
2003
. The role of energy availability in mammalian hibernation: A cost benefit approach.
Physiol. Biochem. Zool
 ,
76
165
-179.
Jefferies
,
R. L.
, R. F. Rockwell, and K. F. Abraham.
2004
. Agricultural food subsidies, migratory connectivity and large-scale disturbance in Arctic coastal systems: A case study.
Integr. Comp. Biol
 ,
44
130
-139.
Johns
,
T. C.
, R. E. Carnell, J. F. Crossley, J. M. Gregory, J. F. B. Mitchell, C. A. Senior, S. F. B. Tett, and R. A. Wood.
1997
. The second Hadley Centre coupled ocean-atmosphere GCM: Model description, spinup and validation.
Climate Dyn
 ,
13
103
-134.
Kerr
,
J.
, and L. Packer.
1998
. The impact of climate change on mammal diversity in Canada.
Environ. Monit. Assess
 ,
49
263
-270.
King
,
J. R.
, and M. E. Murphy.
1985
. Periods of nutritional stress in the annual cycles of endotherms: fact or fiction?
Amer. Zool
 ,
25
955
-964.
Krebs
,
C. J.
, S. Boutin, and R. Boonstra.
2001
. Ecosystem dynamics of the boreal forest: The Kluane project. Oxford University Press, New York.
Krebs
,
C. J.
, K. Danell, A. Angerbjörn, J. Agrell, D. Berteaux, K. A. Bråthen, Ö Danell, S. Erlinge, V. Fedorov, K. Fredga, J. Hjältén, G. Högstdedt, I. S. Jónsdóttir, A. J. Kenney, N. Kjellén, T. Nordin, H. Roininen, M. Svensson, M. Tannerfeldt, and C. Wiklund.
2003
. Terrestrial trophic dynamics in the Canadian Arctic.
Can. J. Zool
 ,
81
827
-843.
Kunz
,
T. H.
, J. A. Wrazen, and C. D. Burnett.
1998
. Changes in body mass and fat reserves in pre-hibernating little brown bats (Myotis lucifugus).
Ecoscience
 ,
5
8
-17.
Lindstedt
,
S. L.
, and M. S. Boyce.
1985
. Seasonality, fasting endurance, and body size in mammals.
Am. Nat
 ,
125
873
-878.
Ludwig
,
D.
, M. Mangel, and B. Haddad.
2001
. Ecology, conservation, and public policy.
Annu. Rev. Ecol. Syst
 ,
32
481
-517.
Lyman
,
C. P.
, J. S. Willis, A. Malan, and L. C. H. Wang.
1982
. Hibernation and torpor in mammals and birds. Academic Press, New York.
MacArthur
,
R. H.
1965
. Patterns of species diversity.
Biol. Rev. Camb. Philos. Soc
 ,
40
510
-533.
Malcolm
,
J. R.
, A. Markham, R. P. Neilson, and M. Garaci.
2002
. Estimated migration rates under scenarios of global climate change.
J. Biogeogr
 ,
29
835
-849.
McNab
,
B. K.
2001
. The physiological ecology of vertebrates: A view from energetics. Cornell University Press, Cornell.
Millar
,
J. S.
, and G. J. Hickling.
1990
. Fasting endurance and the evolution of mammalian body size.
Funct. Ecol
 ,
4
5
-12.
Moen
,
R.
, J. Pastor, and Y. Cohen.
1997
. A spatially explicit model of moose foraging and energetics.
Ecology
 ,
78
505
-521.
Namba
,
T.
1984
. Competitive coexistence in a seasonally fluctuating environment.
J. Theor. Biol
 ,
111
369
-386.
Nedergaard
,
J.
, and B. Cannon.
1990
. Mammalian hibernation.
Philos. Trans. R. Soc. London B Biol. Sci
 ,
326
669
-686.
Odum
,
E.
1963
. Ecology. Holt, Rinehart & Winston. New York.
Oksanen
,
L.
, and T. Oksanen.
2000
. The logic and realism of the hypothesis of exploitation ecosystems.
Am. Nat
 ,
155
703
-723.
Oksanen
,
L.
, S. D. Fretwell, J. Arruda, and P. Niemelä.
1981
. Exploitation ecosystems in gradients of primary productivity.
Am. Nat
 ,
118
240
-261.
Parmesan
,
C.
, and G. Yohe.
2003
. A globally coherent fingerprint of climate change impacts across natural systems.
Nature
 ,
421
37
-42.
Reid
,
D. G.
, and C. J. Krebs.
1996
. Limitations to collared lemming population growth in winter.
Can. J. Zool
 ,
74
1284
-1291.
Robbins
,
C. T.
1993
. Wildlife feeding and nutrition. Academic Press, Inc., San Diego, Calif.
Root
,
T. L.
, J. T. Price, K. R. Hall, S. H. Schneider, C. Rosenzweig, and J. A. Pounds.
2003
. Fingerprints of global warming on wild animals and plants.
Nature
 ,
421
57
-60.
Rosenzweig
,
M. L.
1995
. Species diversity in space and time. Cambridge University Press, Cambridge.
Roy
,
J.
, B. Saugier, and H. A. Mooney.(eds.)
2001
. Terrestrial global productivity. Academic Press, San Diego.
Russell
,
G. L.
, J. R. Miller, and D. Rind.
1995
. A coupled atmosphere-ocean model for transient climate change studies.
Atmos. Ocean
 ,
33
683
-730.
Scholander
,
P. F.
, V. Waters, R. Hock, and L. Irving.
1950
. Body insulation of some arctic and tropical mammals and birds.
Biol. Bull
 ,
99
225
-235.
Simpson
,
G. G.
1964
. Species density of North American Recent mammals.
Syst. Zool
 ,
13
57
-73.
Speakman
,
J. R.
2000
. The cost of living: Field metabolic rates of small mammals.
Adv. Ecol. Res
 ,
30
177
-297.
Vander Wall
,
S. B.
1990
. Food hoarding in animals. University of Chicago Press, Chicago.
Wilson
,
D. E.
, and S. Ruff.
1999
. Smithsonian book of North American mammals. Smithsonian Institute Press, Washington.
Yodzis
,
P.
, and S. Innes.
1992
. Body size and consumer-resource dynamics.
Am. Nat
 ,
139
1151
-1175.