Keywords

1 Introduction

American bison (Bison bison) are the largest extant land animal in North America and have an important history and contemporary role in modern conservation. The determined persecution of the American bison during the nineteenth century across the species’ once continental abundance and distribution, followed by a narrow escape from extinction and the loss of many native people’s lifeways and traditional homelands, is now entrenched in the narratives of the peoples and nations of North America (Aune and Plumb 2019). A diverse suite of enterprises emerged in the early twentieth century to underwrite the species’ increase across four orders of magnitude (e.g., from 100 s to 100,000 s) so that the species is once again widely distributed, albeit extremely patchily, across much of the breadth of its historical range (Plumb et al. 2014; Aune et al. 2017). Rogers (2021) estimates that approximately 350,000–400,000 bison now exist within at least 2,500 distinct herds ranging in size from 10s to 1,000s on a wide array of Tribal, national, state, province, county public, private for-profit and nonprofit lands, and accredited zoos in Mexico, Canada, and the United States (Fig. 23.1). This trajectory has now placed the American bison in an unprecedented position across the full scope of North American rangeland wildlife, with less than 2% of the total abundance functioning as rangeland wildlife. Thus, while the American bison may seem to be an iconic and ubiquitous species, worthy of their designation as the United States national mammal, recovery as a wildlife species is not yet assured. There is hope for ecological restoration of the species (Redford et al. 2016), but time is running out. This chapter focuses on recent advances in understanding of life history and synecology from across an array of herd sizes, rangelands and management approaches, and the challenges and opportunities remaining for full ecological restoration of the American bison as wildlife.

Fig. 23.1
A map of North America with the bison herds in the regions of the United States, Mexico, and Canada.

North American private and public bison herds surveyed by Rogers (2021). Historical range is based on Sanderson et al. (2008), Plumb and McMullen (2018) and updated with information from the Mexican Bison Working group (List pers. comm. 2021). Rangeland ecoregions are Environmental Protection Agency level III ecoregions

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2 Species and Population Status

2.1 Historic Range

American bison historically had the widest continental distribution of all native ungulates (Roe 1970; Gates and Ellison 2010; Plumb et al. 2014; Aune et al. 2017), and now only functions as wildlife under natural selection on < 1.2% of the original range (Sanderson et al. 2008; Aune et al. 2017). The maximum distribution was from southwest desert grasslands (~30° N × 110° W) to the floodplain meadows and northern forests of interior Alaska and Canada (~65° N × 155° W) and from western basin-range systems (~120° W) eastward across the Rocky Mountain cordillera, and then more continuous across the breadth of the Great Plains and Midwest deciduous forest savannahs, eventually dispersing across the Appalachian Mountains into the eastern coastal plains (~75° W), and from near sea level up to 4,000 m elevation (Gates and Ellison 2010; Bailey 2016; Cannon 2018) (Fig. 23.1). Martin et al (2017) and Plumb and McMullen (2018) reassessed multi-disciplinary evidence and concluded that the Colorado Plateau should be included within the southwestern periphery of the species historic range.

The plains bison (B. b. bison) proliferated in North America to an estimated 60 million individuals prior to European contact (Gates et al. 2010; Aune and Plumb 2019). Despite their prolific abundance and vast range in the early 1800s, extreme hunting pressure for profit and intentional subjugation of native peoples resulted in the total bison continental meta-population dwindling to less than 1,000 individuals by the late nineteenth Century (Shaw and Lee 1997). Wood bison (B. b. athabascae) were historically less plentiful than their southern plains relatives. Despite an expansive range from mid-Alberta into interior Alaska, initial estimates based on available suitable habitat suggest a pre-European settlement wood bison population of 168,000 individuals (Soper 1941). Further analysis of historically suitable habitat in Alaska identified a significantly larger range than previously estimated by Soper (1941), suggesting that the population size could have been larger (Stephenson et al. 2001). Wood bison were initially able to avoid extirpation due to their northern distribution (Soper 1941), but by 1840, wood bison began to suffer significant declines caused primarily by human overharvest (Gates et al. 2010). By the early twentieth century, only approximately 300 wood bison survived in the wild near the Great Slave Lakes, Northwest Territories, Canada (Fuller 2002).

There is clear evidence that within this continental historical range, the distribution and abundance of the American bison varied widely and included extensive areas with near-continual presence to areas with only intermittent low-density abundance (Isenberg 2000; Stephenson et al. 2001; List et al. 2007; Gates et al. 2010; Plumb et al. 2014; Flores 2016). Flores (1991) illustrates how extensive droughts resulted in multiple intervals spanning centuries when bison were almost totally absent from the southern Great Plains, including between BCE 5000–2500 and AD 500–1300. Given the highly variable abundance and density of bison across its historical range, it is unclear whether low-density peripheral populations were indeed less viable compared to higher density central populations in accordance with MacArthur and Wilson (1967) and Brown and Kodric-Brown (1977), or whether smaller peripheral populations contributed to species viability via local adaptation (Nielsen et al. 2001; Eckert et al. 2008); and whether the edge of historical range was simply an “expansion threshold” wherein peripherally dispersed small populations persisted through local adaptation or failed either through catastrophic mortality events or when genetic drift reduced genetic diversity below that required for adaptation to a heterogeneous environment (Polechová 2018). In summary, there is strong evidence that extensive variability in temporal and spatial patterns of abundance and distribution occurred throughout the historical range (Gates et al. 2010; Plumb et al. 2014; Aune et al. 2017; Plumb and McMullen 2018), and that these variable patterns should be critically considered anywhere wild bison conservation now occurs, especially at the range periphery (Gates et al. 2010; Plumb and McMullen 2018).

2.2 The Continuum

The resurgence of American bison across the historical range in the past 120 years has occurred through a continuum of diverse sectors of legal status and purposes. Bison as rangeland wildlife occur on an array of exclusive and overlapping governance jurisdictions (e.g., Federal, State, Provincial, County, and Tribes and First Nations). Privately-owned bison have also become core assets of private not-for-profit conservation lands enterprises, zoo and education enterprises. Most bison are now living as privately-held for-profit assets on private lands. The historical and prevailing relationships within and between these higher order sectors are very complex and often conflicting, yet each sector has invested tremendous effort and public and private resources to increase the total abundance of bison to present levels. While there is not yet a comprehensive, rigorous, and comparable accounting or long-term monitoring of the total abundance of the American bison across all sectors and jurisdictions, there are several reliable assessments within key sectors. What we do know is that there is now a continuum of wildness from free-ranging herds, under an array of strong natural selection forces, to captive herds exclusively under an array of non-natural selection forces. These two divergent conditions bracket a continuum of diminishing wildness within bison management from wild to the edge of domestication (Gates 2014; Plumb et al. 2014; Aune et al. 2017). The emergent disparity in abundance over the past 50 years between wild bison (10,000s) to not-wild bison (100,000s) is staggering. Indeed, despite an increase in the number of fenced, conservation-focused herds in the past 50 years, the abundance of free-ranging wild bison has remained relatively constant.

Through the Bison Specialist Group (BSG) of the International Union for Conservation of Nature (IUCN), Gates et al. (2010) completed a comprehensive continental species assessment and conservation guidelines, and Aune et al. (2017) completed an updated Red List Assessment for the species that considers the likelihood of species extinction through accounting of all wild bison within historical range. The United States Department of Interior (DOI) maintains records of bison on lands managed by the National Park Service, U.S. Fish and Wildlife Service, and Bureau of Land Management (see Department of the Interior 2014) and recently completed a comprehensive meta-population viability assessment of all bison herds on DOI lands (Hartway et al. 2020). The Committee on the Status of Endangered Wildlife in Canada published a comprehensive assessment and status report on the plains bison in Canada (COSEWIC 2004, 201). The Mexico National Commission of Natural Protected Areas (CONANP) works with partners to maintain information on public and private bison throughout Mexico. Individual state or provincial wildlife management agencies monitor and maintain information on bison managed as wildlife under their jurisdiction. The National Bison Association (United States) primarily represents private producers and maintains updated information on bison ranching in North America (National Bison Association 2021), along with their northern counterpart, the Canadian Bison Association, and a variety of regional/state/provincial associations. The United States Department of Agriculture and Statistics Canada (the national statistical office) track their respective private bison sectors; the Inter-Tribal Buffalo Council maintains records about how many bison are living on member tribal lands, and the American Zoo Association maintains information on bison in member zoos.

2.3 Publicly-Owned Wildlife Conservation Status

In the most recent IUCN Red List Assessment, Aune et al. (2017) reported an approximate total of 31,000 bison in 68 conservation herds that are managed in the public interest by federal, state, and provincial governments and non-profit environmental organizations across North America, including 20,000 Plains Bison and 11,000 Wood Bison. A Red List Assessment focuses on a species in the wild under natural selection forces, and accordingly Aune et al. (2017) denoted three categories of conservation bison: (1) functioning as wild, (2) functioning as wild with limitations, and (3) not functioning as wild. Aune et al. (2017) classified the species in the wild as “Near-Threatened” and nearly qualifying as “Vulnerable” because the wild species is entirely conservation dependent, e.g., while the wild species is not currently in decline, the number of wild mature individuals could be greatly reduced if current management regimes are changed or removed. Nearly half (30 of 68) of the conservation bison herds, totaling ~ 2,700 bison combined, were denoted as not functioning as wild due to very small (< 300) population size on small, fenced landscapes (< 10,000 acres) for education, public viewing, and research. Another 18 herds, totaling ~ 9,500 bison combined, were denoted as wild with limitation because they are intensely managed behind fences and culled by artificial selection. Thus, only ~ 18,800 bison in 20 herds were denoted as functioning as wild under a range of natural selection forces (Fig. 23.2), and 4000 of these bison live in 12 herds containing < 400 total individuals, the lowest estimate of a minimum viable population size (MVP; Gross and Wang 2005), though it is likely closer to 1,000 (Hedrick 2009; Gates et al. 2010). As juveniles account for ~ 35% of a wild herd, Aune et al. (2017) estimated the total mature wild bison population in North America to be only ~ 12,000 animals. When we focus further on bison in rangeland habitats, that number is even smaller, with only 6 free-ranging wild herds totaling approximately 4000 adults that are subject to natural selection pressures. These 6 herds all occur outside of the historical bison strongholds of the central and northern Great Plains (Fig. 23.2).

Fig. 23.2
A map of North America with wild bison herds on the western coast and southern parts of North America.

Wild North American bison herds according to Aune et al. (2017). Historical range is based on Sanderson et al. (2008), Plumb and McMullen (2018) and updated with information from the Mexican Bison Working group (List pers. comm. 2021). Wild herds are free-ranging, managed as wildlife under natural selection, on large landscapes (> 10,000 acres), and greater than 400 individuals. Wild with limitations herds are limited in population (< 400), have limited predation from large carnivores, and/or are a subspecies outside of the historical range, but function otherwise similarly to wild herds. Rangeland ecoregions are Environmental Protection Agency level III ecoregions

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2.4 Tribes and First Nations

The InterTribal Buffalo Council (ITBC) was convened in 1991 to restore the American bison to Indian Country, and now includes 69 federally recognized Tribes from 19 states with ~ 20,000 Plains Bison on Tribal lands in the United States. The ITBC vision is that reestablishing healthy buffalo populations on Tribal lands will reestablish hope for Indian people; and that returning bison to Tribal lands will help heal the land, the animal, and the spirit of the Indian people. In 2014, dignitaries from U.S. Tribes and Canadian First Nations signed the “Northern Tribes Buffalo Treaty” to establish an inter-tribal alliance cooperating to restore bison on Tribal/First Nations Reserves or co-managed lands within the U.S. and Canada. Collectively these treaty Tribes own and manage ~ 6.3 million acres of grassland and prairie habitats in the United States and Canada and have articulated a goal to achieve ecological restoration of the bison, and, in so doing, reaffirm and strengthen ties that formed the basis for traditions thousands of years old, including youth education and cultural restoration among the tribes.

2.5 Private Lands Commodity

There are now roughly 300,000 American bison under private ownership under agriculture laws and policies of Canada, United States, and Mexico (Statistics Canada 2016; United States Department of Agriculture 2017; National Bison Association 2021), with an unknown relatively small number in private ownership in Europe and Australia (Rogers 2021). Combined, the private bison producer sector slaughters ~ 70,000 bison annually, compared to the annual North American beef slaughter of ~ 45 million cattle (National Bison Association 2021). The National Bison Association (NBA) has initiated a web-based Conservation Management Program for bison farmers and ranchers to monitor conservation practices to improve their overall stewardship outcomes for bison, the land, and surrounding communities, though it is not required of all members. NBA also has announced a “Million Bison” marketing campaign to triple the number of bison across all sectors, with the private sector playing a major role (National Bison Association 2021).

3 Population Monitoring

Population abundance and demography monitoring occurs at the individual herd level across all proprietorship sectors. Most privately-owned bison herds are monitored through direct observation in the field and during annual round-ups, in which vaccinations, disease tests, and pregnancy tests are often administered to assess individual condition and herd health (Rogers 2021). During these round-ups, individuals are often separated for transfer between herds, sale for slaughter, or culling to manage population abundance. Free-ranging herds are monitored through traditional wildlife survey techniques, including ground and aerial surveys. Aerial surveys generally incorporate sightability indices based upon mark-recapture estimates derived from radio-collared animals (see Hess 2002). Distinct age- and sex-specific traits readily allow direct ground sex classification of individual adults and juveniles (Gates et al. 2010). Demographic classification via aerial surveys is possible by visual observation, though use of stabilized digital photography yields higher resolution information that can be assessed following the aerial survey. The free-ranging Henry Mountains bison herd in Utah, for example, does not have an annual round up, and relies purely on ground and aerial surveys (Bates and Hersey 2016; Terletzky and Koons 2016), along with hunter harvest data managed by the Utah Division of Wildlife Resources. Yellowstone National Park has a systemic approach to its annual population surveys, conducting intensive aerial surveys at the beginning of the summer and winter seasons, followed by duplicate aerial surveys, and ground classification surveys to refine estimates of population abundance and demography (Hess 2002).

At the species level, assessments for genetic integrity and extinction risk are completed at semi-regular intervals by government and non-government agencies. Recently the DOI conducted a population viability analysis for all bison herds on federal lands to assess their long-term ability to persist into the future (Hartway et al. 2020). The IUCN BSG conducts Red List Assessments for the American bison based upon data contributed and collated across all conservation herds (see Aune et al. (2017) and plans to publish the next assessment in 2024 (Greg Wilson, personal communication). In addition to Red List Assessments, the IUCN has recently developed the Green Status Assessment, which is a tool that assesses the ecological recovery legacy (e.g., late 1800s–2021) and potential for additional recovery a species has at specific future time intervals (Akçakaya et al. 2018; Grace et al. 2021). The first Green Status Assessment for bison was completed in 2022 (Rogers et al. 2022). 

4 Life History and Population Dynamics

4.1 Description

A compact, large body and a large head set on a strong neck, combined with a pronounced hump and horns curving inwards give the American bison a widely recognized iconic appearance (Fig. 23.3). Weighing up to 1000 kg and with body length of 2.1–3.5 m, and shoulder height 1.5–2 m, it is the largest terrestrial mammal of the Western Hemisphere (Nowak 1991; Shaw and Meagher 2000; Reynolds et al. 2003). Sexual dimorphism occurs among adults with males ~ 20–30% larger, yet females resemble males in color, general body configuration, and presence of permanent horns that are short and sharp from the side of the head that laterally curve upward over the head, with female horns slenderer and showing a greater tendency to curve inward toward the tips (Reynolds et al. 1982, 2003; Nowak 1991; Shaw and Meagher 2000). Allen (1876) first described the species with a narrow muzzle with long pointed nasal bones composed of premaxillae, maxillae, and nasals, with tubular orbits composed of frontals, lacrimals and jugals without preorbital vacuities in the skull. Pelage of the head, neck, shoulders, and front legs is brownish-black, long, and shaggy; while the rest of the body is covered with shorter brown hairs that lighten through sun bleaching (Nowak 1991; Shaw and Meagher 2000). Albino or white/gray pelage is rare but known to occur (Meagher 1973). The bison has a distinctive tufted tail and chin hair that usually resembles a goatee-type beard (Banfield 1974). Young calves are orange-brown to reddish-brown “buff” color that gradually darkens to adult coloration by 4 months (Fig. 23.4; Nowak 1991; Shaw and Meagher 2000). Males present a distinctive shoulder hump suggesting forequarters out of proportion to smaller appearing hindquarters (Reynolds et al. 1982). Bison produce a variety of sounds, including a male “bellow” heard most frequently during the breeding season, and a “snort” and “cough” associated with antagonistic behavior. Cows searching for calves will exhibit a series of snorts, and calves can exhibit bawling (Fuller 1960).

Fig. 23.3
A photograph of bison with shaggy long fur on the front part of the body, a round hump, and 2 horns.

Young adult male bison in the Henry Mountains of southern Utah (photo: Dustin H. Ranglack)

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Fig. 23.4
A photograph of a bison calf with little fur all over the body.

Bison calf in Yellowstone National Park, Wyoming (photo: Dustin H. Ranglack)

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4.2 Growth

Typically, a single calf is born between 15 and 30 kg and begins grazing and drinking water within a week, while continuing to nurse for approximately 7–12 months (McHugh 1958; Fuller 1960; Halloran 1961; Meagher 1986). Calves double their body mass by 3 months of age, and weigh between 135 and 180 kg by 8–9 months of age (McHugh 1958; Halloran 1961; Meagher 1973, 1986; Gogan et al. 2010). The general age of bison may be determined in the field by body size, and horn size and shape. For both sexes, there is strong growth of body mass and inward-curving horns by 4 years of age, with female bison fully grown by 4 years, and males fully grown by 6 years (Banfield 1974). Distinct sex differences in horn size and shape occur earlier than differences in body size and mass. Male horns grow continuously until full development by 7–8 years of age, with horn tips then frequently becoming worn down and rounded, due to rubbing against trees and aggression with other males, and horn bases larger in diameter than their eyes. Female horns grow longer and more curved inward with age, with 20+ year old females retaining sharp horn tips which are generally smaller in diameter.

4.3 Reproduction

Like the European bison (Bison bonasus), the American bison exhibits a polygynous tending-bond mating system wherein non-territorial males court individual oestrous females for up to 3–4 days during an annual rut concentrated in August–September, with July seeing initial increased time spent by mature males beginning to search for receptive females (Plumb et al. 2014). The annual rut often starts later at higher latitudes (Fuller 1962). In herds with an even sex ratio, males between 3 and 6 years old are capable of breeding, but generally are prevented by older, larger, and more experienced males (McHugh 1958; Halloran 1961; Lott 1981; King et al. 2019). Females reach sexual maturity between 2 and 4 years old with a first calf often at 3 years of age (McHugh 1958; Halloran 1961; Fuller 1962; Meagher 1973; Gogan et al. 2010). Estrous lasts 19–26 days with females being receptive for 1–2 days, with only 1–2 ovulations during an annual breeding season (Haugen 1974; Rutberg 1986; Kirkpatrick et al. 1991). Females are often fertile up to 16 years of age (Green 1990) and produce young every 1–3 years depending on their age and physical condition (Kirkpatrick et al. 1993). Gestation is 285 days and fetal sex ratios are often male-biased (Fuller 1960; Meagher 1973; Haugen 1974; Rutberg 1986). Birth synchrony is common with 80% of births occurring during April–June (Haugen 1974; Rutberg 1984; Jones et al. 2010), with synchrony especially noticeable in populations where predation on calves occurs (Gates and Larter 1990). Earlier onset of birth synchrony has been observed in landscapes with earlier onset of spring vegetation growth, and it is thought this adaptation yields increased lactation quality and neonate survival based on higher vegetation quality (Gogan et al. 2005). Conversely, females in poor nutritional condition, with debilitating diseases, or of lower rank may calve later and/or show low synchrony in birthing (Berger 1992; Green and Rothstein 1993a; Berger and Cain 1999; King et al. 2019).

Green and Rothstein (1993b) observed that females born early are more fecund throughout their life compared to calves born towards the end of birth synchrony. Most mature males spend the balance of the year solitary or in small bachelor groups, only joining mixed age-sex herds during the annual rut (Meagher 1973). Competitive mate selection is driven by male competitive dominance through threat displays and short-duration violent pair-wise matches (McHugh 1958; Lott 1981). Females often segregate themselves from the herd or group prior to parturition while lying down, followed by freeing the calf by consumption of the placental membrane and licking amniotic fluid from the calf’s fur. Suckling often initiates within 10 min, and newborn calves can stand and continue to nurse within 30 min of birth (McHugh 1958; Meagher 1986; Lott 2002).

5 Population Dynamics

American bison are generally long-lived, with females occasionally reaching 25 years age and males rarely exceeding 20 years (Gates et al. 2010; Aune et al. 2017; Hartway et al. 2020). Overall adult annual survival rate generally approaches 90% in fenced-protected herds (Gates et al. 2010). In free-ranging populations below carrying capacity, the annual adult survival rate is variable across the continental distribution, wherein an adult survival rate of 75% was observed for wood bison at the Mackenzie Bison Sanctuary, and 95% for plains bison at the Jackson Hole National Wildlife Refuge and Grand Teton National Park (Larter et al. 2000; US Fish and Wildlife Service and National Park Service 2007). Age-specific annual survival rate slightly favors females over males and declines linearly for both sexes from > 75% up to 3 years age to ~ 50% by 12–15 years age, with a subsequent sharply punctuated decline to < 5% between 16 and 20 years age (Hartway et al. 2020). Individual bison infected with brucellosis and tuberculosis at Wood Buffalo National Park exhibited lower age- and sex-specific survival rate than bison with only one of the two diseases, or not infected at all (Joly and Messier 1999, 2004a, b, 2005; Bradley and Wilmshurst 2005).

Annual population growth rate for free-ranging populations also is variable across the continental distribution, from r = 0.08 with an 8-year generation length for relatively wild conditions with predation pressure; to r = 0.15–0.19 with a 9–10 year generation length for herds without dramatic environmental stochasticity and little predation pressure (Hartway et al. 2020). Under stochastic environmental conditions (e.g., drought, wildland fire, winter severity) or reduced genetic variability, growth rates will be lower, especially for smaller herds (Turner et al. 1994; Green et al. 1997; Wallace et al. 2004; Geremia et al. 2008; Hartway et al. 2020). Large mortality events are known to have occurred historically due to wildland fire (see Haley 1936; Hart and Hart 1997) and occasionally still occur at higher latitudes when bison drown after falling through thin ice in spring and fall (Roe 1970; Gates et al. 1991; Mech et al. 1995). Otherwise, bison are capable of swimming short and long distances. Larter et al. (2003) observed bison swimming across a 1.7 km-wide section of the Liard River, taking 27 min with downstream movement to swim a total of 3.6 km.

6 Habitat Associations

As their name suggests, the plains bison is the sub-species that historically occurred across a diverse array of North American rangeland ecoregions (Fig. 23.1). While the plains bison was a dominant keystone species amongst the rangeland ecoregions of the Great Plains (Knapp et al. 1999), the species was also widely distributed across non-rangeland ecoregions throughout much of the North America (Gates et al. 2010; Plumb et al. 2014; Aune et al. 2017). Not-wild plains bison now occur in many diverse rangeland ecoregions, albeit at severely fragmented and significantly reduced spatial scales behind fences subject to agricultural laws and policies (Fig. 23.1).

Free-ranging wild bison are very limited in abundance and distribution throughout their historical range (Rogers 2021), with the IUCN Green Status designating bison as Critically Depleted (Rogers et al. 2022). Out of the 21 free-ranging wild bison herds considered in the most recent IUCN Red List Assessment (Aune et al. 2017), only six were present within western rangelands: the Northern Mixed Grass Prairie, Rocky Mountain, and Colorado Plateau rangeland ecoregions (Fig. 23.2). The other 15 free-ranging wild bison herds outside of western rangelands are present in ecoregions consisting predominantly of aspen parkland, boreal forests, and wetlands in Canada and Alaska (Aune et al. 2017). Concomitant with limited abundance and distribution, the few free-ranging populations are managed to restrict large scale dispersal, range expansion or migratory movement patterns beyond their designated reserve landscapes (Plumb et al. 2014; Aune et al. 2017). Restricting the larger scales of bison habitat associations may thus restrict the fundamental ecological functionality of bison across rangeland types, often yielding continuous smaller-scale grazing patterns inside restricted landscapes rather than large-scale high-intensity and short-duration grazing events (i.e., with potential for prolonged periods of absence) (Gates et al. 2010). Indeed, it has been suggested that American bison are ‘terrestrial castaways’, stranded on ‘island’ ranges within a matrix of inaccessible habitat (Ritson 2019). As such, Augustine et al. (2019) suggest increasing spatial scales for rangeland wildlife through novel partnerships for cross-jurisdiction management that could then support large-scale bison movements.

6.1 Bison Diet

Diet varies by rangeland ecoregion, climate regime and time of year, with overall average peak dietary quality at the height of summer in June (Bermann et al. 2015; Craine 2021). Cool, wet climates produce the highest amount of crude protein and digestible plant organic matter, resulting in larger average body mass compared to hot and dry climates (Craine 2021). Plains bison are predominantly grazers and exhibit optimal foraging ecology in response to dynamic temporal and spatial patterns of graminoid forage availability and quality (Plumb and Dodd 1993; Knapp et al. 1999), typified by shifts from cool season (C3) graminoids during spring to warm season graminoids (C4) during peak summer primary production, and back to cool season graminoids depending on late-summer and early-fall precipitation. Consumption of herbaceous forbs, legumes, and woody half-shrubs is exhibited but is generally indicative of non-selective foraging in relation to total availability (Plumb and Dodd 1993; Knapp et al. 1999), and more common in the spring and fall (Begmann et al. 2015).

6.2 Ecosystem Influences

The primary ecosystem function of plains bison on rangelands is contributing to plant community heterogeneity through patchily distributed short-duration, high-intensity grazing events that create mosaics of grazing pressure (Jonas and Joern 2007; Gates et al. 2010; Tastad 2014). At larger spatial scales, plains bison can contribute to enhanced total primary productivity by stimulating compensatory vegetative growth characterized by seasonal grazing lawns (Coppock et al. 1983; McNaughton 1984; Ranglack and du Toit 2015b; Merkle et al. 2016; Geremia et al. 2019). Geremia et al. (2019) demonstrated how plains bison on montane grasslands along a strong elevational gradient at Yellowstone National Park not only respond to the onset of spring phenology at lower elevations and continue to “surf” the leading edge of high-quality forage “green wave;” but also create large scale grazing lawns along the elevational gradient that optimizes foraging efficiency and quality. When in high enough abundance, short duration intensive bison grazing stimulates plant material regrowth and delays maturation, allowing bison to continue to consume high-quality plant protein even after they fall behind the leading edge of vegetation green-up (Merkle et al. 2016; Geremia et al. 2019).

Beyond their primary ecosystem function, bison exhibit a myriad of other roles in their environment through direct and indirect interactions. Where it is allowed to occur, the decomposition of bison carcasses at the site of mortality produce biochemical hotspots. Rich in calcium and with elevated pH levels, these biochemical hotspots promote plant growth on the landscape for two to four years after mortality (Knapp et al. 1999; Towne 2000; Melis et al. 2007; Bump 2008; Bump et al. 2009). Additionally, these biochemical hotspots can facilitate heterogeneity and reduce forest expansion, as observed by reduction in aspen expansion on fescue grasslands in Riding Mountain National Park of Canada (Knapp et al. 1999; Bump 2008). Bison can also contribute to dispersal of forb and graminoid seeds through shed hair and feces (Dinerstein 1989; Rosas et al. 2008). Both subspecies of bison can create significant abundance of small-scale disturbance within small and large landscapes through wallows, trampling, horning, and grazing (Coppedge and Shaw 1997; Fox et al. 2012). Wallow pits, when filled with water from rain or flooding events, have been identified as essential breeding sites and aquatic habitat for anurans and invertebrates (Gerlanc and Kaufman 2003). Obligate shortgrass prairie bird species’ populations respond positively following bison restoration (Wilkins et al. 2019). Prairie dogs (Cynomys spp.) have a mutually beneficial relationship with bison; bison prefer to occupy sites near prairie dog towns, which provide higher crude protein and nitrogen content, and bison facilitate high plant productivity through fecal deposit, improving later foraging for prairie dogs (Coppock et al. 1983; Krueger 1986; Cid et al. 1991). Bison presence may be critical to restoring and sustaining some unique habitats such as the Canadian Sandhills, which supports several endangered and threatened species (Fox et al. 2012). In the Great Plains, grasshopper and other herbivorous insect species richness is directly and positively related to bison grazing pressure (Joern 2005). Bison landscape disturbance also has a negative impact on woody vegetation and positive impact facilitating the growth of grasses, sedges, and other graminoids in a prairie landscape (Coppedge et al. 1998). Bison fur is utilized by many species including red squirrels (Tamiasciurus hudsonicus) to insulate nests and burrows (Jung et al. 2010). The breadth and importance of roles bison exhibit within their habitat have led some to classify the plains bison as a keystone species in prairie ecosystems (Knapp et al. 1999; Fuhlendorf et al. 2010).

7 Genetics

Perhaps more than with other rangeland wildlife species, genetics play an outsized role in current bison population management. Two major forces have shaped the genetic make-up of all modern-day bison, both being human caused and irreversible. The first of these, and largest in terms of impact, is intentional hybridization of bison and cattle, as early as the late 1800s, that created viable offspring, which were then backcrossed into “pure” bison populations, leaving a legacy of cattle genetics in the resultant bison genome. This has long been of concern in the mitochondrial genome, as female plains bison with cattle mitochondrial introgression may have some indeterminant degree of reduced fitness (Derr et al. 2012). Nuclear cattle genes are also known to have also been perpetuated in the bison genome, and using a panel of 15 microsatellite loci, Halbert and Derr (2008) and Ranglack et al. (2015a) reported that cattle nuclear or mitochondrial introgression had been detected in all U.S. plains bison conservations herds except Yellowstone National Park, Wind Cave National Park, and the Henry Mountain bison herd in Southern Utah. In Canada, plains bison in Elk Island National Park are also considered to be free of cattle introgression (COSEWIC 2013). However, the technological scope of microsatellite testing also limits its inferential power, in that bison have 29 autosomes, and a panel of 15 microsatellite loci leaves entire chromosomes overlooked.

New research, using single nucleotide polymorphism (SNP) tests that examine the entire bison genome, has dramatically restructured our understanding of the consequential scope of historical hybridization, and now confirms that all contemporary plains bison herds have encountered some downstream cattle introgression in the nuclear genome (Stroup et al. 2022); including Wind Cave and Yellowstone National Parks (and therefore also the Henry Mountains herd that was established using individuals from Yellowstone in the early 1940s). Evidence now strongly indicates that 3 male bison that were introduced into the Yellowstone National Park bison herd in 1902 from the Goodnight herd (Texas), as part of an effort to boost the dwindling population in the park, were in fact hybrids that contributed to some unknown scope of breeding and explains the lack of mitochondrial introgression found in the Yellowstone population. As such, evidence now indicates that all modern plains bison herds possess variable levels of artifactual cattle introgression event(s) that occurred over 100 years ago (Stroup et al. 2022). Alternatively, Wang et al. (2018) conclude that introgression in the European bison may represent natural evolutionary legacies, whereas introgression represents incomplete lineage sorting of shared common ancestry with cattle. Still, such introgressed genetic artifacts, from the time of near-extinction of the wild American bison, do not now appear to restrain free-ranging wild bison reproductive processes and phenotypic expression (Dratch and Gogan 2010). Contemporary wild free-ranging bison with relatively higher levels of introgression, such as the Northern Rim population in northern Arizona (Hedrick 2010), exhibit no observable phenotypic traits of cattle (pelage, body conformation, etc.); they look and behave like bison, and they produce viable male and female offspring, indicating essential functionality of chromosomal DNA (Plumb and McMullen 2018).

As we look forward, the interplay between population size, isolation, and genetics is more important than the residual artifacts of a history that cannot be rewritten. A vibrant life history that is rooted in competitive mate selection and natural selection pressures remains the key to avoiding breeding dominance by individual males leading to reduced genetic diversity (Gates et al. 2010).

7.1 Genetic Bottlenecks

The second major impact humans have had on bison is the result of the bottleneck following the near extinction of the species by the end of the nineteenth century, and management practices related to their subsequent recovery (Pertoldi et al 2010). Historical records and genetic analyses indicate that all plains bison today may be descended from only 30 to 50 individuals from 6 captive herds (Hedrick 2009), and the estimated 25 wild bison that remained in Yellowstone National Park (Meagher 1973). Given that these captive herds were typically privately owned, a considerable exchange of animals between herds took place, leading to a relative homogenization of bison genetically. While in some regards this genetic exchange may have enhanced genetic diversity within those source herds, there is no doubt that there was an enormous loss of genetic diversity during this time, and that there were distinct historical lineages of bison that are now extinct (Stroup et al. 2022). While bison have recovered numerically, subsequent bottlenecks resulting from founding new herds with small numbers of individuals (Halbert and Derr 2008) have led to significant concern for the genetic health and long-term survival of certain bison populations (Hedrick 2009; Dratch and Gogan 2010; Hartway et al. 2020). This is exacerbated by the fact that most bison herds have been restricted in size and geographic range, with limited genetic exchange taking place between conservation herds (though exchange of individuals within the private sector is common). There is specific concern over the loss of genetic diversity due to genetic drift, which is the random loss of genetic material from generation to generation (Allendorf et al. 2013). This loss of genetic diversity can have very specific short- and long-term impacts on bison population viability, due to increased risk of inbreeding depression, which was documented in the decline of the Texas state bison herd (Halbert et al. 2004, 2005).

A recent population viability analysis conducted on 12 bison herds by DOI and 2 bison herds managed by Parks Canada revealed that all herds are predicted to lose genetic diversity (measured as both heterozygosity and allelic diversity) over the next 200 years due to genetic drift (Hartway et al. 2020), with smaller herds losing diversity faster than larger herds, as would be expected. This is concerning, as decreases in genetic diversity could also decrease the ability of herds to adapt to changing environmental conditions (Weeks et al. 2011; Ralls et al. 2018). Given that all these herds are subject to management removals to control the herd size due to social, political, or biological constraints, the management strategy directing removals were shown to be particularly important for reducing the amount of diversity loss, with strategies that target younger animals or using mean kinship for removal resulting in lower levels of heterozygosity loss (Hartway et al. 2020). Herd size, however, is the most important driver of genetic diversity loss, with larger herds losing diversity at a lower rate (Gross and Wang 2005; Hartway et al. 2020), however, this is often limited by other constraints, biological or otherwise, highlighting the need for a more comprehensive metapopulation management strategy that re-establishes gene flow between populations, through natural movements and/or translocation of individuals.

7.2 Genetic Augmentation

Genetic augmentation (Frankham et al. 2017) has been shown to be a successful strategy in reversing the effects of inbreeding depression in a variety of species (Bouzat et al. 2009; Johnson et al. 2010). This strategy of increasing gene flow to isolated populations is valuable (Whiteley et al. 2015; Frankham et al. 2017), but not without risk. There must be a balance both in the number individuals and frequency of translocations to ensure that there is enough genetic material transferred as to increase the genetic diversity in the recipient herd (Frankham et al. 2017), but not so much as to swamp out local adaptation or rare alleles (Edmands 2006; Allendorf et al. 2013).

Hartway et al. (2020) evaluated five different scenarios for bison metapopulation management, varying which herds were used as sources and recipients in each scenario. They suggested that smaller, less frequent translocations (i.e., 2 individuals every 10 years, 3 individuals every 7 years) using either the least-related herds as source populations, or alternating which herds are used as the source herd at each translocation event is adequate for increasing genetic diversity in most herds, while minimizing the loss of diversity at the metapopulation level. Smaller, less genetically diverse herds may benefit from more frequent translocations (Hartway et al. 2020). It is now clear that a metapopulation management strategy to re-connect isolated herds through natural or human-facilitated movement of individuals is required for the long-term conservation of bison. Moving forward, bison conservation genetics need to (1) maintain stable population size and avoid large fluctuations in abundance, (2) encourage competitive mate selection by maintaining adult breeding males approaching a 1:1 sex ratio, and (3) mitigate genetic drift by periodically augmenting isolated herds with additional animals as a part of a larger metapopulation (Dratch and Gogan 2010; Hartway et al. 2020).

8 Rangeland Management

Along the continuum of bison management there exists a wide variety of rangeland management techniques, from intensive management of bison movements through the use of adaptive multi-paddock grazing (Hillenbrand et al. 2019) to year-long continuous grazing, with every variation in between. However, as a wildlife species, the rangeland management practices associated with bison have generally focused on disturbance ecology, with an emphasis on bison’s relationship with fire (Fuhlendorf et al. 2009), with a more recent push to understand the impacts of bison grazing at scale (Augustine et al. 2019; Geremia et al. 2019). This question of scale is particularly important given that every bison is found behind a barrier—be it biological, social, physical, political, or otherwise—the movements of bison on the modern landscape are greatly restricted, thus restricting their impacts on rangeland ecology and processes.

8.1 Bison and Fire

The relationship between bison and fire is well documented for plains ecosystems (Fuhlendorf et al. 2009), with bison exhibiting a strong preference for recently burned areas, attracted by the high green vegetation: senescent vegetation ratios and high-quality forage that emerges due to nutrient release follow fires (Allred et al. 2011). Before European settlement, wildfire would have been common on the Great Plains, but in other parts of the historic bison range, the fire return interval would have been longer and more sporadic, with fire return interval estimates ranging from every 8 years to fire being rare (Anderson 2002). These periodic fires would have prevented shrub and conifer encroachment into open habitat types and maintained piñon–juniper (Pinus edulis–Juniperus osteosperma), ponderosa pine (Pinus ponderosa), and other woodlands in a more savanna-like state (West 1984), except in steep and rocky areas that were unlikely to provide significant foraging opportunities for bison. Prescribed burning is commonly used to replicate natural fire return intervals in rangelands used by free-ranging bison, as well as conservation herds on conservation lands behind fences (Plumb et al. 2014). Ranglack and du Toit (2015b) showed that forage quality, as indexed by fecal N, was higher, and that bison spent more time feeding relative to moving in previously burned areas, while mechanically treated areas were more like naturally occurring open habitat types. This response was detected even 10 years after a fire, indicating that bison were likely creating grazing lawns in these areas, thus maintaining the forage in a high-quality state (Ranglack and du Toit 2015b). Indeed, bison not only track plant phenology and respond to differences in forage quality at the landscape scale, but also modify and engineer plant phenological responses through their grazing (Geremia et al. 2019).

8.2 Bison and Cattle

Bison are largely absent from traditional range management systems, as bison and cattle are considered by many to be potential competitors, due to large overlaps in diet and body size (Van Vuren and Bray 1983; Plumb and Dodd 1993), and much research has focused on the ecological equivalence of the two species (Steuter and Hidinger 1999; Fuhlendorf et al. 2010; Allred et al. 2011, 2013; Kohl et al. 2013), given how cattle have largely replaced bison in rangeland systems. Bison dietary overlap with cattle (Vuren and Bray 1983; Plumb and Dodd 1993), combined with the conspicuous nature of bison, underpins continuing concerns over whether the two species should be allowed to share common rangelands (Ranglack et al. 2015b). In the Henry Mountains of Utah, one of the few places where bison and cattle co-mingle on shared rangeland, lagomorphs were found to have twice the impact on forage reduction than bison (Ranglack et al. 2015b), thus presenting a far greater competitive threat to cattle than bison. Indeed, bison grazing caused no significant impact on plant species composition (Ware et al. 2014) and relatively small reductions in forage availability (Ranglack et al. 2015b). This is likely because bison and cattle tend to spatially segregate on shared rangelands, as bison are more likely to range widely across the landscape, using steep slopes and venturing farther from water sources, while cattle focus grazing efforts in areas near water and on more flat terrain (Van Vuren 2001; Allred et al. 2011; Ranglack 2014). Thus, while there is potential for exclusionary competition, at this time there is little evidence that it occurs on shared rangelands, even in areas along the edges of the species range where resources are most limiting, and you would therefore expect competition to be most severe (Ranglack 2014; Ranglack et al. 2015b). While this is still not without controversy, bison and cattle are not incompatible when properly managed, and there may be opportunities to integrate the ranching and wildlife sectors on western public lands through managed bison populations (Ranglack and du Toit 2015a).

8.3 Spatial Scale, Distribution, and Abundance

One of the largest management concerns with bison on modern rangelands is the manipulation and maintenance of appropriate bison distribution and abundances. Understanding the ecological processes, impacts, and interactions (movement, activity and behavior, habitat interactions, population demography, and gene flow/introgression) of bison across time and space is critical to our evaluation of bison as a keystone species and the management of the species. While some of these processes take place over short time scales (foraging decisions; Gogan et al. 2010; Plumb and Dodd 1993) and/or small spatial extents (movements within or between feeding patches; Meagher 1989; Plumb and Dodd 1993), many require large spatial scales (migration, dispersal, range expansion; Gates et al. 2005; Plumb et al. 2009) and/or only occur at decadal time scales or longer (between population gene flow; Dratch and Gogan 2010; Halbert and Derr 2008). Thus, managers need to understand the spatiotemporal scale of the various bison ecological processes to determine what aspects of bison ecology are being conserved at the available spatial scale, and where management needs to allow ecological processes to operate are larger scales than currently available (Augustine et al. 2019).

In wild bison, spatiotemporal variability in overall habitat quality influences many aspects of the behavior of animals in groups, such as group size, composition, and behavior within groups including where, when, and for how long group members forage (Lima and Zollner 1996; WallisDeVries 1996). Optimal foraging theory (OFT) predicts that higher quality foraging patches will lead to larger group sizes (Schoener 1971; Hirth 1977) and more time spent feeding versus vigilance (Lima and Dill 1990; Lima 1995) or moving (Ranglack and du Toit 2015c). Also, ideal free distribution (IFD) theory predicts that the equilibrium distribution of organisms among habitats of different quality, such as results after some patches of rangeland have, or have not, been subjected to natural disturbances or habitat manipulation (e.g., fire or mechanical treatment), will indicate the relative resource qualities of those habitats (Fretwell and Lucas 1969; Fretwell 1972).

Bison meet the main assumptions of IFD theory (Fretwell 1972) in that they are energy maximizers (Van Vuren 2001) and they are long-lived animals that, in most cases, have been present on the landscape of interest for several generations, allowing many foraging patches to be discovered and known. However, learning and memory have also been shown to be important in bison habitat selection (Merkle et al. 2014, 2015a, b; Sigaud et al. 2017). Bison can remember pertinent information about the location and quality of different foraging sites, and thus use that information to choose foraging areas where energy gains could be maximized (Merkle et al. 2014). However, site fidelity still plays an important role, as bison may not always choose the most productive foraging sites, as predicted by OFT and IFD, but show fidelity to previously visited foraging sites (Merkle et al. 2015a). The fusion-fission society of bison herds, however, creates a situation where individuals within a group may have different knowledge of foraging sites, and group familiarity combined with individual knowledge influence decisions on whether to follow the group to a foraging area that may be unknown to the individual but is known by the group, or to leave the group and return to a familiar foraging patch (Merkle et al. 2015b). This memory based foraging strategy allows bison to sample new foraging areas, whether being led there by individuals who have already have knowledge of the site (Merkle et al. 2015b) or through random patch use (Sigaud et al. 2017) that allows for higher energy gains in bison (Merkle et al. 2017). Rangeland managers may therefore use passive techniques for managing distribution through manipulation of habitat quantity and quality across the rangeland, and active techniques such as herding or fencing to manipulate bison distribution both in the present and future.

Abundance is a special consideration for bison, given their unique situation behind barriers—biological, social, physical, political, or otherwise. Rangeland managers must take special care to understand the carrying capacity of the range, both biological and social, and then maintain the appropriate density (Plumb et al. 2009; Steenweg et al. 2016; Cherry et al. 2019), which, given the extremely broad distribution of bison (Figs. 23.1 and 23.2) will vary dramatically and be unique to each bison herd. Unlike many other rangeland wildlife species whose abundance is maintained either largely through sport hunting or natural processes, bison abundances are managed primarily through human activities. In many areas bison populations are maintained through sport hunting (Ranglack and du Toit 2015a), but the most common tool for removing excess bison is using round-ups with excess animals being sold, donated, or culled (Millspaugh et al. 2008; White et al. 2011; Giglio et al. 2018).

9 Disease

Following the catastrophic decimation of wild bison by the late nineteenth century, the remaining few wild bison increasingly encountered domestic cattle and their diseases. By 1917, bison at Yellowstone National Park were infected with the non-native disease brucellosis via contagion from domestic cattle that were kept in the park to provide milk and meat for the US Army stationed at Fort Yellowstone (Meagher and Meyer 1994). Brucellosis and tuberculosis were subsequently introduced into the greater Wood Buffalo National Park area in Canada when plains bison, previously infected via cattle, were relocated there during 1925–1928, and through contagion from local cattle herds (Tessaro 1992). Currently, brucellosis, tuberculosis, and anthrax are focal diseases in some populations of wild bison (Aune et al. 2010, 2017; Plumb et al. 2014; National Academies of Sciences Engineering and Medicine 2020). Malignant catarrhal fever can also occur as acute and chronic cases in individual bison, most often associated with mixing with a carrier such as domestic sheep (Schultheiss et al. 1998, 2001).

9.1 Brucellosis

Bovine brucellosis is caused by the bacterium Brucella abortus that lives as a facultative intracellular parasite (Thorne 2001). Brucellosis causes abortions, still births, and can cause crippling arthritis in infected joints (Williams et al. 1997; Thorne 2001; Geremia et al. 2008; National Academies of Sciences Engineering and Medicine 2020). Only infectious pregnant females have a high probability of shedding live Brucella bacteria, that can be transmitted intra-specifically from fecund females to their offspring or through lactation, and inter-specifically through the ingestion of live bacteria from infected birth tissues (Cheville et al. 1998; Rhyan et al. 2009). Sexual transmission from males to females is rare in bison (Robison 1994). Brucellosis can result in lower pregnancy rates and population growth rates (Geremia et al. 2008), but otherwise does not affect adult wild bison survival or limit population increase in the wild (Fuller et al. 2007a, b; National Academies of Sciences Engineering and Medicine 2020).

9.2 Tuberculosis and Anthrax

Bovine tuberculosis is caused by the bacterium Mycobacterium bovis and is presented as acute debilitating pathology to respiratory, digestive, urinary, nervous, skeletal, and reproductive systems, with transmission by ingestion or inhalation, and from mother to offspring through the placental connection or contaminated milk (Tessaro et al. 1990). Tuberculosis can contribute to reduced population growth rate in wild bison through fetal losses and decreased pregnancy rates due to poorer condition and increased vulnerability of older animals to predation (Tessaro et al. 1990; Joly and Messier 2005). Anthrax is caused by the bacterium Bacillus anthracis and is transmitted by inhalation or ingestion of endospores, that are non-reproductive forms of the bacterium that can remain viable but dormant for decades before reactivating when environmental conditions become favorable (Aune et al. 2010). Anthrax is detected primarily in mature male bison (Dragon et al. 1999) and thus this disease appears to have little influence on bison population dynamics unless operating in conjunction with other limiting factors (Aune et al. 2010).

9.3 Disease Management

Chronic infection of wild bison populations with diseases that can be transmitted to livestock and humans is an important factor affecting potential recovery of bison outside existing reserve boundaries (Gates et al. 2001; Plumb et al. 2009). While the management of brucellosis and tuberculosis in wild bison is warranted based on risk to livestock and human health (Aune et al. 2010), management authorities face difficult challenges in restoring bison and their ecological processes (e.g., migration, dispersal, grazing influences) while preventing transmission to domestic cattle near reserve boundaries (White et al. 2011). In the Greater Yellowstone Area (GYA) of Montana, Wyoming and Idaho, bison were likely the first chronic wild reservoir of brucellosis, yet wild elk are now the primary wildlife host and all recent cases of brucellosis in GYA cattle are traceable to elk, not bison (Scurlock and Edwards 2010; National Academies of Sciences Engineering and Medicine 2020). Elk now maintain chronic infection at the population level, even when there is no direct contact with feed grounds or with infected bison (Cross et al. 2010a, b, 2013; National Academies of Sciences Engineering and Medicine 2020). Despite chronic brucellosis infection at the population level, lack of transmission from bison to cattle is likely a result of spatial and temporal separation management practices outlined in the Interagency Bison Management Plan (2016) combined with fewer cattle operations on some lands adjacent to areas used by bison during late winter, e.g., during the third trimester of pregnancy when potential shedding of the bacteria into the open environment is highest; thereby effectively managing the transmission risk, as opposed to a lack of transmission risk itself (Rhyan et al. 2009; Treanor et al. 2015; National Academies of Sciences Engineering and Medicine 2020). Detection of brucellosis in domestic bison and cattle automatically invokes mandatory national and state domestic animal health regulations including test-slaughter of infected individuals and potential for depopulation of entire herds (Cheville et al. 1998; National Academies of Sciences Engineering and Medicine 2020). However, these domestic livestock regulations do not automatically apply to free-ranging bison managed under federal or state wildlife authorities. In circumstances like the Greater Yellowstone Area, where wild bison and elk are the last remaining chronic reservoir of brucellosis across the United States, federal and state livestock and wildlife authorities are developing and implementing cooperative long-term adaptive risk management strategies to inform timely and evidence-based decisions for reducing the risk of transmission, including risk management through spatial and temporal separation of bison and cattle, iterative hypothesis testing and periodic scientific assessments (Cheville et al. 1998; Plumb et al. 2007; Gates et al. 2010; Nishi 2010; Geremia et al. 2011; Hobbs et al. 2015; National Academies of Sciences Engineering and Medicine 2020).

10 Ecosystem Threats and Conservation Actions

10.1 Ecosystem Threats

The most recent IUCN Red List Assessment for the wild American bison classified the species as “Near-Threatened,” and determined that as the wild species is completely dependent on active conservation protection, as it would likely revert to “Vulnerable” status if government protection was reduced (Aune et al. 2017). Additionally, the first IUCN Green Status Assessment for bison found bison to be critically depleted (Rogers et al. 2022). In Canada, an assessment of plains bison on rangelands in Canada recommended listing them as ‘Threatened’ (COSEWIC 2004, 2013). Otherwise, wild plains bison on rangelands have no formal protected status outside of the authorities of the agency jurisdiction, e.g., wild bison can be hunted on some US Forest Service lands but not on adjacent National Park Service lands. The plains bison has been unsuccessfully proposed several times for designation under endangered species authorities in Canada and the United States (Gates et al. 2010). Key threats to the species in the wild include habitat loss and fragmentation, genetic manipulation of commercial bison for market traits, small population effects in most conservation herds, few herds that are exposed to a wide range of natural selection factors, contemporary effects from historical cattle gene introgression, and the threat of depopulation as a management response to infection of some wild populations hosting reportable cattle disease (Gates et al. 2010; Plumb et al. 2014; Aune et al. 2017).

Approximately 93% of all American bison living on rangelands are legally managed as domestic livestock (Gates et al. 2010), with only four states in the U.S., two provinces in Canada, and one state in Mexico (Arizona, Utah, Montana, Wyoming, Alberta, Saskatchewan, and Chihuahua) that manage free-ranging bison as wildlife. Idaho, Missouri, New Mexico, and Texas designate bison as wildlife, but do not have any free ranging populations, and all other states and provinces across the continental historical range designate the species as domestic livestock. Hunting of wild bison on rangelands as a public trust wildlife resource occurs only in Arizona, Utah, Montana, Wyoming, Alberta, and Saskatchewan (Gates et al. 2010). Alarmingly, across all jurisdictions and legal authorities, less than 3% of all American bison on rangelands are managed as wildlife under a meaningful array of natural selection factors (e.g., competitive mate selection, predation, winter kill); while the vast majority of all bison on rangelands are otherwise subjected to anthropogenic selection for preferred population size and demography, body conformation, ease of management handling, and/or genetic manipulation to enhance profitable commercial bison market traits (Gates et al. 2010; Plumb et al. 2014; Aune et al. 2017). Despite long-term public investment in wild bison conservation, the private sector has far out-stripped wild bison, resulting in a potentially divergent evolution trajectory towards species domestication.

Sanderson et al. (2008) found the full continuum of the species now exists on < 2% of its historic range, and most of these sub-populations are managed in isolation with surplus offtake going to start new small, isolated herds or into food markets. There are only three free-ranging, wild, American bison sub-populations on rangelands that are greater than minimum viable population size (e.g., 400–1,000 total animals), that include only ~ 4,200 mature animals in total (Aune et al. 2017). All other herds managed as wildlife on rangelands live behind fences or are less than 400 animals. Despite this greatly unbalanced meta-population structure, there are no formal, long-term, wild bison meta-population viability strategies in place by any major stewardship sectors. Hartway et al. (2020) modeled several meta-population management strategies and found that translocations of selected age-sex class individuals between herds at the decadal scale could buffer these isolated herds against genetic drift at the century scale. This investigation also highlighted the critical importance of creating and maintaining additional large sub-populations under natural selection as core reserves for long-term genetic viability of the species.

Climate change may represent the next major challenge to bison, as it is expected to directly affect bison through decreased forage and water availability and increased thermal stress (Craine et al. 2009, 2013; Craine 2013; Martin and Barboza 2020a). There are also indirect effects of climate, through changes in the distribution and intensity of parasites (Patz et al. 2000; Kutz et al. 2005; Morgan and Wall 2009) and diseases (Janardhan et al. 2010) that have been shown to reduce reproductive success in bison (Fuller et al. 2007a). It is estimated that, with a 4 °C increase in global temperatures, these stressors could reduce bison body size by as much as 50% by the end of the twenty-first century (Craine 2013; Martin et al. 2018; Martin and Barboza 2020b). This could be compounded by the impacts of climate change on agriculture intensification, land use, and woody plant encroachment (Knapp et al. 1999; Allred et al. 2013; Bowler et al. 2020; Klemm et al. 2020). These threats, combined with the differences in bison management practices between sectors have led Martin et al. (2021) to classify bison as moderately vulnerable to climate change, recommending the creation of a ‘bison coalition’ that could seek climate change adaptation solutions through shared stewardship. That said, changing climate may also open new areas of habitat for bison north of their historical range, leading some to believe that bison may be able to occupy niches currently occupied by moose in boreal regions, though this has yet to be formally evaluated.

10.2 Conservation Actions

While much of the continental historical range is no longer available for recovery due to land use conversion as well as concerns about human safety and property damage, lack of local public support, and lack of funds for management as publicly owned wildlife (Boyd 2003; Plumb et al. 2009; Gates et al. 2010; Ranglack and du Toit 2015a), there are exciting conservation opportunities that are finding voice through a new vision of “Shared Stewardship” that embraces innovative collaboration to work together across jurisdictions and sectors to successfully address the scale, complexity, and ecological and cultural significance of wild bison conservation and restoration (Sanderson et al. 2008; Aune et al. 2017; Aune and Plumb 2018; Augustine et al. 2019; Martin et al. 2021; Pejchar et al. 2021). The US Department of Interior recently updated the charter for the federal “Bison Conservation Initiative” with a vision and commitment to leadership and alliances to ensure the conservation and restoration of wild American bison, focusing on wild, healthy bison herds, genetic conservation, and shared stewardship for ecological and cultural restoration. In Canada, a new free-ranging wild bison population was recently created in the foothill rangelands of Banff National Park, and a new expansive meta-population viability assessment is now being conducted for all federal and provincial bison herds in Canada (Wilson, personal communication). In 2014, the Blackfeet Nation, Blood Tribe, Siksika Nation, Piikani Nation, the Assiniboine and Gros Ventre Tribes of Fort Belknap Indian Reservation, the Assiniboine and Sioux Tribes of Fort Peck Indian Reservation, the Salish and Kootenai Tribes of the Confederated Salish and Kootenai Indian Reservation, and the Tsuu T’ina Nation came together to sign the “Northern Tribes Buffalo Treaty” that formally establishes intertribal alliances for cooperation in the restoration of American bison on Tribal/First Nations Reserves or co-managed lands within the U.S. and Canada. Collectively, these Tribes/First Nations own and manage about 6.3 million acres of grassland and prairie habitats. The Northern Tribes Buffalo Treaty is a formal expression of political unity to achieve ecological restoration of the buffalo tribal lands, and in so doing to reaffirm and strengthen ties that formed the basis for traditions thousands of years old. In Montana, the Blackfeet Nation has recently embarked upon the Iinnii Initiative to restore wild bison to Blackfeet lands in partnership with Glacier and Waterton National Parks in Montana and Alberta. Another example of “Shared Stewardship” is the Sicangu Lakota Oyate nation living on the Rosebud Indian Reservation, South Dakota, who is partnering with the U.S. Department of the Interior and World Wildlife Fund for a herd of 1500 buffalo on tribal lands, which would make it the largest owned by a Native nation.

The Nature Conservancy (TNC) initiated a long-term program of bison stewardship in 1984 at the Samuel H. Ordway, Jr. Memorial Prairie on ~ 8,000 acres of virgin unplowed prairie on the Missouri Coteau in north-central South Dakota. This TNC program has now been expanded to include bison under conservation management on 13 preserves across the Great Plains. The American Prairie Reserve in northeastern Montana is a private lands project of the non-profit American Prairie Foundation, that is envisioned to include over 3 million contiguous acres (12,000 km2) through a combination of both private and public lands to establish a fully functioning mixed grass prairie ecosystem, complete with several thousand migratory bison. In recognition of the ultimately potentially critical role of private lands for bison conservation, the National Bison Association published Conservation Management Guidelines for Herd Managers in partnership with World Wildlife Fund, with the goal to conserve the wild characteristics of bison on private lands through the conservation of the species’ genetic and behavioral traits while at the same time supporting ecosystem function and biodiversity conservation goals on the range the herd inhabits (World Wildlife Fund 2013).

11 Research/Management Needs

The biggest challenges facing the species have been entrenched for decades and are largely driven by socioeconomic and political forces. The continuum of bison ranging from free-ranging wildlife to domestic livestock is structured around multiple legal designations of the species, which also ranges from wildlife to domestic livestock, or in some cases both depending on ownership. Only 10 U.S. states, 4 Canadian provinces, and one Mexican state legally classify bison as wildlife, with all other states and provinces within the historic range designating bison as a livestock species only (Aune and Wallen 2010).

All bison, whether wild or wild with limitations are ultimately restricted in their large landscape movements by a barrier—biological, social, physical, political, or otherwise—as “terrestrial castaways” (Ritson 2019). Fragmentation of the entire species into a disjunct metapopulation needs to be better understood and should be at the forefront of all management decisions. In the private sector where live bison are regularly bought and sold, the mixing of animals of different genetic lineages is common, but the public conservation sector has seen little of this. The recent population viability analysis previously discussed highlighted the need for management of DOI herds within a metapopulation framework to maintain genetic viability over the next 200 years (Hartway et al. 2020). Questions of genetic diversity and viability have previously had to be balanced by concerns over introgression of cattle genes into the bison genome, necessitating that the “pure” bison be managed in a separate meta-population from the introgressed bison. However, given recent advances in genetics, it appears that cattle introgression is more common and widespread than previously thought (Stroup et al. 2022). As such, maintaining genetic diversity and population viability should be the guiding factor in metapopulation management, rather than attempting to eliminate introgression (Dratch and Gogan 2010). With wild bison numbering < 5% of the total bison abundance, it is crucial we monitor the abundance, distribution, and demographics of “wild” and “wild with limitations” bison (see Aune et al. 2017). New quantitative science based upon population viability analyses is needed to identify how many herds of variable size and level of isolation are needed for the long-term conservation of the species as wildlife. These are questions that are rarely asked of other rangeland wildlife, but for bison they are crucial for understanding the status of the species and planning for long-term conservation of bison into an uncertain future, with the ultimate goal of moving bison from “Near Threatened” to “Least Concern” IUCN Red List status (Aune et al. 2017).

12 Summary

Innovative approaches to “Shared Stewardship” of bison across sectors will likely be required to change current paradigms and limitations that imply that the only bison that really “matter” are those under explicit conservation management and that bison must always be behind fences and separate from cattle. Changing these paradigms requires a conscious effort to reach beyond what has typically been considered as conservation and recognize the strength provided by a diverse portfolio of management goals and strategies, along with a recognition, both legally and otherwise, that bison are a valuable native wildlife species that should be allowed to exist in the same manner as other native rangeland wildlife species. At an ecoregional scale, Plumb and McMullen (2018) reviewed whether the Colorado Plateau in northern Arizona should be included within the species historical range and highlighted how difficult it can be to challenge entrenched dogma and move forward with updated perspectives while avoiding confusing scientific inferences with societal value judgments. They further characterized sustainable bison conservation as occurring at the intersection of best available multidisciplinary science, compliance with law and policy, and long-term public interest; so that sustainability embraces historical reference conditions and a balance of local social and ecological concerns, all while demonstrably contributing to continental-scale bison conservation.

Bison have one of the strongest links to human culture of any species in North America, and those cultural connections are important for the restoration of both bison and native peoples, but also allows for bison to serve as an icon for all people. Sanderson et al. (2008: 252) presented an overarching continental vision statement “Over the next century, the ecological recovery of the North American bison will occur when multiple large herds move freely across extensive landscapes within all major habitats of their historic range, interacting in ecologically significant ways with the fullest possible set of other native species, and inspiring, sustaining and connecting human cultures.” In response to scientific evidence and calls for a more inclusive approach to long-term bison conservation, the DOI Bison Conservation Initiative was recently updated with a bold and expansive approach to what will constitute the Department’s bison portfolio moving forward (Department of the Interior 2020). The five main goals of the Bison Conservation Initiative capture key issues for long-term conservation of wild bison on rangelands of North America: (1) conserving bison as wildlife by minimizing artificial selection and allowing natural selection to operate, (2) genetic conservation through metapopulation management, (3) shared stewardship with states, tribes, and other stakeholders, (4) ecological restoration achieved through shared stewardship to establish and maintain large, wide-ranging bison herds and large landscapes where the full ecology of the species can function, and (5) cultural restoration through collaboration with Tribes and First Nations (see Shamon et al. 2022). These goals are universally applicable across the bison continuum, and although ecological recovery of wild free-ranging bison has been relatively static for many decades, there are now new and exciting alliances and opportunities that suggest the era of big conservation is not over; that indeed there is still hope, and after all, hope is a bison (Redford et al. 2016).