Black-Tailed and Mule Deer

Abstract

Black-tailed and mule deer (both designated as Odocoileus hemionus; hereafter referred to as “deer” or “mule deer”) comprise an iconic species that is broadly distributed across western North America. This species occurs in all rangeland types including grasslands, desert shrublands, forests, savannah woodlands, and even portions of tundra. The distribution of mule deer has changed little since Euro-American settlement, but abundance has fluctuated in response to environmental variation and rangeland management practices. These deer are medium-sized, polygynous mammals classified as generalist herbivores (foregut fermenters). Population growth in this species is strongly influenced by survival of adult females and recruitment of young. The management of rangelands has direct influence on deer populations given the wide distribution of this species and measurable responses to rangeland management practices. Rangeland management practices including development of water, grazing by domestic livestock, prescribed fire, energy extraction, vegetation alteration, and others can have positive or negative influences or both on this species. Although mule deer are widely distributed and relatively abundant, conservation of this species is challenged by rapid changes currently occurring on rangelands of western North America. Altered fire regimes due to climate change and invasive plants, competition (with feral horses [Equus ferus caballus], livestock, and other wild ungulates), development of energy, ex-urban and urban expansion, and many other challenges threaten continued abundance of this species. Rangelands and their associated management will continue to play a disproportionally large role in the conservation of mule deer in the future.

1 General Life History and Population Dynamics

Black-tailed and mule deer (both designated as Odocoileus hemionus and hereafter referred to as “deer” or “mule deer”) are polygynous breeders with the breeding season occurring during the late fall or early winter and a gestation period of 199–208 days or approximately 7 months (Anderson 1981). Pregnancy rates for prime-aged (between 2 and 6 years of age) females typically exceed 90% with limited variation in response to environmental or rangeland conditions (Freeman et al. 2014; Montieth et al. 2014). Pregnancy rates for yearling females (1.5-year-olds during fall/winter breeding season), however, are generally lower and more variable across rangelands and in response to environmental conditions (Lawrence et al. 2004; Montieth et al. 2014). Likewise, pregnancy rates for older (≥ 6.5) females are also more variable than prime-aged females and have been reported as low as 73% in arid rangelands (Lawrence et al. 2004). Timing of parturition varies across latitudes with earlier parturition at northern latitudes and later parturition at more southern latitudes, ostensibly in response to the increased influence of monsoon moisture during summer on rangelands in western North America from north to south (Freeman et al. 2014; Stoner et al. 2016).

Typically, one or two offspring are born during the spring or summer following gestation. Number of offspring varies with age (average litter size reduced for yearling females) and condition of parturient females (Montieth et al. 2014). Mean body mass of neonatal mule deer at birth was estimated at 3.4 kg, but varies across rangelands from 2.7 to 4.0 kg depending on condition of females, presence of a twin, location, and year (Lomas and Bender 2006). Neonatal mule deer are weaned over the summer and fall months prior to the subsequent breeding season and most females have offspring each year (Bowyer 1991).

Survival of mule deer varies by age, sex, and across rangelands in western North America. Mean estimates of survival for neonates (from 0 to 6 months of age) range between zero and 62% depending on location and year with high rates of predation commonly observed (Pojar and Bowden 2004; Lomas and Bender 2006; Montieth et al. 2014; Shallow et al. 2015). A weighted average for mean survival to 6 months of age using data collected from across the range of mule deer was 44% (95% CI: 33% to 55%) (Forrester and Wittmer 2013). Survival for fawns (from 6 months of age through first winter) was also averaged at 44% (SE = 3%), but considerable variation was found with estimates ranging between 4% (SE = 3%) and 81% (SE = 7%) depending on location and year (Unsworth et al. 1999). Observed rates of survival for both neonates and fawns are lower and more variable than commonly observed with other ungulates (Forrester and Wittmer 2013). Survival of adult females is less variable than survival of neonates or fawns with a weighted mean estimated from across the range at 84% (95% CI 75–94%) (Forrester and Wittmer 2013). For most populations, annual survival of adult males is strongly dependent on harvest by humans with estimates of survival ranging from 60 to 92% (Pac and White 2005; Bender et al. 2012).

Population dynamics for ungulates are typically most strongly influenced by survival of adult females which is often relatively high and stable (Gaillard et al. 2000). Some estimates suggest adult survival for mule deer is nearly four times more influential than other demographic rates on population growth (Lukacs and Nowak 2023). In other research however, survival of fawns and recruitment were identified as most influential on population growth as 4 of 5 studies found this demographic rate more influential than adult survival (Forrester and Wittmer 2013). This discrepancy may occur because survival of juvenile deer was lower and more variable than observed with other ungulates (Forrester and Wittmer 2013). Moreover, unlike many ungulates, prime-aged deer typically give birth to twins and consequently may rely on relatively high fecundity rates as a driver of population growth (Forrester and Wittmer 2013).

As with other ungulates, predation is the most commonly identified source of mortality for neonates between 0 and 6 months of age and typically accounts for between 50 and 100% of all mortalities (Linnell et al. 1995). Fawns (6 months to 1.5 years of age) are also vulnerable to predation, but high mortality rates are also regularly observed at northern latitudes during winter due to starvation. A diverse group of predators have been reported to prey on mule deer during their first year of life with coyotes (Canis latrans), mountain lions (Puma concolor), bobcats (Lynx rufus), wolves (Canis lupus), and black bears (Ursus americanus) reported as the primary predator depending on years evaluated and location (Forrester and Wittmer 2013). Predation is also the most cited cause of mortality for adult deer with between 22 and 66% of all mortalities due to predation from primarily mountain lions and wolves with secondary sources of mortality identified as disease, malnutrition, vehicle strikes, and miscellaneous other causes (Forrester and Wittmer 2013).

Nonetheless, mule deer do show evidence of density dependence (Bergman et al. 2015). Adult male to adult female ratios, for example, are negatively correlated with ratios of production suggesting intraspecific competition occurs and harvest strategies designed to increase the proportion of males in a population may have a regulating effect on population growth (Bishop et al. 2005a; Bergman et al. 2011, 2015). Likewise, a severe (76%) reduction in density of a mule deer population in Colorado was associated with a more than two-fold increase (31% to 77% in control and treatment areas, respectively) in fawn survival (White and Bartmann 1998). Interestingly, in the same area, reductions in density of 16% and 22% were not associated with increases in overwinter fawn survival (Bartmann et al. 1992).

Mule deer are of tremendous interest to the public for hunting, collection of shed antlers, and wildlife viewing. For many western states and provinces, demand for deer hunting permits greatly exceeds available supply. Sales of hunting permits and excise taxes on purchases of hunting equipment provide millions of dollars annually to state and provincial agencies to be used for conservation and management.

Because mule deer follow a polygynous breeding system, harvest management strategies are often focused on removal of males (less likely to influence population dynamics) and most exploited populations have male/female ratios that are heavily skewed towards females. Selective harvest of males has mixed effects on mule deer populations. Rates of pregnancy (99% vs. 97%) and timing of parturition at comparable latitudes, for example, were similar for populations with relatively low (14 males per 100 females) and relatively high (26 males per 100 females) male/female ratios, respectively (Freeman et al. 2014). This finding suggests that male/female ratios as low as 14 males per 100 females did not limit breeding opportunity or influence dates of conception. Moreover, increases in male/female ratios have been associated with decreased fawn/adult female ratios suggesting that annual production is actually reduced in the presence of high male/female ratios (Bishop et al. 2005a). This effect may be caused by competition for resources between adult females and adult males resulting in poor condition and reduced productivity for adult females.

Hunting can influence other aspects of the ecology of mule deer. Number of males harvested and timing of harvest, for example, were correlated with reduced prevalence of chronic wasting disease, an emerging conservation challenge for this species (Conner et al. 2021). Additionally, selective harvest of large males over the last century has been proposed as the most likely cause of an approximate 3% decline in antler size (Monteith et al. 2013). Further, disturbance associated with hunting can influence movement rates and habitat selection. Female mule deer demonstrated increased movement rates during the daytime during hunting season, but similar movement rates during the night while maintaining high site fidelity (Brown et al. 2020). Similarly, at relatively fine scales (within summer range areas and at stopover sites along migration routes), female mule deer used habitats that retained high-quality forage consistently throughout the hunting season whereas male mule deer selected more secure habitats away from motorized routes (Rodgers et al. 2021). Nonetheless, conservation of mule deer has benefitted from the interest and revenue associated with pursuit and harvest of this species.

2 Species and Population Status

The genus Odocoileus, which contains mule deer along with white-tailed deer (O. virginianus) is one of 18 genera found across the world in the family Cervidae (Wilson and Mittermeier 2011). This genus first evolved in North America approximately 3.5 million years ago with a form similar to white-tailed deer present in North America at least 3 million years ago (Miller et al. 2003). Eleven subspecies of O. hemionus are recognized, but little genetic or morphological variation is found between many of them and they are subsequently grouped into two morphological types: black-tailed deer (O. h. columbianus and O. h. sitkensis) and mule deer (O. h. hemionus, O. h. fulginatus, O. h. californicus, O. h. inyoensis, O. h. eremicus, O. h. crooki, O. h.peninsulae, O. h. sheldoni, and O. h. cerrosensis) (Cronin 1991; Latch et al. 2009). Black-tailed deer are a relatively recent species that likely diverged from white-tailed deer within the last 500,000 years (Polzhein and Strobek 1998). Mule deer (the morphological type that includes all of the subspecies excluding O. h. columbianus and O. h. sitkensis) are even younger and likely evolved within the last 12,000 years following extinction of Pleistocene megafauna, glacial retreat, and putative hybridization between black-tailed deer and white-tailed deer in western North America (Geist 1998). Hereafter we refer to black-tailed deer (O. h. columbianus and O. h. sitkensis) and mule deer (remaining subspecies) as mule deer unless referring to an individual subspecies.

2.1 Historical Versus Current Distribution

Climatic oscillations and the subsequent expansion and retreat of glaciers in North America influenced the distribution of this species (Latch et al. 2009). Genetic analyses suggest that black-tailed deer persisted in a single refugium in the Pacific Northwest near the coast prior to expanding north and inland following the last glacial maximum (Latch et al. 2009). Conversely, mule deer likely persisted in multiple refugia in the southern portion of western North America from which they expanded north following the last glacial maximum (Latch et al. 2009).

More recently, mule deer were first encountered by Lewis and Clark while traveling up the Missouri River (Hays 1869). Lewis and Clark also noted the presence of mule deer as they journeyed across the northern portion of the United States with black-tailed deer (O. h. columbianus) observed west of the Cascade Mountains (Kay 2007). Rock art from multiple Native American cultures also provides evidence that mule deer occurred in locations consistent with their current distribution in the southern portion of their range (Murray 2013). Moreover, mule deer were also noted in the journals of early explorers of the southwestern United States including Dominguez and Escalante (Rawley 1985).

2.2 Distribution Map

Black-tailed deer and mule deer currently occupy portions of 6 Canadian provinces, 6 states in Mexico, and 18 western states in the United States of America (Fig. 17.1; Heffelfinger and Latch 2023). Comparison of the current distribution with early accounts of this species from Euro-American explorers suggest only a few differences since the seventeenth and eighteenth centuries. Lewis and Clark, for example, first encountered mule deer along the Missouri River (Hays 1869) which is near the eastern boundary currently recognized for mule deer (Fig. 17.1; Heffelfinger and Latch 2023). Moreover, pictographs and other archaeological records coupled with early records from Dominguez and Escalante (Rawley 1985) also suggest little change in distribution in recent centuries.

Fig. 17.1

Distribution map depicting the current distribution of mule deer (Odocoileus hemionus) and black-tailed deer (subspecies O. h. sitkensis and O. h. columbianus) modified from Heffelfinger and Latch (2023). The distribution of this species remains largely unchanged since Euro-American settlement of western North America

2.3 Historical Versus Current Abundance

Despite a broad distribution during the period of exploration by Euro-Americans in western North America, mule deer may have persisted at relatively low abundance in many parts of their range. On their journey up the Missouri River and across the northern portion of the western United States, Lewis and Clark harvested more white-tailed deer than all other large mammals combined and mule deer were rare (Kay 2007). This discrepancy may have resulted from differences in habitat selection as the Lewis and Clark expedition navigated along rivers where white-tailed deer were perhaps more abundant. Alternatively, behavioral differences between mule deer and white-tailed deer may also explain the difference as mule deer were thought to have elevated risk of harvest by Native Americans because some of the highest encounter rates for this species occurred along tribal boundary zones (Geist 1998; Kay 2007). In the southern portion of their range, mule deer are outnumbered at least 2 to 1 by bighorn sheep (Ovis canadensis) in rock art suggesting reduced abundance historically relative to that ungulate (Castleton 2002). Additionally, journals from many early Euro-American settlers in areas of western North America where mule deer are now abundant suggest they were only encountered rarely during the late 18th and early nineteenth centuries (Rawley 1985).

Mule deer populations undoubtedly fluctuated in response to environmental conditions and landscape changes (e.g., fire). Some records suggest this species was at least locally abundant in some portions of its range during the nineteenth century. Market hunter Frank Mayer, for example, harvested over 250 ungulates between late August and early November 1878 in Middle Park, Colorado and 89 of these were mule deer (Gill et al. 1999). Expanding settlement by humans and unregulated harvest including market hunting, however, led to severe declines for most ungulate species in western North America during the late 19th and early twentieth centuries (Krausman and Bleich 2013).

Following these severe declines, regulations on harvest coupled with human-induced changes to landscapes and predator communities led to population growth for mule deer during the twentieth century. The famous Kaibab mule deer population in Northern Arizona, for example, was estimated at 4,000 in 1906, but at least 7 times that number in 1924 less than two decades later (Caughley 1970). Moreover, between 1930 and 1960, mule deer populations irrupted in the Intermountain West and likely reached all-time high abundance in recorded history between 1940 and 1950 (Gruell 1986). By the late 1940s, mule deer populations in at least portions of each of the western United States were recognized as over-populated and a cause of rangeland degradation (Leopold et al. 1947; Binkley et al. 2006). Four competing hypotheses have been proposed for this irruption. These hypotheses include: (1) conversion of rangelands to shrub-dominated systems favored by mule deer following overgrazing by domestic animals during the early twentieth century, (2) logging of forested landscapes and subsequent succession of those plant communities to forbs and shrubs, (3) widespread predator control including use of poison, and (4) reductions in livestock grazing following implementation of the Taylor Grazing Act (Gruell 1986). Although some authors favor conversion of rangelands to shrub-dominated systems as a leading hypothesis (Gruell 1986), it is impossible to disentangle that idea from the other concepts as landscape change, predator control, and reductions in grazing occurred simultaneously and may have been synergistic.

Following the historic highs of the 1950s and 1960s, mule deer populations appeared to decline across their range, although good estimates of abundance are not available until the latter part of the twentieth century. During the last half of the twentieth century and early part of the twenty-first century, estimates of abundance for mule deer populations fluctuated in both space and time with periods of general decline (late 1970s, early 1990s) and growth (early 1980s, early 2010s) noted across their range (Mule Deer Working Group 2015). The most recent estimates from governmental entities that participate in the Western Association of Fish and Wildlife Agencies suggest increasing populations in 6 jurisdictions (provinces or states), stable populations in 10, and declining populations in 7 (Mule Deer Working Group 2021). Current estimates of abundance suggest approximately 4 million mule deer occur across their range making this species the most abundant ungulate in western North America (Mule Deer Working Group 2021). Mule deer are listed as “stable” and designated a species of “least concern” by the International Union for Conservation of Nature (IUCN) (Sanchez-Rojas and Tessaro 2016).

2.4 Monitoring

Declines in mule deer populations following historic highs in the 1940s and 1950s prompted state and federal agencies to develop methods to estimate abundance. Mule deer are now monitored extensively on rangelands across western North America using a variety of methods (Keegan et al. 2011). Because management of deer resides with individual agencies within provincial and state governments, there are a myriad of approaches and strategies used to monitor populations (Rabe et al. 2002).

Monitoring efforts often include surveying populations from the ground or by fixed-wing aircraft or helicopter and classification of age and sex composition (Freddy et al. 2004; Keegan et al. 2011). Some agencies use distance sampling (often using fixed-wing or rotary aircraft although this can also be done from the ground) which allows for estimation of density after first estimating detection probability (Koenen et al. 2002). This method assumes perfect detection of deer at the survey point or along the survey transect, limited movement of animals in response to the surveyor prior to detection, accurate measurement of distance from the survey point or perpendicular distance from the transect to each individual deer, and no double counting (Buckland et al. 1993). Consequently, this technique is most often used on rangelands with relatively high visibility where statistical assumptions can be reasonably attained. Detection probabilities can also be estimated on surveys if deer are uniquely marked (e.g., ear tags, collars). Detection probabilities vary from 19 to 86% and are associated with many factors including deer activity, group size, topography, vegetation, and weather conditions (Bartmann et al. 1986; Zabransky et al. 2016).

Harvest surveys are also commonly used to estimate abundance across the range where mule deer occur (Rupp et al. 2000). More recently, advancements with remote cameras have allowed estimation of abundance of deer using instantaneous sampling, space-to-event, and time-to-event models, which may prove particularly helpful on rangelands with reduced visibility (Moeller et al. 2018). Likewise, advancements with unmanned aerial systems and thermal imaging cameras show some ability to estimate parameters such as abundance and survival (Williams et al. 2020). Some agencies have also used DNA collected from fecal pellets (often referred to as fecal DNA or fDNA) to identify individuals and help with estimation of abundance (Furnas et al. 2018). Finally, deer are also commonly monitored via mark-recapture techniques and either GPS or VHF collars, which can provide estimates of fecundity, recruitment, and survival when the sample of marked individuals is monitored closely.

Most provincial or state agencies employ dozens of biologists to collect monitoring data for deer within their jurisdictions. Advancements in statistical analyses and computing power now allow for incorporation of multiple sources of data with varying degrees of uncertainty into estimates of abundance or density (Furnas et al. 2018). Data obtained on populations of deer (e.g., flight surveys, remote camera surveys, mark-recapture data from collared animals, etc.) are now commonly integrated with harvest information into population models to help with decision making (e.g., integrated population models; Riecke et al. 2019). Population models are then commonly used to adjust harvest permits or quotas on an annual or every few years’ basis in an adaptive management framework (Nagy-Reis et al. 2021). These efforts create a robust data set to monitor mule deer populations across their range.

2.5 Migration Ecology and Overcoming Barriers to Movement

Many mule deer populations are migratory including some individuals and populations with long distance (> 100 km) migrations (Sawyer et al. 2005). Nonetheless, tremendous variation in migratory behavior occurs including some individuals and populations that do not migrate (McCorquodale 1999; Van de Kerk et al. 2021). Migratory behavior allows some mule deer populations to maximize intake of nutritious forage on rangelands during spring green-up by prolonging the period when they can consume high-quality forage (Merkle et al. 2016). Conversely, migratory behavior is energetically costly and can be associated with increased risk of mortality as animals navigate migratory routes. Thus, a tradeoff exists for deer between prolonged access to nutritious forage and the ability to escape deep snows (both of which likely have fitness benefits) versus risks associated with migration. Long-distance migrants in one population of mule deer, for example, were exposed to increased risk of mortality from fences and highways, whereas animals that did not migrate as far experienced fewer of those risks (Sawyer et al. 2016). Migratory behavior for another population, however, was associated with higher survival for mule deer that migrated compared to non-migratory animals (Schuyler et al. 2019).

Knowledge of migration routes and timing is thought to be transmitted culturally for ungulates, but has not been specifically tested in mule deer (Jesmer et al. 2018). This cultural transmission of knowledge likely includes learning of both routes and stopover areas, the latter of which are increasingly recognized as important. Mule deer in one population, for example, took an average of 3 weeks to complete migrations, but spent 95% of that time in stopover areas (Sawyer and Kauffman 2011). These animals averaged use of a stopover every 5.3 and 6.7 km during spring and fall migrations, respectively (Sawyer and Kauffman 2011). Stopovers likely act as physiological refugia (forage and rest) and were associated with lower measures of stress hormones (fecal glucocorticoid metabolites) along a lengthy migration on rangelands in Wyoming (Jachowski et al. 2018). Although plasticity on whether to migrate and where to go during migration has been documented (Van de Kerk et al. 2021), many populations show strong (> 80%) fidelity to migration routes and stopover areas across years with limited evidence of plasticity (Sawyer et al. 2019).

Consequently, preservation of migratory routes and stopover areas is critical to conservation of mule deer populations. Terrestrial migrations in general are imperiled around the globe including on rangelands in western North America where deer occur (Middleton et al. 2019). Migratory behavior in mule deer populations is now threatened by rapidly changing landscapes on western rangelands. Anthropogenic activities and structures including fences, highways, homes, and extraction of natural resources can influence migratory behavior and ecology of deer. On a lengthy migratory route on rangelands in Wyoming, for example, it was estimated that mule deer crossed an average of five highways and 171 fences per year during migration (Sawyer et al. 2016). Mule deer have also been shown to increase rate of movement, decrease time in stopover areas, and shift the location of stopover areas in response to energy and residential development (Wyckoff et al. 2018). Likewise, housing development and roadways reduced the effective width of a bottleneck along a migratory route for over 2,500 migratory mule deer to < 0.8 km (Sawyer et al. 2005). Additionally, anthropogenic disturbance associated with development of natural gas resources was associated with delayed departure, but earlier arrival suggesting more rapid transit of migratory routes along with shorter migration distances (Lendrum et al. 2013).

Increased understanding of the importance of migratory routes and stopover areas for mule deer has led to conservation efforts to improve permeability along migration routes by removing movement barriers (e.g., fences) or constructing passage ways such as highway underpasses to facilitate movement. Fences, can alter movement and pose a risk of entanglement for migratory animals. Removal of fencing or replacement of fencing with designs that include a smooth lower wire that is raised (46 cm) and a shorter (107 cm) top wire can increase permeability and mitigate risk of entanglement (Segar and Keane 2020). Likewise, highway underpasses coupled with exclusionary fencing facilitated safe movement of nearly 50,000 mule deer over 3 years under highway 30 in southwestern Wyoming while reducing collisions with vehicles by 81% (Sawyer et al. 2012). In general, passage structures with high “openness ratios” (width multiplied by height/length) are considered most effective (Clevenger and Waltho 2000). Collaborative efforts between governmental agencies and vested stakeholders coupled with improved data from GPS collars have helped identify barriers to movement for migratory mule deer (Middleton et al. 2019). This information should be incorporated into landscape planning so that crossing structures and mitigation efforts preserve migratory routes and stopover areas for mule deer.

3 Habitat Associations

Mule deer occupy grasslands, desert shrublands, savannah woodlands, forests, and even portions of tundra. Deer are medium-sized, generalist herbivores that are foregut fermenters with small mouth parts (concentrate selectors) (Hofmann 1989). These characteristics assist mule deer in selecting the most nutritious parts of the many different plant species they consume across the varied rangelands where they occur.

Mule deer consume a variety of different plants throughout the year including forbs, grasses, shrubs, and trees (Stewart et al. 2003; Berry et al. 2019). Across their range, the relative importance of these functional groups, however, varies seasonally. During spring and summer on rangelands at northern latitudes, mule deer primarily consume forbs, grasses, and deciduous shrubs whereas during fall and winter when herbaceous plants senesce or are covered in snow, evergreen shrubs and trees constitute the majority of consumed forage (Scasta et al. 2016; Berry et al. 2019). On rangelands at southern latitudes, seasonal use of forbs and grasses varies in response to precipitation patterns that can be highly variable and shrubs often constitute a majority of consumed forage annually (Krausman et al. 1997; Marshal et al. 2012).

Although classified as concentrate selectors, mule deer have some characteristics consistent with intermediate foragers. Composition of volatile fatty acids in the rumen, for example, was similar between red deer (Cervus elaphus; intermediate forager) and mule deer (Prins and Geelen 1971). Likewise, papillae density, dry weight of rumen digesta, and intestinal length are all greater in mule deer compared to white-tailed deer allowing them to make use of less nutritious forage (Zimmerman et al. 2006). Relative to white-tailed deer, for example, mule deer required 54% less digestible protein and 21% less digestible energy per day to maintain body mass and nitrogen balance (Staudenmaier et al. 2021).

Moreover, mule deer have adaptations that allow them to process plant secondary compounds such as tannins and terpenes common in shrubs. Mule deer commonly consume multiple different plant species which often allows for overall greater forage intake than consumption of a single species when secondary compounds are present (Freeland and Janzen 1974). Additionally, mule deer, have relatively large parotid salivary glands that produce proteins that bind to tannins and ameliorate the impact of those secondary compounds on digestibility (Hagerman and Robbins 1993). Mule deer were also better able to process forage containing the monoterpene α-pinene compared to white-tailed deer, suggesting a more efficient and less energetically costly method of detoxifying this compound (Staudenmaier et al. 2021). Consequently, mule deer may have an advantage over white-tailed deer on rangelands dominated by low-quality forages that are chemically defended (Staudenmaier et al. 2021).

4 Rangeland Management

The rangelands where mule deer occur (Fig. 17.1) differ greatly in precipitation patterns, plant community composition, and soils. Management of these disparate rangelands also differs regionally and across jurisdictions. Consequently, it is difficult to make definitive statements concerning the influence of rangeland management activities on mule deer, or their habitat. Mule deer, like all species, require cover, food, space, and water in an arrangement where all are accessible (sensu Leopold 1933). When rangeland management activities promote these elements, populations benefit. Conversely, when rangeland management activities eliminate access to or degrade these essential components to habitat, populations decline.

4.1 Livestock Grazing

Dietary overlap between domestic livestock and mule deer covaries with species of livestock, rangeland type, and annual variation in availability of forage. Dietary overlap with cattle (Bos taurus) is often low (e.g., Stewart et al. 2003; Beck and Peek 2005), but can increase during years of low precipitation or with high deer density and heavy stocking rates (Campbell and Johnson 1983; Hansen and Reid 1975). Overlap between mule deer and domestic sheep (Ovis aries) also varies depending on plant community composition and season of grazing. Spring grazing by domestic sheep, for example, resulted in low (15%) dietary overlap with mule deer on a sagebrush (Artemisia sp.) rangeland in Colorado (MacCracken and Hansen 1981). Moderate overlap (22–65%, depending on year), however, was observed in summer in aspen (Populus tremuloides)-sagebrush communities in northeastern Nevada (Beck and Peek 2005).

Grazing by domestic livestock can have both positive and negative influences or both depending on the species of livestock, stocking rate, timing of grazing, and response of plant communities to grazing (Krausman et al. 2011). Influences of grazing by domestic livestock can be both direct and indirect (Chaikina and Ruckstuhl 2006). Presence of livestock, for example, can influence habitat selection leading to avoidance of areas used by livestock (Ragotzkie and Bailey 1991; Loft et al. 1991; Stewart et al. 2002). Avoidance of grazed pastures may occur even beyond the time when livestock are removed (Clegg 1994; Kinka et al. 2021). This interaction and the strength of selection against areas used by livestock, however, are influenced by stocking rate and density of deer with preference for ungrazed pastures diminished at low stocking rates or high deer densities (Austin and Urness 1986).

Mule deer altered their foraging behavior in relation to stocking rates of cattle by feeding for longer durations in areas grazed at high stocking rates when forage was limited, but not when herbaceous plants were abundant (Kie 1996). Removal of vegetation under moderate and high stocking rates can also lead to decreased hiding cover—particularly for neonates (Loft et al. 1987). Additionally, mule deer are more likely to compete with livestock for forage during dry years when forage production on rangelands is limited (Kie et al. 1991). Domestic livestock are potential vectors for exotic and invasive plants through both endozoochory (passage of viable seed through the digestive tract) and epizoochory (transport of seeds on skin and fur) (Chuong et al. 2016). On rangelands in the Northwestern portion of North America, cattle were estimated to disperse (via endozoochory) an order of magnitude more seeds from exotic grasses than either elk (Cervus canadensis) or deer (Bartuszevige and Endress 2008).

Conversely, grazing by livestock can also result in positive outcomes for rangelands. For rangelands dominated by grasses, grazing by cattle has been associated with removal of standing dead vegetation leading to more nutritious forage during the subsequent growing season (Short and Knight 2003; Taylor et al. 2004). Moreover, grazing by cattle produced more nutritious forage during the subsequent growing season than mowing in a rough fescue (Festuca scabrella) rangeland (Taylor et al. 2004). Similarly, grazing by domestic sheep can increase nutritional quality (e.g., protein content and digestibility) of plants—particularly shrubs—during fall and winter when those plants are grazed during spring at moderate (less than 55%) utilization (Rhodes and Sharrow 1990; Alpe et al. 1999).

Changes in vegetation due to grazing by domestic livestock, however, have not been directly linked to increased abundance, condition, or production in mule deer populations and caution is warranted. Nonetheless, suggested guidelines for management of grazing by domestic livestock on rangelands important to deer during winter include grazing during the spring to balance utilization of shrubs by deer during winter with consumption of forbs and grasses by livestock (Austin 2000). Alternating species of livestock along with moderate (50%) utilization in a rest-rotation system have also been suggested as best practices to maintain or enhance rangelands for mule deer (Jensen et al. 1972; Austin 2000).

Grazing of rangelands by domestic livestock can also be used prescriptively to enhance habitat for mule deer by reducing abundance of undesirable species. Targeted grazing of cheatgrass (Bromus tectorum) by cattle during spring, for example, was associated with reduced risk of catastrophic fire in the fall (Diamond et al. 2009). Use of grazing by domestic livestock to reduce invasive plants, however, is challenging across rangelands with large spatial extents because intensive management (e.g., electric fencing, supplemental protein and energy to overcome secondary compounds in targeted plant species) is typically required to reach high utilization rates (Popay and Field 1996; Dziba et al. 2007).

Fencing associated with management of domestic livestock can also influence deer. In western Wyoming, density of fences within mule deer range was estimated at 0.59 km/km² and mule deer encountered fences an average of 119 times per year (Xu et al. 2020). Fences altered normal movement patterns on nearly 40% of encounters (Xu et al. 2020). Moreover, mule deer can become entangled in fences resulting in injury or death (Fig. 17.2). Up to 0.08 mortalities per km of fencing were estimated annually for mule deer in Utah and Wyoming (Harrington and Conover 2006). Fencing mortalities peaked in August and juveniles were more likely to be entangled in fences than adult animals (Harrington and Conover 2006). Increased height of the bottom wire was associated with increased probability of successfully crossing fences (Jones et al. 2020). Woven wire fences with a single strand of barbed wire were the most lethal compared to 4-strand barbed wire fences or woven wire with two strands above the mesh (Harrington and Conover 2006). Replacement of 4-strand barbed wire fences with wildlife friendly fencing where the bottom wire was smooth and raised (46 cm) along with a shorter top wire (107 cm) was associated with an increase of over 18% in successful crossings indicating some ability to mitigate effects of fences with design modifications (Segar and Keane 2020).

Fig. 17.2

Mule deer (Odocoileus hemionus) caught in a fence used to manage livestock on federal rangelands. Photo credit to Jason Nicholes with the Utah Division of Wildlife Resources

4.2 Interactions with Coexisting Feral and Wild Ungulates

Mule deer coexist with several species of native and non-native, free-ranging ungulates throughout portions of their geographic range (Fig. 17.1). Free-ranging native ungulates with potential for interspecific interactions include American elk, white-tailed deer, moose (Alces alces), pronghorn (Antilocapra americana), bighorn sheep, mountain goat (Oreamnos americanus), bison (Bison bison), and collared peccary (Pecari tajuca). In addition, there are several species of free-ranging, non-native animals including feral horses (Equus caballus), feral burros (E. asinus), feral pigs (Sus scrofa), feral sheep (Ovis aries), and feral goats (Capra hircus) that occur on rangelands with mule deer. Based on literature examining interactions between deer and many of these species, it appears that species of greatest interest and potential to influence mule deer include American elk, white-tailed deer, and feral equids (horses and burros). There is also significant risk for future interactions with feral pigs if that species continues to expand and increase in abundance (O’Brien et al. 2019).

American elk have generally increased in population size across western North America in recent decades. Concurrent with this increase in abundance of elk has been a general decrease in abundance of mule deer in many areas. This concurrent and inverse relationship in abundance has led many to postulate that elk may be responsible for the decrease in abundance of deer. Both species occupy the same rangelands across much of western North America. However, results from many studies examining the potential for competition between mule deer and elk are inconsistent. Because of their smaller body size, deer require higher-quality forage than elk (Wickstrom et al. 1984). Elk are considered more generalist foragers and their diet is typically comprised of more low-quality food such as grasses, except during spring when diet is more similar—likely due to both species needing higher quality forage to recover from winter and reproduce (Stewart et al. 2003; Sandoval et al. 2005; Torstenson et al. 2006). In addition, elk are much larger than deer and appear to be socially dominant and capable of physically displacing deer, which provides some evidence for interference competition between these two species (Stewart et al. 2002). However, the evidence for displacement is not universal (Sallee et al. 2023).

White-tailed deer are the most widespread ungulate in North America and co-occur with mule deer across much of western North America. Similar to patterns observed with elk over recent decades, there appears to be an overall decline in abundance of mule deer that is concurrent with an increase in abundance of white-tailed deer. Among sympatric populations, mule deer typically demonstrated population decline concurrent with population increase by white-tailed deer (Robinson et al. 2002) even though mule deer appear to be competitively dominant (Anthony and Smith 1977). Genomic analyses suggest this pattern has a much older origin with effective population size for white-tailed deer increasing over the last 500,000 years while that same metric has declined for mule deer (Lamb et al. 2021). Because of the similarity in body size and digestive systems between both species, there is potential for forage competition. Indeed, studies have generally demonstrated a high degree of dietary overlap between these two species, but there is evidence for partitioning of food resources (Berry et al. 2019). Mule deer, for example, were better able to process forage containing the monoterpene α-pinene compared to white-tailed deer which may broaden forage options relative to white-tailed deer (Staudenmaier et al. 2021).

Mule deer also coexist on rangelands with feral equids in large portions of western North America (Stoner et al. 2021). Over 131,000 km² of mule deer habitat in the western United States is occupied by feral equids and 97% of management units for feral equids contain mule deer (Stoner et al. 2021). Moreover, over 80% of federally managed herds of feral equids exceed population objectives (BLM 2018). Overabundance of feral equids has been associated with habitat degradation and loss of biodiversity in some areas where mule deer occur (Zeigenfuss et al. 2014; Davies and Boyd 2019). Dietary overlap between feral equids and mule deer is limited in most seasons, but varies regionally and seasonally (Scasta et al. 2016). In a Sonoran Desert rangeland where both feral asses and mule deer consumed primarily browse species, overlap was highest and biologically significant during periods of abundant forage (summer and early fall seasons) compared to periods of relatively low forage availability when each species focused on forage that maximized physiological differences (Marshal et al. 2012). Feral horses appear to interfere with mule deer access to drinking water when limited availability of water occurs on arid and semi-arid landscapes (Hall et al. 2018). Conversely, feral burros may improve access to drinking water for mule deer on some rangelands in the southern portion of their range by digging wells in dry washes (Lundgren et al. 2021). Because of the ongoing range expansion and increase in abundance of feral equids, there is considerable potential for competition.

4.3 Fire

Historically, fire played a large role in structuring plant communities on rangelands where mule deer occur (Block et al. 2016). Historical fire-return intervals vary across rangelands in western North America from every few years to more than 300 years (Rollins 2009; Stevens et al. 2020). Tremendous variation (10–200 years) in return intervals can even occur within the same rangeland type (Miller and Tausch 2001). Variation in extent, severity, and return intervals associated with fire would have created a shifting mosaic on rangelands with associated plant communities in various stages of succession. This historical heterogeneity, however, has been reduced since Euro-American settlement. Suppression of fires coupled with changes in rangelands due to climate change and plant invasions (e.g., annual grasses) has led to increased size and severity of wildfires and an overall reduction in rangeland heterogeneity where mule deer occur (Dennison et al. 2014; Jolly et al. 2015).

The influence of fire (both prescribed and wild) on mule deer populations depends on the responses of rangeland plant communities to this disturbance which can be both positive and negative or both (Block et al. 2016). Both above-ground biomass and nutrients in plants can increase following fire (Rau et al. 2008; Roerick et al. 2019). Prescribed fire, for example, increased both crude protein and in vitro digestible organic matter in plants available to mule deer (Hobbs and Spowart 1984). Likewise, nitrogen in forbs and grasses regularly consumed by mule deer also increased following fire (Rau et al. 2008).

Changes in availability of plants and nutrition within plants following fire can influence habitat selection by mule deer (Fig. 17.3). On southern rangelands, mule deer selected for burned habitats unless they had been impacted recently (< 5 years) by high-severity fire and then they were avoided (Roerick et al. 2019; Bristow et al. 2020). Increased availability of nutritious plants on rangelands following fire should lead to increased health of adult females with cascading effects on reproduction including litter size and birthweight of neonates (Shallow et al. 2015). Indeed, recruitment rates over 13 years were positively correlated with acreage burned when precipitation patterns were also favorable (Holl and Bleich 2010). Selection for rangelands following fire, however, does not always translate into population growth (Klinger et al. 1989).

Fig. 17.3

Habitat selection during winter (3rd-order selection or habitat patches within a home range) for an adult female mule deer (Odocoileus hemionus) in relation to fire with the black polygon representing the burn boundary for a portion of the Pole Creek fire in central Utah which occurred during fall of 2018. Figure shows preference (warm colors) for edges and avoidance of areas on the interior of the fire polygon during the initial two winters following the fire. Red to green colors represent high, medium–high, medium, medium–low, and low probabilities of selection, respectively

Moreover, fire has the potential to negatively impact deer populations when rangeland plant communities respond negatively. Where cheatgrass or red brome (Bromus rubens) occur in combination with nonsprouting shrubs such as sagebrush (Artemisia spp.), for example, fire can alter rangelands in a way that is detrimental to deer. When these invasive grasses become dominant, they can create a negative feedback loop that leads to increased frequency of fires and eventual elimination of shrubs (Pilliod et al. 2017). Conversion of shrublands to grasslands across large acreages of western North America has been facilitated by invasive annual grasses and fire (D'Antonio and Vitousek 1992). These changes have the potential to negatively impact mule deer populations. Consequently, managers should carefully consider potential responses of rangeland plant communities prior to use of prescribed fire (Block et al. 2016).

4.4 Vegetation Management—Chaining and Mastication of Conifers

Rangeland managers have a long history of vegetation treatment to improve rangelands using a variety of methods. Mechanical methods such as chaining, lop and scatter, mastication (also sometimes referred to as shredding), and mowing have been used across the range where mule deer occur. Likewise, treatment of vegetation with herbicide has also occurred, often in conjunction with post-fire restoration or mechanical treatment. Each of these methods is designed to improve rangelands by reducing risk of catastrophic fire, increasing forage for domestic livestock, improving habitat for wildlife including mule deer, and general promotion of rangeland health.

Chaining and mastication are most frequently used to reduce cover of pinyon (Pinus spp.)-juniper (Juniperus spp.) and promote forbs, grasses, and shrubs in this cover type (Monaco and Gunnell 2020). Unlike chaining where woody debris is moved into piles, mastication allows rangeland managers to turn woody material into mulch and spread it onto the soil where it can facilitate positive responses from the plant community (Bybee et al. 2016; Havrilla et al. 2017; Monaco and Gunnell 2020). In recent years, efforts to reduce conifers from sagebrush rangelands have matched estimates of expansion (1.5% per year) for this cover type (Sankey and Germino 2008; Reinhart et al. 2020).

Mule deer respond to conifer removal positively when forage plants respond well and adequate concealment and thermal cover remains within or adjacent to treatment areas. Positive outcomes for mule deer were noted in some treated areas, whereas others showed no increase or even declines depending on response of plant communities and presence of adequate cover (Bombaci and Pejchar 2016). Because reduced use of chained habitats by mule deer has been noted beyond 120 m from cover, suggested guidelines include interspersing food and cover such that areas where chaining or mastication has occurred are no more than 200 m from cover (Fairchild 1999). Managers also need to provide mule deer with a diversity of forage options for consumption across seasons. Mule deer selected for mastication treatments in New Mexico during summer 1–4 years after treatment, but switched to patches > 4 years old in winter, presumably due to preference for herbaceous plants in the summer and browse species in the winter (Sorensen et al. 2020).

When deer populations are resource limited, vegetation treatments including mastication can increase population growth rates if plant communities respond favorably. Over-winter survival of 6-month old fawns in pinyon-juniper habitats treated with mastication and herbicide to control invasive grasses averaged 77% (SE = 8%) compared to 68% (SE = 11%) for areas without treatment or with treatment and no control of invasive grasses (Bergman et al. 2014a). Likewise, measures of condition for adult females (ingesta free body fat; Cook et al. 2007, 2010) were higher in areas with these same vegetation treatments (Bergman et al. 2014b). Similarly, mechanical removal of pinyon-juniper trees on a rangeland composed of perennial grasses interspersed with pinyon-juniper stands and limited shrublands influenced space-use patterns (smaller) and metrics of condition (greater) suggesting treatments improved forage for mule deer (Bender et al. 2013).

Conversely, many studies involving mule deer response to reduction of pinyon-juniper woodlands on rangelands have failed to identify a positive response (Bombaci and Pejchar 2016). Evidence that mule deer selected for juniper trees on sagebrush rangelands in the northwestern United States sparked a lively debate about the value of mastication projects for this species (Coe et al. 2018; Clark et al. 2019; Maestas et al. 2019). Coe et al. (2018) found selection for trees by mule deer at multiple spatial and temporal scales in a sagebrush-dominated rangeland and concluded that mastication of western juniper (J. occidentalis) may not improve habitat for this species. Maestas et al. (2019) countered that habitat selection was imprecise and unreliable, and response of mule deer to mastication could only be evaluated by looking at demographic responses (see Clark et al. 2019 for a rebuttal).

4.5 Vegetation Management—Mowing of Shrubs

Managers also treat rangelands dominated by shrubs where deer occur with mechanical implements such as a mower or Lawson aerator. Treatment of shrubs is often intended to reduce shrub cover in favor of herbaceous plants and create increased availability of forage for livestock and wildlife. Shrub response to vegetation treatments varies with species and environmental conditions. Positive outcomes have more potential in mountain big sagebrush (A. tridentata vaseyana) communities but have rarely been shown in Wyoming big sagebrush (A. tridentata wyomingensis) leading to calls for caution and more research (Beck et al. 2012). Consideration of seasonal use of rangelands by mule deer should also be considered in relation to treatment of shrubs. Mowing alters the structure and height of sagebrush with effects that can persist for more than 20 years (Davies et al. 2009). Consequently, mowing could have detrimental effects for mule deer if done on winter ranges where treated shrubs could then be covered in snow and unavailable (Davies et al. 2009).

4.6 Vegetation Treatment—Herbicide

Herbicides such as imazapic or Plateau^® (BASF, Ludwigshafen, Germany), indaziflam or Rejuvra^® (Bayer, Cary, NC, USA), and tebuthiuron or Spike^® (Dow AgroSciences, Indianapolis, IN, USA) are also used for vegetation treatment where deer occur. Tebuthiuron has been used to thin browse species and can be applied aerially resulting in highly variable mortality rates for shrubs (Scifres et al. 1979). Imazapic and indaziflam are specific herbicides developed for control of annual grasses (Mealor et al. 2013). These two herbicides can reduce biomass of undesirable annual plants such as cheatgrass, medusahead (Taeniatherum caput-medusae), or red brome and they are often used in combination with mechanical methods to restore rangelands (Elseroad and Rudd 2011; Burnett and Mealor 2015).

Responses of mule deer to vegetation treatments with herbicide are variable and dependent on the response of preferred forage species. Treatment of sagebrush with Tebuthiuron resulted in greater crude protein in leaves compared to plants in both control plots and sagebrush that was mowed during the initial year following treatment, but increases were modest (Smith et al. 2022). These treatments may have improved palatability of sagebrush for mule deer because plant secondary metabolites (e.g., terpenes) were unchanged in relation to treatment while crude protein increased, but marginal increases likely do not compensate for loss of cover or density of plants in sagebrush systems (Smith et al. 2022). Herbicide treatments in forested habitats were associated with reduced bite sizes and reduced digestible energy, but increased digestible protein for black-tailed deer (Ulappa et al. 2020). Similarly, mule deer in the Great Plains selected for sites treated with tebuthiuron five years earlier, ostensibly due to improved quality of forage (Gage 2011). Mastication of pinyon-juniper followed by control of annual grasses with imazapic was associated with increased condition of adult female mule deer and increased survival of mule deer fawns compared to untreated areas and areas treated with mastication alone where invasive grasses were not controlled (Bergman et al. 2014a, b).

4.7 Water Development

Water is an essential element for all life on earth including deer. Water, however, is available in forms other than drinking water and this species does not always need to drink. Three forms of water are recognized including metabolic water, preformed water, and free or drinking water. Metabolic water is produced when compounds such as carbohydrates, fats, and proteins are oxidized (Robbins 1983). A single gram of carbohydrate, for example, produces over half (0.56 g) a gram of metabolic water and the conversion ratio is essentially one to one for fatty compounds (Gill 1994). Deer also consume preformed water which is available in plants. Average moisture content of plants consumed by mule deer can vary from < 10% for some seeds and senescent grasses to > 80% for forbs and succulent plants (Cain et al. 2008). When metabolic and preformed water are adequate to meet the needs of deer, they do not need to drink. When metabolic and preformed water are inadequate to meet physiological needs, mule deer access free or drinking water from natural sources such as springs or streams and anthropogenic sources such as water developments (Larsen et al. 2012).

Although some have called for more research on the influence of water developments on deer populations (Simpson et al. 2011), water development has the potential to benefit deer—particularly on arid rangelands. On arid rangelands, mule deer regularly use water developments including troughs and wells (Fig. 17.4; Krausman 2002). Mule deer visited water sources every 1–4 days and consumed between 1 and 6 L of water per visit with higher consumption rates occurring during the summer months (Hazam and Krausman 1988; Shields et al. 2012). Frequency of visits to water sources was higher for females compared to males—particularly during summer when females were lactating (Fig. 17.5; Hervert and Krausman 1986; Shields et al. 2012).

Fig. 17.4

Image of a male mule deer (Odocoileus hemionus) at a livestock water development designed as a rangeland improvement practice to increase availability and distribution of free or drinking water

Fig. 17.5

Image of a female mule deer (Odocoileus hemionus) with two neonates at a wildlife water development designed to increase density and distribution of this species. Note the fence posts, but lack of wire as it has been removed to facilitate access by this species

During summer months, mule deer in arid rangelands are often located closer (typically within 5 km) to sources of water than they are during other seasons (Ordway and Krausman 1986; Krausman and Etchberger 1995). Moreover, availability of water was a factor associated with migration in arid environments; mule deer migrated to areas with available water during the summer months (Rautenstrauch and Krausman 1989). During summer months, availability of water was also associated with reduced movements in arid rangelands suggesting that mule deer were able to meet their resource needs in smaller areas when water was available (McKee et al. 2015). Access to drinking water may allow mule deer in arid rangelands to consume a wider variety of forage plants including species with low pre-formed water content. Moreover, higher densities of mule deer in the arid Chihuahuan Desert of Mexico were noted near available drinking water, although those densities may reflect habitat selection as opposed to increased abundance (Sánchez-Rojas and Gallina 2000a, b). Nonetheless, many of the arid rangelands where mule deer occur are considered water-limited (Cox et al. 2009).

Water developments may receive very little use by mule deer in arid rangelands—particularly if they are newly constructed, fenced, or heavily used by livestock or feral equids. Water developments available for longer than 3 years received more use than those built more recently suggesting mule deer required some time to find and acclimate to newly constructed developments (Marshal et al. 2006a, b). Fencing that successfully deterred feral horses from water developments was also associated with reduced use by mule deer—particularly when the area fenced around water sources was relatively small (Larsen et al. 2011). When water developments are not fenced, feral equids may outcompete mule deer and other wildlife for access to drinking water (Fig. 17.6; Hall et al. 2016, 2018). Competition for water can be particularly acute at relatively small water sources where feral horses can monopolize access for most of a 24 h period (Hall et al. 2018).

Fig. 17.6

Feral horses (Equus caballus) at a livestock water development. Feral horses have been shown to limit access to water for mule deer (Odocoileus hemionus) and other species at water sources

Water developments have also been used to mitigate loss of water resources due to anthropogenic effects (e.g., urban, agricultural, transportation, industrial development; Rosenstock et al. 1999; Krausman et al. 2006). In Joshua Tree National Park, both the number of available springs and the volume of water flowing from those springs have declined over the last 50 years (Longshore et al. 2009). Wildlife water developments, however, were able to partially offset the predicted loss of suitable habitat for bighorn sheep over those same years (Longshore et al. 2009). Projected loss of naturally occurring water sources will make water development an increasingly important rangeland management practice for mule deer (Seager et al. 2013). To maximize value to mule deer in arid rangelands, water developments should be spaced at 3.2–4.8 km from other sources of water (Heffelfinger 2006; Krausman et al. 2006).

4.8 Predator Management

Mule deer are vulnerable to a suite of predators and predation is typically the most common identified cause of mortality across all age and sex classes, except during severe winters and in the presence of high harvest by hunters (Linnell et al. 1995; Unsworth et al. 1999; Bishop et al. 2005b; Forrester and Wittmer 2013). The relationship between deer and their predators, however, is complex and nuanced with predation rates that vary across age and sex categories, in relation to animal condition, with availability of alternative prey for predators, in response to habitat conditions (e.g., availability of hiding cover), and by type of predator. Coyote predation on neonates during the summer, for example, was lowest when abundance of microtine rodents was high (Hamlin et al. 1984). Predation can also be considered compensatory or additive depending on where deer populations are in relation to carrying capacity and condition of rangelands (Ballard et al. 2001; Forrester and Wittmer 2013).

Given the complex ecological relationships between deer and their predators, it isn’t surprising that the results of predator control on population growth is highly variable and the literature equivocal. Removal of coyotes over a large area (> 10,500 km²), for example, was not associated with increased production of mule deer as measured by fawn/adult female ratios (Brown and Conover 2011). Likewise, although removal of coyotes and mountain lions showed some short-term increases in survival rates depending on age class, it did not appreciably change the long-term dynamics for a population of mule deer on rangelands in southeastern Idaho (Hurley et al. 2011). Conversely, removal of wolves on Vancouver Island led to increased survival, higher production, and positive population growth rates for black-tailed deer (Hatter and Janz 1994). Similarly, removal of coyotes was associated with increased survival of neonates and population growth when that predator was removed over consecutive years from fawning habitat when the deer population had room to grow (McMillan et al.2023).

Variation in outcomes associated with deer populations (along with ungulates in general) in response to predator removal highlight uncertainty surrounding this management action, the need for better science, and the requisite nature of information on limiting factors affecting populations targeted for increase following predator removal (Clark and Hebblewhite 2020). Conditions under which predator control was effective at influencing population growth in deer include deer populations below carrying capacity, predation identified as limiting and additive, control efforts adequate to significantly reduce predator densities for the species exerting “top-down” control, and control efforts conducted at optimal spatial and temporal scales (Ballard et al. 2001; Forrester and Wittmer 2013; Mahoney et al. 2018). When these conditions are not present, reductions in predators are unlikely to influence population growth.

5 Impacts of Disease

Similar to other species, deer are susceptible to numerous pathogens including bacteria, fungi, parasites, prions, and viruses. Diseases associated with a bacterial vector include rain rot (often associated with ticks and flies), necrobacillosis, gangrene, keratoconjunctivitisrosis, and others (Mule Deer Working Group 2014). Many different parasites including bot flies, fleas, lice, round worms, tapeworms, ticks, and others also occur in mule deer (Mule Deer Working Group 2014). Chronic wasting disease (CWD) is an emerging disease in mule deer caused by an infectious prion (modified protein). This disease is similar to bovine spongiform encephalopathy (sometimes referred to as “mad cow disease”) in cattle and scrapie in domestic sheep and is spreading rapidly in mule deer populations (Haley and Hoover 2015). Diseases caused by viral pathogens include hemorraghic diseases (blue tongue and epizootic hemorraghic disease or EHD), and fibroma tumors caused by the papilloma virus (Mule Deer Working Group 2014). Up to 40% of sampled white-tailed deer were found to have antibodies for Covid-19 and mule deer are likely also susceptible (Chandler et al. 2021).

5.1 Impacts of Disease on Populations

Most of the diseases listed above affect individuals with little influence on populations. Exceptions with potential to influence population growth include the hemmoraghic diseases and chronic wasting disease. Hemmoraghic diseases result from a viral infection spread by an insect vector (several species of Culicoides midges). Infection usually occurs in late summer or early fall and can result in significant mortality—particularly at northern latitudes (Howerth et al. 2001). In one outbreak on rangelands in California, over 1,000 mule deer were estimated to have perished with pathology for a sample of these animals consistent with bluetongue or EHD (Woods et al. 1996).

Impacts to mule deer populations have also been observed with CWD which is a relatively new disease currently found in mule deer and spreading rapidly (Haley and Hoover 2015). Chronic wasting disease presents as degenerative because it affects the central nervous system of host animals. The source of CWD in mule deer is not completely understood, however, it seems likely it may have originated in north-central Colorado and southeastern Wyoming because original diagnosis in captive herds occurred there in the late 1970s (Williams et al. 2002). Prevalence of CWD in cervids has increased exponentially over the last 6 decades and it has now been detected in 4 Canadian provinces and 29 US states including many where mule deer occur (Otero et al. 2021).

Although no differences in susceptibility to this disease have been identified between female and male mule deer, prevalence of CWD is higher for males and increases with age, ostensibly due to behavioral differences and higher contact rates for mature males in this species (Miller and Conner 2005). In areas with high prevalence rates (> 20%), population-level impacts have been noted (DeVivo et al. 2017). Mule deer infected with CWD are more susceptible to predation and vehicle strikes and large differences in annual survival (e.g., 32% compared to 76%) and population growth rate (e.g., λ = 0.79 compared to λ = 1.00) of CWD positive versus CWD-negative animals has been noted (Miller et al. 2008; DeVivo et al. 2017).

5.2 Disease Interactions with Livestock

Livestock including cattle and domestic sheep can be infected with hemorraghic diseases, but the disease is usually subclinical for these species (Howerth et al. 2001). The Culicoides midges that spread the virus associated with hemmoraghic diseases reproduce in water and some species (e.g., C. sonorensis) appear to thrive in stale waters or mud enriched by fecal material and urine from domestic livestock or wild animals (Pfannenstiel et al. 2015). Congregation of both livestock and wildlife including mule deer at water sources in the summer may help create favorable conditions for midges and spread of this disease (Pfannenstiel et al. 2015). Consequently management of livestock (e.g., stocking rate, group sizes) and water resources (e.g., protecting spring heads) may provide an option to reduce outbreaks of hemmoraghic diseases, but the etiology of the diseases are not completely understood and specific recommendations are not available (Pfannenstiel et al. 2015).

The prions associated with CWD can persist in the environment for years or even decades and contraction of the disease by mule deer from the environment is documented and perhaps more frequent than transmission between animals (Miller et al. 2004, 2006). Although natural transmission of CWD from mule deer to livestock has not been observed, passage of CWD to sheep can be induced via intracranial inoculation suggesting some potential to cross species barriers to domestic livestock (Cassman et al. 2021). Nonetheless, understanding of CWD and the potential for interactions with livestock are limited. Monitoring programs for CWD can be found in most states and provincial agencies. These programs often include collection of samples from harvested animals and are typically “hunter-based” (Smolko et al. 2021). Diagnostic tests for CWD have evolved and improved rapidly over the last few decades, but tests with high accuracy remain relegated to those where tissues (e.g., tonsil, lymph nodes) are collected postmortem and limitations persist for effective testing of live animals (Haley and Richt 2017). Despite several decades of research on CWD, much remains to be learned about the etiology of this disease and how to successfully manage it in mule deer populations on rangelands in western North America.

6 Ecosystem Threats

Despite a broad distribution (Fig. 17.1) and relatively high abundance, mule deer populations face many threats. The human population in western North America has grown rapidly over the last century with concomitant changes to landscapes including habitat loss, degradation, and fragmentation. In 2008, the estimated human footprint (physical area occupied by humans as housing, roads, intensively managed agricultural lands, and other infrastructure) covered 13% (402,000 km²) of the land area in the Western United States (Leu et al. 2008). The size of this footprint has increased rapidly in recent decades. Between 1980 and 2010, for example, residential land-use increased by 37% and impacted more than 1,000,000 ha of western Colorado (Johnson et al. 2017). Likewise, the estimated land area occupied by energy extraction infrastructure (wells, well pads, roads, storage facilities) in the central portion of North America including many areas where mule deer occur on the eastern portion of their range increased by 3 million ha between 2000 and 2012 leading to an estimated loss of 10 Tg of plant biomass (Allred et al. 2015). Changes in land use and increases in habitat loss in western Colorado were associated with an average decrease of 0.5 fawns per 100 adult females per year with a strong negative trend noted in relation to increased residential development (Johnson et al. 2017).

Direct loss of habitat can reduce availability of forage for deer, but these effects are often magnified by avoidance of areas near human structures creating indirect effects. Mule deer, for example, avoided areas within 2.7 and 3.7 km of well pads during winter suggesting that indirect effects of energy extraction were greater than direct effects (Sawyer et al. 2010). Avoidance of areas near roads and well pads magnifies the impact of surface disturbance with recent estimates suggesting a multiplier of 4.6 should be applied for this species to account for indirect effects associated with avoidance of energy infrastructure (Dwinnell et al. 2019). Avoidance of wells and well pads was most pronounced during the active drilling phase when presence of humans, vehicles, noise, and artificial light was greatest (Northrup et al. 2021). Nonetheless, even after 15 years this species failed to habituate to energy infrastructure in some areas suggesting effects can be long-term and persistent (Sawyer et al. 2017).

Moreover, fragmentation of landscapes used by mule deer can disrupt migration routes and timing (Lendrum et al. 2013). Avoidance of human structures when selecting stopover areas, for example, has been documented for migratory mule deer (Wyckoff et al. 2018). Additionally, decreased movement efficiency and increased energy expenditure were noted for mule deer while migrating through areas impacted by surface mining potentially leading to fitness consequences (Blum et al. 2015). Likewise, there can be genetic consequences of habitat fragmentation evidenced by genetic structure that corresponded to highway boundaries (Fraser et al. 2019). Impacts including avoidance of areas disturbed by energy development have been observed for migratory mule deer after surface disturbance reached only 3% of rangelands (Sawyer et al. 2020; Lambert et al. 2022). Disruptions to migratory routes and reductions in fitness for deer migrating across disturbed and fragmented landscapes could lead to complete loss of migratory knowledge by populations if information on routes and timing is transmitted culturally in this species as it is with other ungulates (Jesmer et al. 2018).

Additional indirect effects associated with growth in human populations in western North America include altered dynamics between deer and their predators and responses to increased recreation by humans on rangelands. Interactions between mule deer and mountain lions, for example, were influenced by urbanization and presence of artificial light (Benson et al. 2016; Ditmer et al. 2021). These altered interactions may trigger trophic cascades that lead to changes in plant composition on rangelands (Waser et al. 2014). Additionally, hikers and in particular—hikers off established trails with dogs—have been shown to influence mule deer with increased vigilance and energy expenditure common responses (Miller et al. 2001). Likewise, approximately 1/3rd of mule deer with GPS collars responded to people searching for and collecting shed antlers by leaving established home ranges which increases energy expenditure and may lead to increased predation risk (Bates et al. 2021).

6.1 Climate Change

Deer demonstrate a wide thermal tolerance zone which is reflected in their broad distribution (Wallmo 1981). The thermo-neutral zone for mule deer has been estimated at operative temperatures (temperatures experienced by the animal after accounting for wind and radiation) between − 20 °C and 5 °C (− 4 °F to 41 °F) in winter and < 25 °C (77 °F) in summer (Parker and Robbins 1994). Outside of these operative temperatures, mule deer must alter behavior (i.e., seek shade) or use water and energy to maintain homeostasis (Parker and Robbins 1984). As temperatures warm under predicted climate change scenarios, deer may be forced to adjust behavior and alter resource selection. Because operative temperatures experienced by deer are strongly influenced by cover, changes to rangelands (e.g., conversion from shrubland to grasslands due to increased fire frequency) that result in reduced cover may exacerbate the effect of increased temperatures (Parker and Gillingham 1990). Consequently, the relative role of thermal cover and water to resource selection, for example, may increase for this species and rangelands with limited thermal cover may require mule deer to expend more energy to meet their needs, which may influence fitness.

Nonetheless, the biggest impacts to this species from climate change will likely be indirect. Massive conifer mortality across much of western North America, for example, is predicted due to increased temperatures (McDowell et al. 2016). Increased temperatures will likely result in more frequent and larger wildfires (Schoennagel et al. 2017). Increased frequency and severity of wildfires, in turn, will lead to loss of forests and shrublands in favor of grasslands—particularly where annual grasses such as cheatgrass have invaded. Once established, annual grasses such as cheatgrass prolong the fire season and increase availability of fine fuels in a negative feedback loop that leads to more frequent fire that furthers conversion of shrublands and forests to rangelands dominated by annual grasses. Conversion of rangelands dominated by forests and shrubs to grass-dominated systems will have far-reaching and cascading consequences for mule deer. These changes are likely to favor more generalist ungulates (e.g., elk) which may exacerbate potential competition with mule deer.

Moreover, long-standing benefits associated with migration for mule deer including prolonged access to nutritious forage for migrating animals are likely to be reduced or eliminated (Aikens et al. 2020). Massive losses in surface water are also predicted for western North America which will reduce availability of free or drinking water for this species and likely increase competition at remaining water sources where mule deer are at a competitive disadvantage to feral horses (Hall et al. 2018). Increased aridity and drought on rangelands in western North America may also influence disease dynamics associated with hemorrhagic diseases as outbreaks typically occur in summer and are often most severe during drought years. Increased temperatures may congregate animals at water sources and favor conditions for the midges associated with transmission of this disease. Indirect effects including ecosystem change associated with climate change represent a serious threat to mule deer populations across their range.

7 Conservation and Management Actions

Deer have a long history of management by provincial and state agencies. Agencies typically develop conservation plans to guide management decisions with spatial resolution for specific plans often at the population level (i.e., individual conservation or management plans for specific herd units). Conservation and management planning, however, can be challenged by migratory mule deer that regularly cross administrative boundaries (Middleton et al. 2019). In many jurisdictions, habitat restoration efforts are also coordinated between provincial or state management agencies and land owners. Mule deer are of tremendous interest to the public and many non-governmental organizations work closely with provincial, state, tribal, and federal agencies to conserve and manage this species. Most state and provincial agencies have active programs to conserve and restore habitat for mule deer.

In recent decades, many of these programs have focused on preservation and restoration of winter ranges. The working hypothesis for many mule deer populations is that they are nutritionally limited and many in particular across the northern portion of their distribution are thought to be limited by quality of winter range (Bergman et al. 2015). Consequently, millions of dollars have been spent to conserve and restore rangelands used by mule deer during winter. Likewise, large investments in restoration of natural springs and construction or maintenance of wildlife water developments for mule deer have occurred in more arid portions of their range. Total expenditures for development and maintenance of wildlife water developments across western states in the US exceeded $1,000,000 in 1999 (Rosenstock et al. 1999). Recent work has also highlighted the importance of maternal condition on population growth suggesting that populations may also be limited by quality of summer range (Lamb et al. 2023).

Most recently, advances with tracking technology including GPS collars have allowed for unprecedented understanding of migration ecology for this species including the ability to identify migration routes, stopover areas, and barriers to migration. State and provincial agencies have been able to use this information to construct highway crossing structures (e.g., underpasses and overpasses) which have successfully facilitated crossing of major roadways (Sawyer et al. 2012). Likewise, information derived from GPS collars has been valuable for conservation and management of populations that use rangelands on multiple land ownerships throughout the year (Middleton et al. 2019).

Mule deer are of tremendous interest to the public and many non-governmental organizations (e.g., Mule Deer Foundation). These conservation organizations provide resources including funding, equipment, and volunteers that can be leveraged with state or federal resources to complete more conservation projects at the larger scales needed to benefit this species. Engagement with the public and these organizations including local stake holders is crucial to sustain conservation actions (Middleton et al. 2019). Furthermore, production and sharing of information with the public in a format that is easily accessible to a diverse group of constituents and stakeholders can facilitate conservation planning and implementation of management actions that benefit mule deer (Middleton et al. 2019). Although the challenges to long-term conservation of mule deer are immense, so is the interest amongst the public associated with this species.

8 Research/Management Needs

Despite a long history of management and research on mule deer in North America, much remains to be learned about this species. Tremendous variation in habitat selection and migratory behavior exists across populations of mule deer and thus information at relatively fine spatial scales is often needed for effective conservation and management actions—particularly in the face of rapid growth of human populations in western North America. Information on the response of mule deer to landscape changes currently influencing rangelands is urgently needed for conservation and management of this species. Recent research on response of mule deer to oil and gas development, for example, suggests impacts when as little as 3% of the landscape is disturbed (Sawyer et al. 2020), but no information is available regarding the applicability of that threshold to renewable energy. Thus, information on response of mule deer to solar or wind energy developments would help with conservation planning for this species. Likewise, we lack clear understanding of the impacts to mule deer from recreational activities occurring on rangelands. Increased vigilance and some displacement of individuals has been noted (Miller et al. 2001; Bates et al. 2021), but these effects have not been tied to maternal condition, demographics, or population growth.

It is also clear that there is potential for the management of co-occuring species and the rangelands where deer occur to influence this species. There is a need to better understand the potential for apparent competition between co-occuring ungulates and mule deer. Apparent competition occurs when top-down effects on a population are mediated by a common predator. For example, mule deer are a preferred food source by a generalist predator (e.g., mountain lion, Cooley et al. 2008). If co-occurring species of ungulates such as elk or feral equids provide a supplemental food source such that mountain lions can maintain a higher population size or even maintain a consistent population size during times of declining abundance for mule deer, then there is potential for coexisting ungulates to have an indirect effect on mule deer. Indeed, there is some evidence for such an interaction between mule deer and white-tailed deer (Wielgus 2017), but no information on mule deer with elk or feral equids.

Additionally, aggressive measures to reduce prevalence of CWD have been implemented or proposed in many areas including altered hunt structures to impose increased harvest on mature males. Many of the proposed mitigation measures, however, are untested and deeply unpopular with consumptive users. Spread of CWD has increased exponentially since it was discovered and this disease shows little sign of abating (Otero et al. 2021). Chronic wasting disease has the potential to disrupt management of deer populations along with funding to provincial and state governments. Thus, there is an urgent need to identify measures to mitigate the spread of CWD.

Land managers across western North America are faced with increasing challenges to rangelands due to invasive plants, increased frequency and severity of wildfires, and climate change with all of the interacting and cascading effects. Improved understanding of how mule deer, and their habitats are likely to respond to these disturbances including climate change would help with conservation planning. More importantly, however, are solutions that can be employed on rangelands to mitigate and restore these ecosystems. Land managers would also benefit from help with prioritization so that resources can be maximized for conservation value. In recent decades, for example, substantial energy, effort, and money have been invested in restoration of winter ranges for mule deer, but it is unclear if those efforts represent the best use of limited resources. Some argument can be made that restoration of summer habitats may impact population growth of mule deer as much or more than restoration of winter habitat (Clements and Young 1997). Consequently land managers would benefit from increased understanding of the relative value of summer versus winter habitat quality on population growth in deer.