Abstracts
| Can’t People Ever Be Happy with the Species They Have?—
Impacts of Introduced Fish and Game Daniel Simberloff1 Perhaps the most devastating exotic sport fish in the U.S. is the common carp, introduced in 1877 by the U.S. Fish Commission. Its greatest impact is through uprooting vegetation and releasing sediment. At least 6 species listed under the U.S. Endangered Species Act are threatened by hybridization with introduced rainbow trout. Predation by introduced northern pike has even eliminated native species locally. Largemouth bass and brown trout among introduced sport fish have also greatly affected native fishes and invertebrates through predation. Many other introduced sport fish have damaged native species and ecosystems. Among introduced game, deer stand out as changing entire ecosystems through browsing and trampling, and the combination of introduced deer and introduced beaver has destroyed ecosystems of the Queen Charlotte Islands, Canada. Rooting by introduced wild boar, feral hogs, and their hybrids has also caused widespread habitat destruction, while nutria have literally turned thousands of hectares of wetlands and riverbanks into waterbodies. Introduced American mink contribute to the decline of the endangered European mink through hybridization, even though the sterile matings do not lead to introgression. Ecosystem-wide impacts are unknown for introduced game birds, but impacts on particular native species can be severe. For instance, mallards and other introduced ducks hybridize with native species, driving some to the verge of extinction. A few introduced columbiforms are nuisance pests. 1Department of Ecology and Evolutionary Biology, University of Tennessee, Knoxville, TN 37996, USA |
| Wildlife Reintroductions: Conceptual Development and Application of Theory Olin E. Rhodes Jr.1*# and Emily K. Latch2 Reintroductions remain a vital tool for the conservation and management of wildlife populations, and as the design of reintroduction programs has evolved, wildlife managers increasingly have become interested in the conservation of genetic diversity in reintroduced populations. High levels of genetic diversity within populations is thought to be desirable because the reservoir of existing genetic polymorphisms could directly influence a population’s ability to rapidly adapt to new environmental conditions and thus affect its long-term persistence. In this presentation, we identify and discuss three stages of the reintroduction process that directly influence the genetic variability present in a reintroduced population: 1) sampling efficiency of the available genetic diversity within source population(s), 2) reproductive success and survival of newly founded individuals, and 3) interactions among individuals and populations following a release. When individuals are removed from source populations for relocation, the effective sampling of genetic diversity present in those source populations varies as a function of a variety of factors, including the number of individuals removed, the presence of family groups in the sample, and the sex ratio of translocated individuals. Once animals are released, the newly established population may be at risk for further reductions in genetic diversity as stochastic differences in reproductive success and survival cause some alleles to be lost or to change dramatically in frequency. This process, referred to as genetic drift, may be particularly strong if the reintroduced population is slow to become established. Finally, reintroduced individuals may interact with individuals from resident populations near the release site. A variety of outcomes, including hybridization or deterioration of native genetic diversity may result if reintroduced individuals mix with nearby populations. By considering these processes prior to the initiation of a reintroduction program, wildlife managers can design management plans that maximize long-term success of reintroduced populations. 1Department of Forestry and Natural Resources, Purdue University, West Lafayette, IN 47906, USA 2Department of Biological Sciences, University of Wisconsin-Milwaukee, Milwaukee, WI 53211, USA *To whom correspondence should be addressed, E-mail: rhodeso@purdue.edu #Presenting Author |
| Bringing Moose Back to Michigan: Unrealistic Expectations Meet the Realities of Biology Scott R. Winterstein1*# , Dean E. Beyer, Jr.2, William B. Dodge Jr.1, Thomas D. Drummer3, Henry Campa IIIIII, Stephen M. Schmitt2, Thomas M. Cooley2, Robert W. Aho2, and James J. Maskey4 Moose (Alces alces) are native to MI, but were nearly extirpated by the end of the 19th century, primarily because of over-hunting and habitat destruction. By the mid-1960s, only 25–50 moose were estimated to exist in the UP. Efforts that began in the early 1970s to supplement the UP moose herd culminated in the relocation of 29 animals in 1985 and 30 animals in 1987 from Algonquin Provincial Park in Ontario, Canada to the west-central UP. The stated goal was to establish a self-sustaining population. At the time of the reintroductions, claims surfaced that were widely accepted by the citizens of Michigan that there would be 1000 moose in the UP by 2000–"1000 by 2000" had such a nice ring to it. Additional claims surfaced that at 1000 animals the moose population would be large enough to support a recreational hunt. Although not specifically stated as a project goal, a recreational moose hunt in MI was clearly a goal of many stakeholders. In 1995, deterministic population models based on demographic data gathered on the trans-located animals projected that the population would contain over 850 animals by 2000; acceptably close to the "promised" goal. Unfortunately, sightability models based on aerial surveys conducted in the same year, estimated a significantly smaller moose herd (less than 200 animals) in the west-central UP. The disparity in the population estimates raised serious questions about what, until that point, appeared to be a successful reintroduction of moose to the west-central UP. Issues were also raised concerning the use of a so-called "fudge factor" (i.e., the sightability generated correction factor) to "inflate" (i.e., statistically adjust) the number of moose "really there" (i.e., the count from the aerial survey). An investigation of the ecology of moose in the west-central UP was initiated in 1999 to estimate demographic parameters and refine the aerial-based sightability models. Intensive collaring operations were conducted in 1999–2001 and 2003. As a result of these operations, 109 moose were fitted with radio collars, of which 49 (45%) were still alive on 24 August 2005 when data collection ceased. Forty-five (41%) of the collared moose were known to have died by 24 August 2005. Disease and accidents were the primary sources of mortality; few suspected predator deaths were observed. We were unable to determine the fate of 15 of the collared moose. Calf survival from drop to 6 months of age was about 85%, while survival from 6-months to 1-year of age averaged 0.94 (n = 48; 95% CI = 0.85 – 1.00). Calf survival did not differ by sex. Yearling survival averaged 0.85 (n = 50; 95% CI = 0.74 – 0.95). Yearling female (n = 26) survival averaged 0.91, while male yearling (n = 24) survival averaged 0.79. Annual adult survival (1 June – 31 May) averaged 0.88 (n = 85; 95% CI 0.84 – 0.92). Annual adult male survival (n = 21; 0.85) did not differ from annual adult female survival (n = 64; 0.88). The probability of surviving from birth to age 5 is estimated to be about 0.51. We have used direct observations of radio collared females and analysis of pregnane concentrations in fecal samples (ug/g dry feces) from adult cows with active transmitters to estimate reproductive success and calf:cow ratios. Reproductive success (cows successfully producing at least one calf) ranged from 64% to 77%. The observed twinning rate, calculated as the number of births involving twins divided by the number of females giving birth, ranged from 11 - 28%. Overall, we estimated the calf:cow ratio to be about 0.6 - 0.8:1. The 2009 moose survey yielded an estimate of 420 moose in the west-central UP. Our demographic data suggest that the population is growing 0.05 – 0.10% per year. This slower than expected rate of growth is more likely attributable to a low reproductive rate (itself probably attributable to a lower than expected twinning rate), than it is to low survival rates. In response to growing pressure from various stakeholders (including members of the MI legislature) for a limited recreational moose hunt, population projection models were developed to determine the impact of various hunt scenarios on the future growth rate of the moose population in the west-central UP. Our results indicate that a small selective hunt (e.g., less than 10 males) will have negligible impact on the annual average growth rate of the population. However, it is unclear that even if such a hunt is proposed that it will be socially acceptable. 1Department of Fisheries and Wildlife, Michigan State University, East Lansing, MI 48824 USA. 2Michigan Department of Natural Resources – Wildlife Division, Lansing, MI 48909 USA. 3Department of Mathematical Sciences, Michigan Technological University, Houghton, MI 49931 USA 4Department of Biology, University of North Dakota, Grand Forks, ND 58202 USA *To whom correspondence should addressed. E-mail: winterst@msu.edu #Presenting Author |
| Freshwater Mollusk Population Restoration in the Cumberlandian Region and Mobile River Basin Paul Hartfield1#, Paul Johnson2*, Jeff Powell3, and Robert Butler4 The Southeastern US has the highest diversity of freshwater mollusks in the world. The Cumberlandian Region Basin (CRB) (comprising the Cumberland and Tennessee River basins) and Mobile River Basin (MRB) (comprising the Alabama and Tombigbee River basins) harbor the majority of the fauna. The molluscan faunas of these basins have also suffered an inordinate level of imperilment due to habitat fragmentation from large impoundments, historical point source discharges, and unregulated land use changes. Combined, at least 40 species of snails and 21 mussels became extinct over the last century in these watersheds. However, the basins still contain numerous federally listed mussels (31 CRB and 17 MRB), snails (4 CRB and 7 MRB) and candidates (6 CRB and 2 MRB). Surviving populations now occur in a highly fragmented landscape that continues to be stressed by human activities in some watersheds. However, other locations have improved due to human demographic shifts, land use changes, and better environmental regulatory mechanisms. However, natural emigration from stressed habitats to improved or restored reaches is usually prevented by large impoundments. Because of the high degree of habitat fragmentation and isolation, the survival and management of these imperiled aquatic communities will require continuous and careful husbandry of the species and management of their habitats. Plans have been developed for 57 mussels and 25 snails in the CRB and 26 mussels and 30 snails in the MRB. These plans provide guidelines for population restoration activities through artificial propagation technology and adult translocation in priority streams. Due to low population sizes of most species and to better measure success and minimize possible impacts of hatchery propagules to natural populations, these plans prioritize reintroduction over augmentation or translocation activities. Additionally, a habitat management and monitoring component is currently under development for both basins. 1US Fish and Wildlife Service, Jackson Field Office, Jackson, MS 39213, USA. 2Alabama Aquatic Biodiversity Center, Alabama Department of Conservation and Natural Resources, Marion, AL 36756, USA 3US Fish and Wildlife Service, Daphne Field Office, Daphne, AL 36526, USA 4US Fish and Wildlife Service, Asheville Field Office, Asheville, NC 28801, USA. *To whom correspondence should be addressed. E-mail: Paul.Johnson@dcnr.alabama.gov |
| Black-footed Ferret Recovery Progress and Continued Challenges Pete Gober 1#, Scott Larson2, and Paul Marinari1 The endangered black-footed ferret (Mustela nigripes) is a member of the weasel family. It weighs approximately two pounds and has a long, slender body marked by black feet and a black mask. It is one of the rarest animals in North America and for a time was thought to be extinct. Its recovery program is one of the oldest in the U.S. The ferret is an extreme specialist that depends on prairie dogs (Cynomys spp.) for food and shelter (Biggins 2006). Historically, it was found throughout the Great Plains, mountain basins, and semi-arid grasslands of North America wherever prairie dogs occurred. The ferret’s close association with prairie dogs was an important factor in its decline. Historically, prairie dogs occupied approximately 100 million acres. Over the past century occupied habitat has declined by 98% (Mac et al. 1998). This decline was largely due to the conversion of native grassland to cropland, widespread poisoning, and the inadvertent introduction of a non-native disease (sylvatic plague). In 1979, the ferret was presumed to be extinct after the last few individuals from a population in South Dakota died in captivity. Fortunately, in 1981, a small population was discovered near Meeteetse, Wyoming. Unfortunately, disease outbreaks occurred at Meeteetse in the early 1980s. Eighteen surviving ferrets were removed into captivity. Seven of these animals produced a captive population lineage that is the foundation of present recovery efforts (Hutchins et al. 1996). Extant populations, both captive and reintroduced, descend from these seven founder animals. The National Black-footed Ferret Conservation Center, managed by the U.S. Fish & Wildlife Service, and five zoos affiliated with the American Zoological Association, now maintain separate captive breeding facilities for approximately 290 ferrets. An estimated 7,000 ferret kits have been produced in captivity since 1987, and over 2,500 ferrets have been released into the wild. There have been 19 ferret reintroduction projects initiated since 1991 in eight states, Mexico, and Canada. A minimum of 400 breeding adults occur at these reintroduction sites (approximately 25% of the downlisting goal). Populations are currently self-sustaining at four sites. Many diverse partners have contributed to the recovery of the ferret including foreign governments, state and federal agencies, tribes, the American Zoological Association, conservation groups, and private landowners. Most of these partners are members of the Black-footed Ferret Recovery Implementation Team. Team members meet regularly to coordinate recovery efforts and address challenges to the recovery of the species. Some obstacles to ferret recovery have been successfully addressed, including the development of captive breeding and field reintroduction techniques. However, many challenges remain including providing enough secure prairie dog habitat to support ferrets in the wild and developing tools to manage sylvatic plague, which is usually lethal to both ferrets and prairie dogs. Many stakeholders consider prairie dogs a pest species. Large prairie dog complexes of a size necessary to support self-sustaining populations of ferrets are particularly at risk from poisoning. Incentive programs to conserve prairie dogs where appropriate and control them in other areas will be needed to achieve ferret recovery in the western U.S. The quality of ferret habitat is also limited by sylvatic plague. This disease was accidentally introduced into San Francisco in 1900 (Gage and Kosoy 2006). It was first detected in prairie dogs in 1932 and now occurs in all 12 states within the range of the ferret. Several potential management tools are being evaluated including direct vaccination of ferrets, dusting prairie dog burrows with an insecticide that kills the plague-bearing fleas, and vaccination of prairie dogs via oral bait. Despite the radically altered environment that reintroduced ferrets face today, recovery of this species is within reach. The challenge will be to continue ferret and prairie dog management efforts in order to complete the job. References
1U.S. Fish & Wildlife Service Natl. Black-footed Ferret Conservation Center, P.O. Box 190, Wellington, CO 80549, USA 2U.S. Fish & Wildlife Service, 420 South Garfield Ave., Suite 400, Pierre, SD 57501, USA *To whom correspondence should be addressed. E-mail: pete_gober@fws.gov #Presenting Author |
| Maximizing Success of Martes Reintroductions: Models, Data, and a New Hypotheses for Mating Patterns Roger A Powell1#, Jeffrey C Lewis2, Brian Slough3, Scott Brainerd4, Niel Jordan3, Alexei Abromov5, Vladimir Monakhov7, Pat Zollner8, Takahiro Murakami9 Historically, over-trapping for fur, loss and fragmentation of forest habitats, and predator control caused decreases and local extinctions of Martes populations. Protection allowed population recovery in some places but not everywhere. Because these predators are important components of ecological communities and can be valuable furbearers, they have been reintroduced to re-establish populations. Animals have also been released to augment low populations and to establish populations at new sites. Not all such translocations have been successful. We modelled reintroductions to predict criteria for success and tested model predictions using data from real reintroductions. The model predicted that more adult females released across several sites increases the probability of re-establishing a population. The number of males released should not affect success beyond a minimum number. American martens (M americana) have been translocated over 50 times with over 50% success. Fishers (M pennanti) have been translocated over 30 times with over 80% success. Pine martens (M martes) and house martens (M foina) have each been translocated 6 times, the former with at least 67% success and the latter with no confirmed successes. Japanese martens (M melampus) were were accidentally released once, successfully. Nearly 20,000 sables (M zibellina) were translocated in the former Soviet Union, re-establishing many populations. For real reintroductions, the 2 variables most strongly linked to success were the total number of animals released and the number of release sites. Sex-specific analyses for fishers, American martens and sables linked number of females released, number of release sites, and number of males released strongly with success. The contradiction between model and data regarding males may relate to the assumption in the model that all males are equal. We hypothesize many males must be released so that sufficient numbers of good breeders are released, possibly big males. 1Dept Biology, North Carolina State University, Raleigh, North Carolina, 27695-7617 USA 2Washington Dept Fish & Wildlife, Olympia, Washington, 98501 USA 3Whitehorse, Yukon Territory, Y1A 5S9 Canada 4Alaska Dept Fish & Game, Fairbanks, Alaska, 99741 USA and Norwegian Institute for Nature Research, Tungasletta 2, NO-7485 Trondheim, Norway 5Vincent Wildlife Trust, Waterside, Lowick Bridge, Ulverston, Cumbria, LA12 8EF UK 6Zoological Institute, Russian Academy of Sciences, St Petersburg, 199034 Russia 7Institute of Plant and Animal Ecology, Russian Academy of Sciences, Ekaterinburg, 620144 Russia 8Department of Forestry and Natural Resources, Purdue University, West Lafayette, Indiana, 47907 USA 9Shiretoko Museum, Hokkaido, 0994113 Japan *To whom correspondence should be addressed. E-mail: newf@ncsu.edu |
| Recolonization of Blacknose Dace after Restoration of an Acidified Stream Alexandra Fitzgerald 1#*, Mark Hudy 2, Chas Kyger2, and C. Andrew Dolloff 3 Restoration of acidified streams from acid deposition has focused on recovery of trout species with little attention to native non-salmonid species. Recovery of the entire fish assemblage is an important metric for success. I documented survival, dispersal and reproductive success of multiple attempts (1993-2009) to reintroduce blacknose dace (Rhinichthys atratulus) in an acidified headwater stream that had been restored through addition of limestone sand. Although brook trout (Salvelinus fontinalis) recolonize quickly after liming, blacknose dace have taken over fifteen years to produce an age-class. The re-establishment of blacknose dace in restored acidified streams may be very difficult and require frequent large stockings over multiple years to reestablish a reproducing population. My results provide insight into the difficulty and long-term recovery efforts that may be needed to restore the entire native fish assemblage to acidified streams. 1Virginia Polytechnic Institute and State University, Blacksburg, VA 24060 2U.S. Forest Service, Fish and Aquatic Ecology Unit, James Madison University, Mail Stop Code 7801, Harrisonburg, VA 22807 3USDA Forest Service, Department of Fisheries and Wildlife Sciences, Virginia Polytechnic Institute and State University, 350 Latham Hall, Blacksburg, VA 24061 *To whom correspondence should be addressed. E-mail: fitzgeralda@vt.edu #Presenting Author |
| Evaluating Success of Brown-headed Nuthatch and Eastern Bluebird Reintroductions in Everglades National Park
Gary L. Slater1#*, John D. Lloyd1, and Skip Snow 2 Despite the widespread use of reintroductions to reestablish populations of native species extirpated by habitat degradation or overexploitation (Wolf et al. 1996), rigorous, well-documented assessments of their outcome are rare. Such assessments are critical for determining success of the reintroduction and for identifying management actions needed to ensure the persistence of reintroduced populations. In this 9-year study, we evaluated the success of a reintroduction program for two bird species, Eastern bluebird (Sialia sialis) and brown-headed nuthatch (Sitta pusilla), in Everglades National Park, Florida. Both species were extirpated by the mid-1950s, as part of a larger wave of local bird extinctions, triggered by the widespread elimination and degradation of south Florida’s pine rockland ecosystem (Snyder et al. 1990). The reintroduction of Eastern bluebirds and brown-headed nuthatches was viewed as a test of the progress made in restoration of this unique, fire-dependent ecosystem in Everglades National Park. We evaluated the nuthatch and bluebird reintroduction program at 2- and 5-years following the cessation of translocations. Our objectives were to compare results of the evaluations at the two time intervals and to determine whether the reintroductions resulted in self-sustaining populations. During the period 1997-2001, 53 Brown-headed Nuthatches and 47 Eastern Bluebirds were translocated to ENP. We collected demographic data (population size, productivity, and survival) from the reintroduced populations in each of the breeding seasons from 1998 to 2007, excluding 2004, during which we collected no data. Thus, these data cover 4 years during which we translocated individuals to Long Pine Key and 5 years post-translocation. We also collected productivity and survival data from the donor site, considered a high quality reference site, from 1998- 2003. To evaluate short-term success of the reintroduction, we compared demographic parameters between the reintroduced and reference populations, with the criteria that similar estimates between populations were indicative of success. To evaluate long-term success of the reintroduction effort, we examined post-translocation demography using the reverse-time, capture–recapture models of Pradel (1996) to examine realized growth rate of each reintroduced species population (λt). Our criteria for long-term success was λt >1.0. At the end of the 2003 breeding season, the reintroduced brown-headed nuthatch population consisted of 46 adults and the reintroduced Eastern bluebird population consisted of 39 adults. Productivity and survival of both species were similar between the reintroduced and high-quality reference population, indicating short-term reintroduction success, even though populations remained small. In 2007, the reintroduced nuthatch population size equaled 52 adults, while bluebird population size was 35 adults. For the brown-headed nuthatch, the reverse-time, capture–recapture models indicated support for time-dependent survival probabilities, and thus we derived estimates of (λt) for individual years. Estimates of λt (mean = 1.04, range = 0.67-1.32) indicated that the population grew from 2001 to 2005 and then declined the following two years as adult survival declined. For bluebirds the model averaged estimates of (λt) was 0.92 (95% CI = 0.83–1.00), indicating that the reintroduced population of Eastern bluebirds was either stable or slowly declining from 2001–2007. Based on results from our short-term evaluation, reintroductions were considered a success. However, results from our long-term evaluation were mixed. Nuthatches appear to have positive population growth indicating a self-sustaining population, while bluebirds were stable or most-likely declining. Results from this study indicate the importance of long-term monitoring to evaluate reintroduced efforts. References
1Ecostudies Institute, P.O. Box 703, Mount Vernon, WA 98273 2Everglades National Park, 40001 SR 9336, Homestead, FL 33030 *To whom correspondence should be addressed. E-mail: glslater@ecoinst.org #Presenting Author |
| Response of a Re-introduced Chiricahua Leopard Frog Metapopulation to Bullfrog Removal C.R. Schwalbe1#*, P.C. Rosen2, D.O. Suhre2, B.H. Sigafus1, and D.H. Hall2 Breeding populations of the American Bullfrog (Rana catesbeiana) were removed from Buenos Aires National Wildlife Refuge, Altar Valley, Arizona, in 1999-2001, primarily by hand capture and draining of ponds, in some cases with pond-encircling fencing. Maintenance removals of bullfrogs that immigrated each year from off-refuge populations has prevented them from reproducing on-refuge for the subsequent decade. Marked bullfrogs were recaptured up to 11 km from the pond where originally marked (1). In 2003, larvae of the federally threatened Chiricahua Leopard Frog (R. chiricahuensis) were re-introduced onto the refuge at three of the bullfrog-free locations and have since been found at 12 refuge ponds, with reproduction observed at 6 of those sites. Bullfrog control has been facilitated by recent, ongoing drought, which has limited dispersal opportunities for the bullfrogs. However, rainfall in summer 2007, the wettest summer yet recorded for the refuge, allowed some normally ephemeral ponds to hold water through the winter, and bullfrogs moved into several sites, including four leopard frog breeding sites. Volunteers from multiple stakeholders responded to requests for assistance in removing these immigrant bullfrogs before they could breed in 2008. In 2009 we received permission from a ranch manager to remove all bullfrogs from the 29,000 ha ranch bordering the west boundary of the refuge that harbors the last source ponds of bullfrogs in the approximately 20 km X 70 km U.S. portion of the Altar Valley. There are 53 earthen watering ponds (stock tanks) on the ranch, nine of which contained bullfrogs; none contained Chiricahua Leopard Frogs. Removing bullfrogs by hand capture and shooting was extremely effective during the summer and fall of 2009, with more than 2500 bullfrogs removed, including all the adults. We observed < ten juvenile bullfrogs remaining in two of the nine ponds by the end of the bullfrog active season in October. Abundance of bullfrogs at one of those two ponds was reduced from a mark-recapture estimate of 7710 bullfrogs in 2003 (1) to < 10 seen in late 2009. We will remove those few remaining bullfrogs during spring 2010, before breeding can occur, and confirm that bullfrogs no longer occur in the valley. The rancher intends to sign a Safe Harbor Agreement (SHA) for the Chiricahua Leopard Frog with the U.S. Fish and Wildlife Service, with a zero-frog baseline, which we determined in 2009. There were significant hurdles involving trust and legal protections earlier in the project, but these issues appear to be resolving themselves satisfactorily. Other ranchers in the valley have now also expressed interest in obtaining SHAs and are exploring ways to finance the baseline surveys and participate in leopard frog conservation. There are at least four landscape-level bullfrog removal projects designed for leopard frog conservation ongoing in Arizona at present. Funding and support provided by DOI Amphibian Research and Monitoring Initiative, U.S. Fish and Wildlife Service Refuges, National Fish and Wildlife Foundation, Arizona Water Protection Fund, U.S. Geological Survey, and University of Arizona. References
1USGS Sonoran Desert Research Station, University of Arizona, Tucson, AZ 85721 2School of Natural Resources and the Environment, University of Arizona, Tucson, AZ 85721 *To whom correspondence should be addressed. E-mail: cecils@email.arizona.edu #Presenting Author |
| Using Structured Decision Making to Evaluate Re-introduction as a Tool to Recover an Arctic Sea Duck, the Steller’s Eider, in Alaska Tuula E. Hollmén1,2*#, James B. Grand3, Charles J. Frost1, James D. Nichols4, Ted R. Swem5, and Angela C. Matz5 Due to significant declines in numbers and reductions in nesting range, the Alaska-breeding population of Steller’s eiders (Polysticta stelleri) was listed as threatened in 1997 under the US Endangered Species Act. In Alaska, the species currently breeds intermittently and in low numbers on the Arctic Coastal Plain, and has nearly disappeared from its historical nesting areas in western Alaska. Due to the near-disappearance from historical nesting areas in western Alaska and high extinction risk based on population viability analysis, re-introduction is being considered as a tool to re-establish the western Alaska population and aid recovery of the Alaska-breeding population. First, a feasibility study was conducted to evaluate biological and socioeconomic feasibilities of re-introduction, develop plans for captive breeding and release methodologies, and assess uncertainties and risk factors of re-introduction of Steller’s eiders within their historical range. A captive reservoir population was established at the Alaska SeaLife Center and development of breeding techniques is underway. Uncertainties identified included lack of knowledge about original population declines, habitat suitability and future projections in face of climate change, and those relating to success of alternative release methodologies. Critical risk factors included 1) risk of disease transmission from captive populations to natural populations and the ecosystem, and 2) risk of loss of genetic diversity with captive breeding. Next, a structured decision-making (SDM) framework was applied to further evaluate re-introduction and associated uncertainties. Through the SDM process, we defined the decision problem, defined objectives, developed and considered alternative management actions and release methodologies, compared cost estimates, and evaluated the consequences of alternatives. The decision problem was to determine whether re-introduction is a cost-effective way to meet our objectives: re-establish a breeding population in western Alaska, enhance the viability of the Alaska-breeding Steller’s eiders, and aid recovery of the species. We used two population models to quantitatively evaluate alternative methodologies. Initially we used a population viability model incorporating current information about Steller’s eider vital rates and abundance. We also developed a simulation model of population growth for re-introduced and naturalized populations. We used both stage-specific and individual-based models to compare alternatives, to evaluate costs and benefits, and to identify key monitoring needs. We continue to use individual-based models to account for differences in survival for both sexes, over all age-classes, and geographic locations, and in productivity in relation to a suite of environmental variables, including location. We are also simulating variation in the above rates based on status (native, newly introduced, or introduced and subsequently naturalized) in the individual-based framework. Due to uncertainties surrounding alternative release methodologies and factors limiting growth of a re-established population, a rigorous adaptive learning strategy will be essential to monitor success, to refine re-introduction methodologies, and to maximize learning. Thus, development of appropriate monitoring tools and ability to adapt methodologies based on learning was identified as a key to further decision-making and ability to implement a successful program. 1Alaska SeaLife Center, Seward, Alaska 99664, USA. 2School of Fisheries and Ocean Sciences, University of Alaska Fairbanks, Fairbanks, AK 99775, USA. 3Alabama Cooperative Fisheries and Wildlife Research Unit, School of Forestry and Wildlife Sciences, Auburn, AL 36849, USA. 4Patuxent Wildlife Research Center, U.S. Geological Survey, Laurel, MD 20708, USA. 5Fairbanks Fish and Wildlife Field Office, U.S. Fish and Wildlife Service, Fairbanks, AK 99701, USA. *To whom correspondence should be addressed. E-mail: tuula_hollmen@alaskasealife.org #Presenting Author |
| Double trouble? Synergies among invasive exotics in south Florida Jerome A. Jackson1 and Bette J.S. Jackson2 Recognizing problems created by invasive exotic species is usually not difficult. Dealing with those problems can be especially troublesome unless the full range of species interactions within the invaded ecosystem is considered. We present a case history of a web of interactions involving several invasives in south Florida, including the Black Spiny-tailed Iguana (Ctenosaura similis), European Starling (Sturnus vulgaris), and Brazilian Pepper (Schinus terebinthifolius) and supporting interactions among these and other species. Phenology of flowering and fruiting of Brazilian pepper coincides with timing of low availability of native flowers and fruits as food resources and maximum potential for dispersal of seeds by resident iguanas and winter resident starlings. In evaluating ecosystem impacts, however, benefits of Brazilian pepper to resident and migrant native species must be weighed against the synergistic negative impacts of these species. 1Department of Marine and Ecological Sciences, Florida Gulf Coast University, Ft. Myers, FL 33965 2Department of Biological Sciences, Florida Gulf Coast University, Ft. Myers, FL 33965 |
| The Effects of Chinese Tallow (Triadica sebifera) on Lithobates sphenocephalus Hatching Cory Adams1#*and Daniel Saenz1 The idea of global amphibian decline was first introduced in the late 1980s (1), and is a major concern and topic of study. Some potential causes of amphibian decline that have been documented are habitat loss, fragmentation, disease, climate change, invasive species, and chemical contamination (2). Native and exotic invasive species could play a role in amphibian declines. Chinese tallow (Triadica sebifera) is a subtropical deciduous tree native to China and Japan (3). Chinese tallow was first introduced into the United States in the late 1700s and in Texas in the early 1900s (4). Chinese tallow is extremely abundant in parts of eastern Texas and has the capability of producing monocultures which can be in or near wetlands that are utilized by breeding amphibians. The impact Chinese tallow has on most amphibians is currently unknown. However, Leonard (5) conducted experiments on the effects of Chinese tallow on three species of anuran tadpoles and found that two of the species had lower survival when raised with Chinese tallow leaf litter. The two objectives of this study were to determine if Chinese tallow leaf litter affects hatching success of Lithobates sphenocephalus and if differential hatching success was observed; to determine the mechanisms that affected hatching. Lithobates sphenocephalus is a common species in eastern Texas which breeds in a variety of habitats ranging from ephemeral to permanent; however, the permanent sites typically do not contain fish. This species is also known to have the ability to breed year round in this part of its range. This is particularly important in understanding the potential negative effects that Chinese tallow can have on this species. Since L. sphenocephalus can breed any month of the year, it has the potential to lay its eggs in pools that have recently filled with water that contain Chinese tallow leaf litter. Chinese tallow leaves were collected from 4 November to 4 December 2009 from multiple trees located on the campus of Stephen F. Austin State University in Nacogdoches, Texas. Leaves were collected by stripping the loose leaves off of low hanging branches just prior to abscission. Leaves were then air dried indoors. We collected 15 L. sphenocephalus egg masses in varying stages of development ranging from 1 – 8 (6) from the Stephen F. Austin Experimental Forest in eastern Texas on 20 January 2010. Experiment 1 consisted of four treatments: 1 - a control with 20 L. sphenocephalus eggs; 2 - 20 L. sphenocephalus eggs with 4g/L of Chinese tallow leaf litter and aged tap water; 3 - 20 L. sphenocephalus eggs with 8g/L of Chinese tallow leaf litter and aged tap water; and 4 - 20 L. sphenocephalus eggs with 16g/L of Chinese tallow leaf litter and aged tap water. Eggs were placed in plastic containers (14 X 9 X 14) with mesh covered holes to allow water flow. Each treatment was replicated 15 times using the 15 different egg masses; therefore each replicate was blocked by egg mass. As a result, each replicate within a block contained full siblings. To determine hatching success the eggs were observed every 12 hours and the number of hatchlings recorded. We considered an egg mass hatched once half (10 eggs) of the eggs hatched. We collected an additional 7 L. sphenocephalus egg masses (stage 5 – 15) from the Stephen F. Austin Experimental Forest on 22 January 2010. Experiment 2 consisted of 8 treatments: 1 - a control with 10 L. sphenocephalus eggs; 2 - 10 L. sphenochalus eggs with 4g/L of Chinese tallow leaf litter and aged tap water; 3 - 10 L. sphenocephalus eggs with 8g/L of Chinese tallow leaf litter and aged tap water; and 4 - 10 L. spheoncephalus eggs with 16 g/L of Chinese tallow leaf litter and aged tap water. The additional four treatments contained the same number of L. sphenocephalus eggs and Chinese tallow leaf litter but were aerated. Each treatment was replicated seven times. Eggs and replicates were maintained and blocked the same as experiment 1. Hatching success was determined by observing the eggs every 4 to 8 hours until half (5 eggs) of the eggs hatched. We used a Hach Hydrolab Quanta to collect water quality parameters for each experiment. We collected water quality every 12 hours in experiment 1 and twice (once when eggs were placed in treatments and once when the experiment was complete) in experiment 2. Hatching success differed among treatments in experiment 1. All egg masses in the controls hatched and none of the egg masses in any of the Chinese tallow treatments hatched. Water quality also differed among treatments. Oxygen and pH were lower in treatments with higher concentrations of Chinese tallow. In experiment 2 we found differences in hatching success among treatments. Treatments with aeration had higher hatching success than non-aerated treatments. Water quality also differed among treatments. Obviously aerated treatments had higher oxygen concentrations and pH was lower in treatments with higher concentrations of Chinese tallow. We found that Chinese tallow reduces the hatching success in L. sphenocephalus. This reduced success is likely due to lower oxygen and lower pH caused by Chinese tallow leaf litter. In aerated treatments we observed higher hatching success than non-aerated treatments of the same concentration; however, no eggs survived in the highest concentration regardless of aeration. This is likely due to pH. More research needs to be done in order to determine if this invasive species could cause significant amphibian declines. References
1Southern Research Station, USDA, Forest Service, Nacogdoches, TX 75964, USA *To whom correspondence should be addressed. E-mail: coryadams@fs.fed.us #Presenting Author |
| Improving Predictive Models for Ecosystem Invasibility James T. Thorson1* and Nick Lapointe2# Invasive species represent a considerable challenge to the conservation and management of freshwater resources. Although studies have recently identified ecosystem characteristics (e.g., area, elevation, urban land use, etc.) that contribute to nonnative fish species richness (NNSR) at different spatial scales (Gido and Brown 1999; Leprieur et al. 2008; Marchetti et al. 2004), their designs were largely exploratory and results await replication in other systems. Few studies have used predictive modeling to identify regions with lower NNSR than expected based on their characteristics, and this identification can be useful to generate hypotheses about important effects that are not currently being considered in the study of ecosystem invasibility. In this study, we have two objectives. First, we compare the ability of several modeling methods to predict NNSR in watersheds in the mid-Atlantic region of the U.S. Second, we use the optimal model to identify watersheds with different NNSR than expected. We use a dataset of 78 watersheds (Hydrologic Unit Code-8), including information on NNSR per drainage and a variety of predictive variables representing propagule pressure, native species richness, and natural and anthropogenic abiotic factors. Predictive models include two parametric methods – generalized additive mixed models (GAMM, Wood 2004) and generalized linear mixed models (GLMM) – and two nonparametric methods – classification and regression trees (CART, Breiman et al. 1984) and random forest (RF, Breiman 2001). Spatial models including (1) correlation by distance using semivariograms, (2) region-specific heteroskesticity, and (3) region-specific random effects were included in the GLMM. Model selection for GLMM and GAMM was conducted using the Akaike information criterion and the unbiased risk-estimator, respectively.We use crossvalidation to contrast predictive ability among CART, RF, GLMM, and GAMM models. The differences between predicted and observed numbers of invasive species from crossvalidation for all four methods are contrasted in terms of root-mean squared error (RMSE), bias, and rank correlation. Rank correlation is a non-parametric measure of predictive accuracy, and is included because GLMM and GAMM incorporate non-Gaussian (Poisson distribution) assumptions for NNSR while RMSE may inherently favor models that implicitly assume normality (CART and RF). We then use the optimal model to compare a jackknife estimate of NNSR (i.e. excluding each watershed unit in turn and predicting it from a model including all other watersheds) with observed NNSR in each mid-Atlantic watersheds, and identify watersheds that have significantly different NNSR than would otherwise be predicted. Results show that RF outperforms CART and GAMM outperforms GLMM in terms of average error and rank correlation tests . RF provides realistic estimates of prediction error, while GAMM underestimate prediction error. Jackknife comparisons of RF predictions and observed NNSR identify two watersheds that have lower and one watershed that has higher NNSR than predicted. Based on crossvalidation results, we recommend that future studies of ecosystem invasibility use GAMM for parametric prediction and RF for nonparametric prediction. Parametric assumptions will in some cases make GAMM for efficient than RF, but can also lead to overestimation of accuracy. Prediction error is low for both of these models (±4 non-native species), suggesting that the variables in this study represent the majority of important factors regulating ecosystem invasibility within this region of the U.S. However, further improvements in prediction could be obtained through further study of the watersheds for which these models do not provide accurate estimates. These watersheds deserve additional analysis to identify possible factors that would further explain ecosystem invasibility. References
1School of Aquatic and Fisheries Science, Box 355020, University of Washington, Seattle, WA 98195-5020 2Dept. of Fisheries and Wildlife Sciences, 100 Cheatham Hall, Virginia Tech, Blacksburg, VA 24061-03 *To whom correspondence should be addressed. E-mail: JimThor@uw.edu #Presenting Author |
| Invasive Rodents in the United States and Their Management and Eradication Gary W. Witmer1* Many invasive rodents have become established in the United States and its territories. The species list includes:
These introductions occurred for a variety of reasons and by various pathways. Most occurred accidentally as a result of shipping and shipwrecks. Some were introduced as a source of subsistence food for people (possibly hoary marmots) or other animals (arctic ground squirrels as a food source of foxes introduced to islands for fur harvest). Nutria were introduced to numerous states for the fur industry. Gambian giant pouched rats were introduced indirectly due to the pet industry. These rodents have caused serious impacts to native flora and fauna and other resources (Howald et al. 2007, Hygnstrom et al. 1994). Damage occurs to crops and stored foods, structures, soils, and water quality. Endangerment or extinction has occurred to numerous species of native flora and fauna, especially on islands. There has a worldwide effort to eradication invasive rodents from islands with many successes (Howald et al. 2007). Since the early 1990s, United States agencies and collaborators have been eradicating rodents from various islands, primarily for conservation purposes. Of about 40 eradication attempts, 22 (55%) appear to have succeeded. For several islands, however, it is too early to determine if the attempted eradication has been successful or not. In the case of failed, eradications, rapid re-invasion by rats from nearby islands may be the reason. Numerous additional eradications are planned. We review the eradications, both successful and unsuccessful, that have occurred in the United States. Most rodent eradications in the United States involved the use of the anticoagulant rodenticides diphacione and brodifacoum. Rodenticides have been applied by hand-broadcast, bait station deployment, and aerial broadcast. In a few cases, cage traps alone were used on very small islands. This was done to lessen the potential impacts to an endangered lizard species on those islands. I briefly review the strategies and methods used in eradication projects and the efforts to mitigate potential non-target and environmental impacts. Finally, I consider some of the issues remaining invasive rodent management and eradication in the United States. Some of the challenges faced include the use of toxicants, land access, public attitudes, resource availability, and monitoring difficulties. References
1USDA/APHIS/WS National Wildlife Research Center, 4101 Laporte Avenue, Fort Collins CO 80521-2154, USA *To whom correspondence should be addressed. E-mail: gary.w.witmer@aphis.usda.gov |
| Describing the impact of nonnative fishes: spatial patterns and species traits Richard Pendleton1*, Nicolas W.R. Lapointe1, and Paul L. Angermeier1,2 Ecological and socioeconomic impacts of nonnative species invasions can range from negligible to severe. Patterns of variation in impact are scarcely documented, due in part to the lack of a concrete definition and quantifiable metrics of impact (Daehler 2001). Previous studies have used a single metric to estimate the degree of impact by a species (e.g., invader abundance, surveys of biologists’ opinions, or literature review; (Kolar and Lodge 2002, Marchetti et al. 2004, Ricciardi and Cohen 2007), yet no study has evaluated the strengths, weaknesses, or concordance of these approaches. Species traits are important factors in all stages of invasion, including impact. Individual metrics of impact may weight particular species traits differently. For example, small fishes may receive higher ratings from an abundance-based metric because such species are often very abundant. The impact of a species may vary across a region, depending on physicochemical, biotic, or socioeconomic contexts. The ability of various impact metrics to represent such spatial variation remains largely unexplored. In this study, we compare five approaches to estimate impact of nonnative fishes: 1) literature review of documented impacts (review metric), 2) abundance estimates based on agency collection records (collection metric), and 3-5) surveys of fish biologists regarding the abundance, ecological impact, and socioeconomic impact of a species in a specific drainage (abundance-, ecological-, and socioeconomic- survey metrics). We examine differences in: a) traits among species identified as having high impacts, as measured by each approach, b) impact ratings among major river drainages for individual species, and c) patterns of spatial variation among metrics. The study area included 11 USGS 6-digit hydrologic units (HUC-6s) in the mid-Atlantic region. A list of established freshwater nonnative fishes (n = 73) and their documented impact histories was obtained from the U.S. Geological Survey’s Nonindigenous Aquatic Species database (USGS-NAS). Each species’ documented impact was rated as: 1) none or unknown impacts, 2) mild or indirect decline in native species, or, 3) major decline in a single native species or moderate decline in multiple native species (Ricciardi and Cohen 2007). Collection records were used to estimate relative abundance of species by HUC-6 (absent = 1, rare = 2, common = 3). Fish biologists were surveyed online regarding the abundance, perceived nuisance status, and ecological impact of each species. Respondents rated the impact (low = 1, medium = 2, high = 3) of each species by HUC-6. To compare traits of species ranked highly among approaches, we chose the upper 25 % of species rated by each method. We paired metrics and compared highly ranked species unique to each method. We then tested for differences in traits between pairs of metrics, using the unique species as samples. We explored particular traits hypothesized to be favored by given metrics, instead of conducting exhaustive tests of every trait for each pair of metrics. We compared species-specific impacts among HUC-6s, focusing on three species found in all 11 HUC-6s; Cyprinus carpio, Lepomis macrochirus, and Micropterus salmoides. Nonparametric ANOVA was used to test differences in impact rating among HUC-6s, separately for each species and metric (DISTLM v.5, Anderson 2004). The review metric was excluded because it did not provide drainage-specific estimates of impact. Most traits tested did not differ among metrics. For example, abundance-based metrics (abundance-survey and collection metrics) did not favor species with shorter maximum total length when compared with the other three metrics. The socioeconomic-survey metric identified fewer top predators as having high impacts than review or collection metrics. Impact ratings did not differ among drainages (HUC-6s) for any survey metric (n = 115 – 128, pseudo-F = 0.934 – 1.377, P = 0.198 – 0.598). Ratings by the collection records metric differed among drainages for M. salmoides and L. macrochirus (n =121, pseudo-F = 1.785, 9.280, P = 0.061, 0.001), but the test could not be performed for C. carpio due to a lack of variation in ratings (some drainages did not contain samples with carp). The socioeconomic survey-metric was less likely to give top predators a high impact rating, probably because many introduced sport fish are not considered a socioeconomic nuisance. Though lists of high-impact species differed considerably among methods, no other traits differed significantly among metrics. The power of these tests was limited by small sample sizes of species unique to each metric; however our results show that top predators are less likely to be viewed as a socioeconomic nuisance. We did not observe strong spatial patterns in survey metrics, possibly because expert perceptions of impacts of these species did not differ throughout the region. The use of collection records may be the most appropriate approach for describing drainage-specific impacts; whereas, all approaches appeared reliable at a regional scale. References
1Dept. of Fisheries and Wildlife Sciences 100 Cheatham Hall, Virginia Polytechnic Institute and State University Blacksburg, VA 24061-0321, USA 22United States Geological Survey, Virginia Cooperative Fish and Wildlife Research Unit, Virginia Polytechnic Institute and State University, Blacksburg, VA, 24061-0321, USA *To whom correspondence should be addressed. E-mail: rmp0124@unt.edu |
| An Evaluation of Feral Cat Management Options Using a Decision Analysis Network Kerrie Anne T. Loyd1#* and BJayna L. DeVore1 Domestic cats (Felis catus) have been identified as "one of the world's worst invasive species" (Lowe et al. 2000) and non-native, invasive species are widely considered to be a leading cause of native species endangerment in the United States (Clavero and Garcia-Berthou 2005). The number of feral cats (abandoned, stray cats and those born in the wild) is estimated to be 70 to 100 million in the US today (Mott 2004). The issue of how to manage these populations is a topic of growing concern to diverse stakeholders in communities hosting populations of feral cats (Longcore et al. 2009), with implications for wildlife populations, cat welfare and human health. While other invasive predators are often controlled without incurring substantial debate (e.g. foxes, brown tree snakes), the issue of feral cat control remains controversial. Despite the increasing availability of scientific literature on feral cats and their impact on wildlife, the issue of feral cat management is currently influenced by the opposing interests and beliefs of passionate stakeholder groups concerned with protecting wildlife or preventing lethal control of cats. Management is a hotly debated and highly publicized topic in municipalities throughout the US. Often, stakeholders are exerting greater influence over the decision makers than the available scientific information on feral cat populations and environmental effects. For example, as a result of campaigns by the public and cat advocacy groups, Trap-Neuter-Release has been adopted by some cities (ex. Baltimore, MD) and states (ex. Illinois) as the sanctioned method of feral cat control. A model-based approach is necessary to quantify and evaluate outcomes of such management decisions. Because the number of cats is growing and the costs of dealing with them are great, planning for the best use of scarce resources is essential. One common management approach involves the permanent removal of cats, which are then euthanized (Trap-Remove [TR]) or removed to permanent enclosed sanctuaries (Trap-to-Sanctuary), while a second approach involves the capture, sterilization, rabies vaccination, and return of cats to their colonies (Trap-Neuter-Return [TNR]). In some cases basic TNR programs may be augmented with disease testing, vaccination and monitoring efforts (Trap, Test, Vaccinate, Alter, Return and Monitor [TTVARM]) or by the permanent removal of kittens through adoption (TNR+) (Hughes and Slater 2002). We have created a structured decision support model representing multiple stakeholder groups in order to facilitate the coordinated management of feral cats. A structured decision network was chosen to address this management issue because these can be particularly helpful in difficult management situations, including those involving invasive species, where the outcomes of management alternatives are uncertain and where numerous stakeholders with opposing parties are involved (Maguire 2004). We employed a Bayesian Belief Network to connect ecological findings to management decisions. In addition to predicting future population status of cat populations under 6 alternative management alternatives, our model incorporates analyses of costs, stakeholder values, and wildlife take by cats, thus considering both scientific and social aspects of the feral cat management problem. We informed our network by conducting population modeling exercises (stochastic model created in STELLA) that investigated the impact of alternative feral cat management options on cat populations and through the use of published data on cat life history and population dynamics. Our BBN results indicate that Trap-Remove or Trap-to-Sanctuary would be the optimal management decisions in most situations as removal would reduce feral cat populations quickly and prevent cats from taking a large number of wildlife prey. This decision balances a public interest in cat non-lethality with the value that stakeholders place upon the conservation of native wildlife. Our probabilistic network predicts that, in many circumstances, the high probability of substantial wildlife take under TNR management strategies outweighs the possible benefits of cat non-lethality. The use of a Bayesian Belief Network allows our analyses to become a tool for adaptive management, one to be updated as additional information on cat population dynamics, the impact of cat predation on wildlife, and attitudes of stakeholders becomes available. References
1Warnell School of Forestry and Natural Resources, The University of Georgia, Athens, GA 30602 *To whom correspondence should be addressed. E-mail: loydk@warnell.uga.edu |
| History of Introductions and Governmental Involvement in Promoting the Use of Asian Carps Anita Kelly1*#, Carole Engle2, Mike Armstrong3, Mike Freeze4, and Andrew Mitchell4 Numerous natural resource agency and media reports have alleged that Asian carps were introduced into the wild through escapes from commercial fish farms. This presentation chronologically traces the introductions of Asian carps (grass carp Ctenopharyngodon idella, silver carp Hypophthalmichthys molitrix, bighead carp H. nobilis, and black carp Mylopharyngodon piceus) and discusses the likeliest pathways of their introduction to the wild. Grass carp were first introduced in 1963 by the U.S. Fish and Wildlife Service. After that, state and federal agencies, universities, and private fish farmers interacted in efforts to introduce Asian carps, develop technologies for production, and promote their use in both public and private sectors in several states. These purposeful and legal introductions were to take advantage of the unique food preferences of Asian carps (planktivory by silver carp and bighead carp, herbivory of grass carp, and molluscivory by black carp). The U.S. Fish and Wildlife Service in Stuttgart, Arkansas had the first accidental release of diploid grass carp in 1966. Other early reports of grass carp in the wild were from waters in Alabama, Georgia, and Florida. Grass carp were first reported occurring in the wild in 1970, two years prior to the first private hatchery possessing grass carp. By 1972, 16 different states had stocked grass carp in open water systems. Silver carp and bighead carp were first imported purposely by a commercial fish producer in Arkansas in 1973. All fish were transferred to the Arkansas Game and Fish Commission (AGFC) by March 1974. The AGFC first successfully spawned silver carp and bighead carp later that year. The first report of silver carp in the wild was in Arizona in 1972, although strong evidence suggests this may have been a misidentification, followed by reports in the wild in Arkansas in 1975. The Arkansas report occurred two years prior to bighead carp and silver carp being returned to private hatcheries for commercial production. By 1977, silver carp and bighead carp had been imported to Alabama, Arizona, Arkansas, Illinois, and Tennessee. Research and stockings of silver carp and bighead carp were conducted by at least six state and federal agencies and three universities in seven states in the 1970s and 1980s. Public-sector agencies, which were successful in encouraging development and use of Asian carps that today are in commercial trade, are the likeliest pathways for the earliest escapes of grass carp, silver carp, and bighead carp. 1Aquaculture/Fisheries Center, University of Arkansas at Pine Bluff, Pine Bluff, Arkansas 2Arkansas Game and Fish Commission, 2 Natural Resources Drive, Little Rock, Arkansas 3Keo Fish Farm, P.O. Box 123, Keo, Arkansas 72083 4Harry K. Dupree Stuttgart National Aquaculture Research Center, Stuttgart, Arkansas *To whom correspondence should be addressed. E-mail: akelly@uaex.edu #Presenting Author |
| Guidelines for Propagation and Translocation for Freshwater Fish Conservation: Summary Principles and Examples of Source Population Selection Considerations Patrick L. Rakes1#*, Anna L. George2, Bernard R. Kuhadja3, James D. Williams4, Mark a Cantrell5, and J.R. Shute1 Principles and guiding rules for propagation, translocation, reintroduction, and augmentation (PTRA) efforts for conservation of freshwater fish (George et al 2009) are reviewed. The growing need for such actions for conservation purposes due to the continuing decline and fragmentation of fish populations, coupled with well-meaning but unsuccessful or even disastrous past PTRA projects, led to the publication’s list of recommendations and considerations, some of which are uniquely exemplified by the diverse fauna of the southeastern United States. A leading principle is "do no harm". No action at all may be better than a poorly considered effort that actually increases the risk of loss of a population or species. Essentially all PTRA activities are also secondary to and dependent upon habitat recovery and protection for any possibility of success. Until causes of decline are corrected, PTRA activities are just expensive, potentially risky stop-gap measures. Eight guiding "rules" for PTRA efforts include 1) determining whether PTRA is truly necessary and what forms are appropriate and justifiable, 2) assuring appropriate approval at all governance levels with a well-constructed plan based on expert advice from relevant scientists, 3) choosing and managing source populations wisely, designing propagation efforts appropriately, and protecting both captive and wild stocks, 4) conducting propagation and husbandry in as natural, minimal risk, and selection-neutral a manner as possible, 5) planning release methods, protocols, and sites well in advance, with contingency plans for unanticipated production or outcomes, 6) evaluating and monitoring target populations with appropriate adaptive responses, 7) informing the public in a variety of ways to reach and educate supporters, stakeholders, anglers, government officials, etc. , and 8) recording and disseminating all data that might assist similar efforts as well as the conservation and management of the target species. Current and recent case examples are discussed to illustrate Rule 3 with the taxonomic and biogeographic complexities influencing the selection of source stock or broodstock for reintroduction efforts of small non-game fish species in the southeastern United States. Reintroduction projects with madtoms (Noturus baileyi, smoky madtom, N. flavipinnis, yellowfin madtom, and N. flavus, stonecat), darters (Etheostoma sitikuensis, Citico darter, E. vulneratum, wounded darter, and Percina burtoni, blotchside logperch), a minnow (Erimonax monachus, spotfin chub), and a sucker (Moxostoma sp. "sicklefin redhorse", Moyer et al. 2009) in the Little Tennessee River basin in North Carolina and Tennessee demonstrate the role of species- and population-level physiographic province endemism on decisions. The use of genetic and taxonomic criteria as guides, often with surrogate species, is illustrated with several examples: the Barrens topminnow, Fundulus julisia (itself imperiled by an introduced non-native fish, Gambusia affinis) on the Barrens Plateau in middle Tennessee; the spotfin chub, Erimonax monachus, in Shoal Creek in south-central Tennessee, and the yellowfin madtom, Noturus flavipinnis, in the North Fork Holston River in southwest Virginia. References
1Conservation Fisheries, Inc., Knoxville, TN 2Tennessee Aquarium Conservation Institute, Chattanooga, TN 3Department of Biological Sciences, University of Alabama, Tuscaloosa, AL 4Florida Museum of Natural History, Gainesville, FL 5U. S. Fish and Wildlife Service, Asheville, NC *To whom correspondence should be addressed. E-mail: xenisma@gmail.com #Presenting Author |
| Allee Effects in Small Carnivore Populations: A Case Study of Recovering Gray Wolves Jennifer L. Stenglein1*#, Timothy R. Van Deelen1, Adrian P. Wydeven2, David J. Mladenoff1, Ted R. Swem5, and Angela C. Matz5 Allee effects threaten small populations with extinction when a component of individual fitness or growth rate is positively related to population size or density [1]. Multi-component Allee effects in reproduction and/or survival can interact in ways that are not well-understood, but have been documented in at least 15 species and in both natural and exploited systems [2]. One or more component Allee effects can lead to a demographic Allee effect when a population exhibits inverse density dependence at low population densities [1,2]. Allee effects may be particularly influential in reintroducted, newly established, or struggling carnivore populations because carnivores inherently exist at low densities, have elaborate social structures, and are sensitive to human activities. Gray wolves (Canis lupus) in Wisconsin, USA, are currently listed as ‘endangered’ as part of the Western Great Lakes Distinct Population Segment of gray wolves. The reestablishment of wolves began in the mid-1970’s and the first 15 years of recovery showed very little population growth. Once the population reached a critical size of ~50 wolves in 1994, the population moved into a robust growth phase. We assessed different mechanisms leading to Allee effects in the pre-growth phase of recovery (pre-1995) compared to the growth phase of recovery (1995-2009). More specifically, our objectives were to: 1) test for presence of a demographic Allee effect, 2) determine what mechanism(s) support an Allee effect, and 3) make recommendations for carnivore restoration when Allee effects are probable. We confirmed a strong demographic Allee effect by fitting a second-order polynomial to the plot of the per capita growth rate by population density. The Allee threshold was 12.1 wolves/1000km2 and per capita growth rate maximized at 25.7 wolves/1000km2. To understand the component Allee effects, we tested a number of survival and reproductive factors that could contribute to individual fitness [2]. We did not detect a difference in survival of adults (χ2= 0.66, P=0.41), yearlings (χ2= 0.03, P=0.86), or pups (χ2= 0.93, P=0.33) between the periods of recovery [3]. Additionally, we did not detect a difference in recruitment in terms of the proportion of packs in the population (t-test= -0.28, P=0.78) or the number of surviving pups per pack (t-test= -0.52, P=0.61). But, there was evidence that the proportion of dispersing wolves in the population and habitat saturation were mechanisms generating an Allee effect. Wolf packs have been called "dispersal pumps" because wolves are reared and often immediately leave the pack [4]. High rates of yearling dispersal have been linked to increased potential for colonization in unoccupied range [4]. We detected this pattern in our study with proportionally (t-test=6.74; P<0.0001) many more lone wolves (i.e., dispersers) in the population before 1995 (10.4% ±2.66) when the habitat was unsaturated, compared to three-times fewer lone wolves in the population (3.9% ±2.67) since 1995. Wolves were reaching the boundaries of preferred habitat by mid-1990’s, which was evidenced by the establishment of packs in the central forest region of the state [5]. This area is disjunct from the northern forest by 100km of agriculture, but had been predicted as preferred wolf habitat [5]. With primarily agricultural habitat left, wolf densities increased in areas of preferred habitat, which led to substantial population growth that has recently begun decelerating in a density dependent fashion [4,5]. To date, the reduced fitness of dispersing wolves has not been documented as a primary driver for a component Allee effect, but dispersers have reduced fitness because of lower survival and lower reproduction in low density wolf areas. African wild dogs (Lycaon pictus) have been well-studied in light of Allee effects and they are obligate cooperator breeders, similar to wolves [6]. There is strong evidence through simulations that when emigration is greater than immigration for wild dogs, their minimum group sizes drop below the Allee threshold [6]. High wolf dispersal for a population at low density may lead to an unstable population until a critical threshold is reached (>12.1 wolves/1000km2). Reintroduced and small carnivore population face numerous challenges because these populations typically exist at low densities and many have elaborate social structures. Potential for Allee effects should not be overlooked, especially because of the confounding nature of interactions between multiple Allee effects [2]. Demonstrating a strong demographic Allee effect is often done retrospectively, as in our study. Therefore, it is vital for managers and conservationists to understand mechanisms leading to component Allee effects in other, similar systems. References:
1Department of Forest & Wildlife Ecology, University of Wisconsin-Madison, Madison, WI 53706, USA. 2Wisconsin Department of Natural Resources, Park Falls, WI 54552, USA *To whom correspondence should be addressed. E-mail: jstenglein@wisc.edu #Presenting Author |