Conservation Genetics https://doi.org/10.1007/s10592-018-1075-6
RESEARCH ARTICLE
Genetic diversity, effective population size, and structure among black bear populations in the Lower Mississippi Alluvial Valley, USA Sean M. Murphy1,2 · Jared S. Laufenberg3,4 · Joseph D. Clark5 · Maria Davidson1 · Jerrold L. Belant6 · David L. Garshelis7 Received: 28 November 2017 / Accepted: 23 May 2018 © Springer Science+Business Media B.V., part of Springer Nature 2018
Abstract Multiple small populations of American black bears Ursus americanus, including the recently delisted Louisiana black bear subspecies U. a. luteolus, occupy a fragmented landscape in the Lower Mississippi Alluvial Valley, USA (LMAV). Populations include bears native to the LMAV, bears translocated from Minnesota during the 1960s, and recently reintroduced and colonizing populations sourced from within the LMAV. We estimated population structure, gene flow, and genetic parameters important to conservation of small populations using genotypes at 23 microsatellite markers for 265 bears from seven populations. We inferred five genetic clusters corresponding to the following populations: White River and western Mississippi, Tensas River and Three Rivers, Upper Atchafalaya, Lower Atchafalaya, and Minnesota. Upper Atchafalaya was suggested as the product of Minnesota-sourced translocations, but those populations have since diverged, likely because of a founder effect followed by genetic drift and isolation. An admixture zone recently developed in northeastern Louisiana and western Mississippi between migrants from White River and Tensas River, resulting in a Wahlund effect. However, gene flow among most populations has been limited and considerable genetic differentiation accumulated (global FST = 0.22), particularly among the three Louisiana black bear populations that existed when federal listing occurred. Consistent with previous bottlenecks, founder effects, and persisting isolation, all LMAV bear populations had low genetic diversity (A R = 2.08–4.81; HE = 0.36–0.63) or small effective population size (NE = 3–49). Translocating bears among populations as part of a regional genetic restoration program may help improve genetic diversity and increase effective population sizes. Keywords Bottleneck · Founder effect · Genetic drift · Isolation · Louisiana black bear · Ursus americanus
Introduction
Electronic supplementary material The online version of this article (https://doi.org/10.1007/s10592-018-1075-6) contains supplementary material, which is available to authorized users. * Sean M. Murphy
[email protected] 1
Louisiana Department of Wildlife and Fisheries, 646 Cajundome Boulevard, Suite 127, Lafayette, LA 70506, USA
2
Present Address: Department of Forestry and Natural Resources, University of Kentucky, 214 T.P. Cooper Building, Lexington, KY 40546, USA
3
Department of Forestry, Wildlife and Fisheries, University of Tennessee, 274 Ellington Plant Sciences Building, Knoxville, TN 37996, USA
Habitat loss associated with human population expansion has severely fragmented many populations of large carnivores (Wolf and Ripple 2017). Consequently, multiple species of large carnivores are comprised of small, isolated 4
Present Address: U.S. Fish and Wildlife Service, 1011 East Tudor Road, Anchorage, AK 99503, USA
5
U.S. Geological Survey, Northern Rocky Mountain Science Center, Southern Appalachian Field Branch, University of Tennessee, 274 Ellington Plant Sciences Building, Knoxville, TN 37996, USA
6
Carnivore Ecology Laboratory, Forest and Wildlife Research Center, Mississippi State University, 251 Thompson Hall, Mississippi State, MS 39762, USA
7
Minnesota Department of Natural Resources, 1201 East Highway 2, Grand Rapids, MN 55744, USA
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populations that may have heightened vulnerability to stochastic genetic, demographic, and environmental processes (Lande 1993). Bottlenecks and genetic drift can erode genetic diversity and effective population size in small, isolated populations, resulting in reduced population fitness and adaptive potential (Jamieson and Allendorf 2012; Frankham 2015), which can substantially influence long-term population viability (Pierson et al. 2015; Whiteley et al. 2015; Benson et al. 2016). To mitigate deleterious genetic and demographic consequences of small population size and isolation, conservation planning typically focuses on habitat restoration to improve landscape connectivity and thus facilitate gene flow among populations (Dixon et al. 2006; Gilbert-Norton et al. 2010). Population reintroduction or augmentation can also be used to improve genetic diversity of imperiled large carnivores (Seddon et al. 2005; Hayward and Somers 2009; Johnson et al. 2010), but logistical and financial constraints often result in releases of small numbers of animals with a limited subset of alleles from source populations, which can cause harmful founder effects (Nei et al. 1975; Weise et al. 2014). Widespread habitat loss and fragmentation greatly reduced American black bear Ursus americanus range and caused local extirpation from many parts of the southeastern United States by the early to mid-twentieth century (Wooding et al. 1994). Local extirpation was particularly prevalent in the Lower Mississippi Alluvial Valley (LMAV; St. Amant 1959), a portion of which was inhabited by Louisiana black bears U. a. luteolus, one of 16 recognized American black bear subspecies. To counter the decline of Louisiana black bears, 161 bears were translocated from Minnesota and released in the Tensas River Basin (n = 31) and Upper Atchafalaya River Basin (n = 130) of Louisiana during the 1960s (Taylor 1971). Within the presumed historical range of the subspecies, bears were restricted to three small, isolated populations in the Tensas, Upper Atchafalaya, and Lower Atchafalaya River Basins by the 1980s, and were federally listed as Threatened in 1992 (U.S. Fish and Wildlife Service [USFWS] 1992). Recovery criteria were subsequently developed for Louisiana black bears that focused on demographic viability, habitat protection and restoration, and the creation of interconnecting habitat corridors between populations (USFWS 1995). Based on an approximately fivefold increase in suitable bear habitat, results from analyses that suggested high demographic viability of the three populations that existed when listing occurred (> 0.99 over 100 years; Laufenberg et al. 2016), and radio-telemetry monitoring data that supported some movement among populations had occurred (Clark et al. 2015), the Louisiana black bear was delisted in 2016 (USFWS 2016). Genetic factors were not explicitly incorporated into recovery criteria or the population viability analysis for Louisiana black bears, though connectivity criteria in the
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Conservation Genetics
recovery plan implied that gene flow among populations should occur (USFWS 1995, 2016; Laufenberg et al. 2016). Previous genetic studies of Louisiana black bears primarily focused on two main issues: (1) whether bears in Louisiana warranted classification as a distinct subspecies (Hall 1981; Pelton 1991), and (2) whether the translocated Minnesota bears influenced the genetic makeup of bear populations in Louisiana. The conclusions of those studies were generally incongruous, primarily because different genetic markers used among studies precluded direct comparisons, the small number of markers or small sample sizes used had limited power to detect differences, or markers or analytical approaches were insufficient for quantitatively answering the questions posed (Miller et al. 1998; Warrillow et al. 2001; Csiki et al. 2003; Triant et al. 2004; Van Den Bussche et al. 2009). Multiple events occurred since those genetic studies that could have influenced genetic diversity and structure of black bear populations in the LMAV. First, 23 adult female bears and their 56 dependent young from the White River Basin in Arkansas were translocated to Felsenthal National Wildlife Refuge along the Arkansas–Louisiana border during a 2000–2002 reintroduction (Wear et al. 2005). Bears in Arkansas were classified as U. a. americanus despite the close geographical proximity to U. a. luteolus (Hall 1981; USFWS 1992), and the White River population likely had been isolated for nearly a century (Warrillow et al. 2001; Csiki et al. 2003; Van Den Bussche et al. 2009). Second, a population was reintroduced at the Three Rivers Complex in east-central Louisiana by translocating 48 adult female bears and their 104 dependent young from the Tensas River (n = 46 adults with 99 cubs) and Lower Atchafalaya (n = 2 adults with 5 cubs) populations during 2001–2009 (Benson and Chamberlain 2007; Laufenberg et al. 2016; Murphy 2016). The Three Rivers population was established in part to facilitate movement between the disjunct Tensas River and Upper Atchafalaya populations, which were separated by ~ 180 km. In addition to the reintroduction, 22 bears (18 males and 4 females) from the Lower Atchafalaya population were subsequently translocated to the Three Rivers Complex during 2010–2014 for management of human-bear conflicts (Murphy and Davidson 2015; Louisiana Department of Wildlife and Fisheries, unpublished data). Third, bears began naturally colonizing and reproducing in western Mississippi within the last decade along the state’s borders with Louisiana and Arkansas, but the exact source population(s) for that expansion was unknown (Simek et al. 2012). Thus, a complex history of population declines and growth, range contraction and expansion, and reintroduction and augmentation have resulted in multiple disjunct bear populations in the LMAV with a number of potential but unconfirmed genetic outcomes. Our goal was to evaluate the contemporary genetic status of black bears in the LMAV,
Conservation Genetics
testing four hypotheses focused on genetic diversity, structure, and gene flow within and among extant populations. First, we hypothesized that the Minnesota bears released in Louisiana during the 1960s survived and reproduced in the Upper Atchafalaya population but not the Tensas River population (Csiki et al. 2003; Triant et al. 2004). Thus, contemporary genetics would be influenced by Minnesota bears in the former, but solely be the product of bears native to the LMAV in the latter. Second, we hypothesized that bears in the Three Rivers population would be the exclusive product of Tensas River bears, given that Tensas River was the primary source for reintroduction, and that bears in western Mississippi would also be the product of Tensas River bears, given the geographical proximity (~ 45 km). Third, we hypothesized that gene flow has been insufficient to counter genetic drift in the three Louisiana black bear populations that existed when the subspecies was federally listed (Tensas River, Upper Atchafalaya, and Lower Atchafalaya), which would result in genetic structuring among populations. Fourth, we hypothesized that LMAV bear populations would consequently exhibit signs of genetic erosion, including low genetic diversity or small effective population size.
Materials and methods Sample collection, DNA extraction, and microsatellite genotyping We obtained hair samples collected from individual black bears ≥ 1 year of age in seven populations that corresponded to those present in the LMAV when listing occurred, those established by reintroduction or colonization, and the sources for augmentation and reintroduction. We used DNA from samples collected noninvasively (via hair traps) from bears during 2012–2014 as part of demographic studies in the Tensas River Basin (TRB), where 19% of Minnesota bears were released; Three Rivers Complex (TRC), where bears from Tensas River and Lower Atchafalaya River Basin (LARB) were recently translocated; Upper Atchafalaya River Basin (UARB), where 81% of Minnesota bears were released; and Lower Atchafalaya River Basin, where no bears from other populations were released (Fig. 1; Laufenberg et al. 2016; Murphy 2016). We also used DNA from samples collected from bears that were live-captured using culvert traps and Aldrich spring-activated foothold snares during a 2006–2012 bear resource-use study in western Mississippi (MISS; Simek et al. 2015), and DNA from samples collected noninvasively from bears via hair traps in the White River Basin (WRB) in southeastern Arkansas during a 2006–2007 demographics study (Clark et al. 2010). Samples from bears in the recently reintroduced population at Felsenthal National Wildlife Refuge along
the Arkansas–Louisiana border were not available. To evaluate the influence of the 1960s releases of Minnesota bears, we obtained DNA from samples collected from livecaptured bears in northeastern Minnesota (MINN) during a 2008–2013 demographics study (Garshelis and Noyce 2010). The area in Minnesota from which we collected samples (Itasca County) was ~ 200 km west of the area where the 1960s translocated bears originated (Cook County; Csiki et al. 2003; Van Den Bussche et al. 2009). Based on analysis of both contemporary and historical samples, bears throughout northern Minnesota assigned to a single genetic cluster and a single clade (Puckett et al. 2014). Thus, our sample of Minnesota bears was representative of the population from which the 1960s translocated bears were sourced. To avoid erroneous inferences caused by uneven sample sizes among populations (e.g., underestimation of the number of genetic clusters via structure analyses; Puechmaille 2016), we randomly selected 30–40 individuals from each of the seven aforementioned data sets [265 total individuals (129 males; 136 females)]. These sample sizes exceeded the recommended number of individuals needed from each population to accurately estimate parameters using allele frequency-based analyses of a large number of loci (25–30 individuals and ≥ 20 loci; Hale et al. 2012; Landguth et al. 2012; Puechmaille 2016). We stored all collected hair samples in individually labeled paper coin envelopes at room temperature, and we sent samples to Wildlife Genetics International, Inc. (Nelson, British Columbia, Canada) for DNA extraction and PCR amplification following the standardized methods described by Paetkau (2003). Each sample was comprised of ≥ 5 hairs, and laboratory technicians used Qiagen DNeasy Tissue kits (Qiagen, Valencia, CA, USA) to extract DNA from clipped guard hair roots or entire clumps of underfur hairs. Microsatellite genotyping was conducted following the protocols described by Paetkau and Strobeck (1994) and Paetkau (2003). Genotypes at the following 23 microsatellite markers were produced for individual bears: CPH9, CXX20, CXX110, D1A, D123, G1A, G1D, G10B, G10C, G10H, G10J, G10L, G10M, G10P, G10U, G10X, MSUT2, MU23, MU26, MU50, MU59, REN144A06, and REN145P07.
Genetic diversity, effective population size, and bottlenecks We grouped bears and analyzed their genotypes by the population from which they were sampled. We used MicroChecker v2.2.3 (Van Oosterhout et al. 2004) to test for null alleles, allelic dropout, and scoring errors at each locus. We tested for Hardy–Weinberg equilibrium and quantified linkage disequilibrium using the R software package genepop (Rousset 2008; R Core Team 2017). We applied Bonferroni corrections for multiple comparisons at each locus for Hardy–Weinberg equilibrium tests (α 1 population (Morin et al. 2010; Puckett et al. 2014). We detected four and 77 private alleles in the Lower Atchafalaya and Minnesota populations, respectively (Table A2 in Online Resource 1). Genetic diversity (AR and HE) and NE estimates were significantly lower for all populations in the LMAV than Minnesota (Fig. 2; Table A3 in Online Resource 1). Most genetic variation occurred within individuals (77%) followed by among populations (23%). Within the LMAV, White River had the lowest AR (2.08, 95% CI = 1.86–2.30) and HE (0.36, 95% CI = 0.28–0.44), whereas Three Rivers and Upper Atchafalaya had the highest AR and HE, respectively (AR(TRC) = 4.81, 95% CI = 4.36–5.26; HE(UARB) = 0.63, 95% CI = 0.59–0.67). Estimates of NB and NE were smallest for western Mississippi (NB = 3, 95% CI = 2–4; NE = 3, 95% CI = 2–4) and Three Rivers (NB = 9, 95% CI = 8–10; NE = 8, 95% CI = 7–9), and largest for White River (NB = 57, 95% CI = 28–85; NE = 49, 95% CI = 24–73) and Tensas River (NB = 46, 95% CI = 35–64; NE = 39, 95% CI = 30–55). We detected elevated FIS (0.11, 95% CI = 0.02–0.19) and evidence of non-random mating (Hardy–Weinberg exact test: χ246 = 126.03, P
