Abstract Woodland caribou ( Rangifer tarandus ) are one of the most iconic wildlife species in Canada. These large ungulates depend on undisturbed, interconnected assemblages of mature and old‐growth conifer forests for survival, and exhibit life‐history traits typical of K‐selected species (i.e., low reproductive rates, long generation times, and slow population growth rates). Many of their local populations are threatened by habitat alteration and loss, increases in predation rates, climate change, and the cumulative effects of these stressors. Three of Canada's Designatable Units for caribou, which represent irreplaceable components of its biodiversity, are listed as threatened or endangered under the Species at Risk Act owing to their rapid population declines. There are rising concerns about the feasibility of their recovery, as many caribou populations become increasingly small and isolated. While most of the discussion on caribou conservation has focused on the relationship between habitat, human disturbances, and ensuing changes to predation pressure, here we broaden the conversation by exploring the influences of evolutionary forces, ecological processes, and population‐level demographic and genetic parameters on the vulnerability and viability of critically small caribou populations. Our objective is to support the management and recovery of caribou in Canada through an in‐depth, consolidated assessment of scientific information on the conservation of small wildlife populations in general, and of small caribou populations in particular. We begin by introducing small caribou populations of concern in Canada, including their population status and ongoing recovery efforts. Next, we examine several factors broadly relevant to the persistence of any small wildlife population, namely population size, connectivity among populations, habitat suitability, and population response variability, and discuss how they apply to caribou. We discuss ways by which demographic and environmental stochasticity interact with population‐level demographic and genetic parameters, and how certain evolutionary forces and ecological interactions increase the vulnerability of small populations. We conclude by examining key but difficult questions surrounding the viability and recovery potential of small caribou populations, including relationships between population size, sustainability, and recovery potential, interventions available to help small populations, and what could be lost when small caribou populations become extirpated. We emphasize the importance Résuméof combining information about population size, connectivity, and habitat availability in determining probability of persistence of small populations, but also the urgency of acting even when this knowledge is incomplete. We underline the urgent need for concomitant interventions to address both proximate (e.g., increased predation pressure) and ultimate drivers of decline (e.g., habitat alteration and loss). Protection of large tracts of undisturbed habitat will be crucial for persistence over the long term, but immediate conservation actions are required to ensure that small caribou populations are able to survive until suitable habitat becomes available across the landscape.
Translocations have been widely used to restore and conserve bighorn sheep (Ovis canadensis) populations in North America. Some translocations have been successful, but many populations remain small and genetically isolated. Population structure can influence the viability and long-term success of reintroductions. Social ungulates often function as interconnected subpopulations (metapopulations); however, few studies evaluate subpopulation sizes, connectivity, and genetic diversity within metapopulations. To address this gap, we conducted a comprehensive study of a reintroduced Rocky Mountain bighorn sheep (Ovis canadensis canadensis) population in Dinosaur National Monument in Colorado and Utah, USA, between 2006-2020. We analyzed global positioning system (GPS) radio-collar data, genetic samples, and results of health testing to evaluate abundance, distribution, genetic structure and diversity, habitat use, movement and connectivity, and presence of or exposure to respiratory pathogens. We integrated these analyses to evaluate the outcomes of a reintroduction effort that began in 1952, over 70 years ago, and to inform management decisions in Dinosaur National Monument. We also provide a framework for evaluating metapopulation processes, including a non-invasive approach that links genetic structure with Bayesian spatial capture-recapture analyses to estimate subpopulation sizes. Despite models indicating continuous suitable habitat, we found a spatially structured population with at least 4 subpopulations with constrained connectivity. Evidence from step selection and density analyses suggested that movement among subpopulations may be limited by semi-permeable barriers including rivers and human disturbance, which could contribute to maintenance of spatial structure over time. In 2006, antibody to Mycoplasma ovipneumoniae was detected in all geographically and genetically distinct subpopulations. Widespread clinical signs of disease and a confirmed exposure to M. ovipneumoniae in 2019 indicate a long-term disease challenge. Proximity to domestic sheep creates repeated opportunities for introduction of new M. ovipneumoniae strains. We estimated abundance in 2019 at 109 (95% CrI = 87-133), composed of subpopulations ranging from 18-39 animals (95% CrIs from 11-50). Genetic diversity was relatively high compared to other reintroduced and native Rocky Mountain bighorn sheep populations, which is likely a consequence of multiple translocations from different sources. Three of 4 subpopulation centers generally aligned with the locations of original translocation release sites. Persistence in the presence of pathogens may be facilitated by metapopulation structure and moderately high genetic diversity. Conversely, metapopulation structure can also facilitate pathogen persistence. Our approach offers a path to advance understanding of the population ecology of reintroduced bighorn sheep and can inform effective conservation and management of their populations.
There has been global concern regarding the effects of selective harvest on the size of horns and antlers. Many government agencies have imposed regulations on hunters based on the size of antlers or horns to protect young males. This has caused phenotypic and genetic decline in the horn size of one bighorn sheep (Ovis canadensis) population and concern about the decline of antler size in white-tailed deer (Odocoileus virginianus). In Texas, USA, there has been interest in selective harvesting to increase genetic potential for antler size of white-tailed deer based on the results of several studies of penned deer. Designed studies on effects of selective harvest on horn or antler size in ungulates are rare, especially studies with a control area. We conducted 20 years of experiments on the effects of selective harvest on antler size of white-tailed deer using 2 research sites in South Texas, USA, each with a control area. Our first experiment was on the King Ranch in southeastern South Texas, where our objective was to see if we could change the phenotypic size of antlers in a population as one of the prerequisites for evolution. We picked a selective harvest area and a nearby area with only incidental recreational harvest as a control, each of about 3,800 ha. There were no barriers to deer movement on or off the study areas. We harvested the treatment area for 6 years (1999-2004) and monitored 1 year post-harvest (2005). For the selective harvest treatment, we used Texas legal methods for hunters and harvested yearling males (1.5 years of age) with <6 antler points and older males with <= 8 antler points. We monitored antler size by age class each year with pre-harvest captures in both areas using a helicopter and net gun. We harvested 135 males in the treatment area and an additional 33 were harvested by recreational hunters. Eight of the recreational harvests met our criteria. During the study, we caught and aged deer, measured antlers, implanted a microchip for recognition of recaptures, and released 436 and 303 males in the treatment and control areas, respectively. Selectively harvested males had smaller antlers than captured males for 1.5- and 3.5-4.5-year-olds but not other ages. Despite an average 8.6% annual harvest rate, antlers of all ages did not differ from the beginning to the end of the study. We concluded that insufficient harvest rate, immigration and emigration of yearling males, and rainfall variation affecting recruitment contributed to our results. We used results of the King Ranch experiment to guide design of follow-up experiments on the Comanche Ranch in southwest South Texas. We used 3 existing game-fenced areas to prevent or restrict deer exchange. A control area of 20 km(2) was completely high-fenced, as was an intensively harvested area of 14 km(2). A third area with a moderate selective harvest treatment encompassed 73 km(2) and was 85% enclosed by high fence. Deer on all study sites were provided with a pelleted nutritional supplement, which a concurrent study on Comanche Ranch showed prevented density-dependent population dynamics. The intensive treatment had harvest criteria similar to the King Ranch experiment, targeting 1.5-year-olds with <6 antler points and 3.5-4.5-year-olds with <= 8 antler points. Males 2.5 years old were harvested if <8 points, as were >= 5.5-year-olds with a gross Boone and Crockett score (GBC) <145. For the moderate harvest treatment, we released all 1.5- and 2. 5-year-olds and harvested older males with the same criteria as for the intensive treatment. To increase harvest rate, instead of using hunters, we harvested males at capture if they met selective harvest criteria. Males with antlers larger than the selective harvest criteria were released. We monitored antler size through yearly captures by helicopter and net gun in each area. In addition to measuring antler size, we aged and weighed captured males and inserted a microchip at the base of an ear of those that were released. We applied selective harvest treatments for 7 years (2006-2012) and continued capturing for 6 years post-harvest (2013-2018) to monitor for any lag effects of treatments. We estimated adult sex ratios using annual surveys by helicopter. Our initial objective, like the King Ranch study, was to see if harvest treatments increased the population antler size as one of the prerequisites for evolution. We captured 6,675 males over the 13 years of the study, including 3,346 recaptures. During the treatment period, we selectively harvested 365 males in the intensive treatment and 878 males in the moderate treatment. Harvested males of all age classes averaged smaller antlers than released males. We estimated a 19% yearly harvest rate in the moderate treatment (no harvest of 1.5- and 2.5-year-olds) and 48% of all ages in the intensive treatment. During the treatment period, adult sex ratio increased from 2:1 to 5:1 (female:male) in the intensive treatment but did not change in the moderate treatment, despite a male population size decrease of 65%. There was a large female harvest by Comanche Ranch during the harvest period that influenced sex ratio results in the moderate treatment. The male population decreased 88% during the harvest period in the intensive area. At the end of the harvest period (2013), antler size was larger compared to control in the 3.5-4.5-year (+10.67 GBC) and >= 5.5-year (+8.87 GBC) age classes in the intensive treatment and the >= 5.5-year (+3.70 GBC) age class in the moderate treatment. However, there was little evidence of phenotypic change and thus no detectable trait evolution by the end of the post-harvest period (2018), as all ages averaged smaller antlers or did not differ compared to the control. Additionally, only the 3.5-4.5 age class in the moderate treatment had larger antlers in 2018, versus the study beginning in 2006. Males changed state across years from eligible for harvest to the do-not-harvest category according to our criteria, and vice versa, likely as a result of stochastic rainfall. We used the experimental design, capture data, and ear tissue from the Comanche Ranch experiment to conduct 3 genetic studies. First, we analyzed male breeding success using 15 microsatellite loci. Our objective was to determine if male breeding success by age class changed in the treatment areas versus the control site during the treatment period (2006-2012). We genotyped 162, 432, and 1,504 captured males in the control, intensive, and moderate areas, respectively. Mature males (>= 5.5 years of age) comprised most of the sires in the control (62%) and intensive areas (42%). However, young males (1.5-2.5 years old) outnumbered mature males in the moderate treatment by 40% versus 28%, respectively. We assigned sires for 650 of 938 (69%) sampled offspring (1.5 years of age). Successful males sired 1-6 male offspring/year. During the post-harvest period (2013-2018), mature males comprised most of the candidate sires in the control (81%) and moderate (40%) areas. In contrast, young males slightly outnumbered mature males, 39% versus 34% in the intensive treatment. During post-harvest, successful males sired 1-4 male offspring that were recruited and averaged 1.1, 1.8, and 1.5 offspring/year for the control, intensive, and moderate areas, respectively. The objectives of our second genetic study were to 1) estimate heritability for antler traits, the proportion of variation in a trait within a population that is attributable to genetic differences rather than environmental factors, by age class, and 2) whether Comanche Ranch selection differentials were sufficient to cause evolution in antler size. We constructed pedigrees based on paternity assignments and then used an animal model procedure to estimate variance components and heritability of antler traits. Estimated heritability (h(2)) for antler points and GBC was low (0.06-0.16) for males 1.5-2.5 years of age. For males >= 3.5 years of age, heritability for antler points ranged from 0.25-0.32, and for GBC ranged from 0.31-0.68. Repeatability for antler points and GBC score ranged from 54-58%. Our annual estimates for generation time across treatments and the control area ranged from 3.8-7.3 years. Estimated response to selection was <= 0.2 for antler points for all ages and 0.8-5.8 GBC units for males >= 3.5 years of age. Given the long generation times and modest predicted responses, if there were genetic changes due to our selective harvest treatments, they appeared to be minor in these high-fenced populations. Our third genetic study was to estimate if selective harvest for antler size at Comanche Ranch was correlated with the distribution of major histocompatibility complex (MHC) DRB II genotypes in sires and offspring. The MHC DRB complex is associated with the immune system in animals. Previous research in Oklahoma, USA, has shown a relationship between MHC DRB and antler size in white-tailed deer. For this study, we used tissue from captured males of known age that were born in 2011 and 2012. No association between MHC DRB II alleles and antler development was detected. We speculated that the semi-arid climate of South Texas may result in fewer pathogens than the more mesic climate of previous research. We concluded that selective harvest to alter gene frequencies at the population scale to increase antler size in white-tailed deer, if possible, would require decades and would be unlikely in open populations with dispersal and no supplemental feed. Our results also showed that antler size restrictions on harvest by governmental agencies would be unlikely to cause evolution in white-tailed deer.
Wildfire has been a fundamental component of ecosystem dynamics for millennia. However, climate change, fire suppression, and land use practices have produced larger, more severe, and more frequent fires that threaten a range of wildlife species. Land managers in some fire-prone forest systems use fuels reduction treatments, including prescribed burning and mechanical thinning, to reduce vegetation density and the risk of large, high-severity fires. The strategy is to create forest landscapes that are more resilient to wildfire, which in turn will lessen the negative impact of uncharacteristically severe wildfire on wildlife. Yet there is uncertainty regarding the extent to which fire-adapted species are threatened by novel wildfire characteristics. The spotted owl (Strix occidentalis) has become a focal point for this controversy owing to its use of fire-suppressed forests that have been and will be vulnerable to changing fire regimes in the future. In this study, we attempted to address the question of how spotted owls respond to high-severity fire in California, USA, by leveraging 1) over 3 decades of detection-non-detection surveys in 1,514 historical spotted owl territories and 2) global positioning system (GPS) movement data from 171 individual spotted owls. Our study area included 7 national forests, 3 national parks, and extensive privately owned forests in California. Dynamic occupancy analysis indicated that spotted owl territories were more likely to become vacant and less likely to be colonized if they experienced recent and extensive high-severity fire. Consistent with the occupancy analysis, our movement analysis indicated that individual spotted owls avoided using severely burned forest. Although spotted owl territory persistence was higher in pyrodiverse territories (i.e., those with a mixture of low, moderate, and high severity burned areas), foraging owls showed mixed responses to local pyrodiversity, with some evidence that use of severely burned forest increased when pyrodiversity was high. We conclude that large and severe wildfires threaten the persistence of spotted owls in California. That said, forest restoration and fuels reduction treatments that reduce the risk of large and uncharacteristically severe fires while promoting multi-scale heterogeneity and pyrodiversity will likely benefit spotted owls and other species that occupy mixed conifer forests in California.
Highly pathogenic avian influenza viruses (HPAIV) have had disastrous, worldwide effects on wild birds and domestic poultry since the emergence of the A/goose/Guangdong/1/1996 (Gs/GD/96) lineage. The currently circulating H5N1 clade 2.3.4.4b has an expanded set of susceptible hosts, including many migratory wild birds, and is associated with higher transmission rates, increased susceptibility among wild bird hosts, and a greater number of wildlife reservoirs. Certain wild bird life-history strategies and behaviors have been suggested to explain avian hosts' susceptibility and exposure to HPAIV. These biological traits include gregariousness, such as colonial nesting and mixed flock foraging, predation or scavenging on wild birds, and association with aquatic habitats. Variation in host infection responses (e.g., infectability, shedding rates and duration, mortality rate, antibody development) informs the overall infection risk across avian species, yet the specific role of biological traits is often inconsistent and unclear across taxa. Moreover, the interactions and potential compounding effects among these biological traits remain largely unknown. To develop a more holistic understanding of cumulative risk across bird species, we integrate existing information on infection risk factors (i.e., susceptibility, immunological response, and behavioral traits) into a qualitative multivariate analysis. This approach enabled us to examine how infection risk factors relate to biological traits (e.g., phylogeny, physiology, behavior, species range) and to begin disentangling their complex interactions. We quantified and summarized these risk factors across host species and qualitatively ranked species by their viral responses along a proposed HPAIV response continuum, guided by expectations of traits and metrics associated with competence or vulnerability to HPAIV. In doing so, we aimed to better understand how viral responses and biological traits synergistically interact to influence cumulative risk across wild bird species. This work broadly expands on the previous avian influenza literature, which has focused on Anseriformes and Charadriiformes as primary viral reservoirs. We tie our findings to effective disease management responses with links to risk components, including descriptions of potential surveillance strategies applied to research and One Health goals, as well as a fuller understanding of how resources may be better deployed for rapid response when spillovers do inevitably occur. Additionally, we identified numerous areas where vital epidemiological information is lacking to best characterize the spread of these viruses. Ultimately, this improved understanding will help identify and inform disease management needs and decision making.
