Considerations for the Propagation and Conservation of Endangered Lake Suckers of the Western United States

Artificial propagation has become a key component in preserving biodiversity and preventing extinction of compromised species, particularly those listed as Threatened (likely to become endangered within the foreseeable future throughout part of its range, as defined in the Endangered Species Act) or Endangered (in danger of extinction throughout all or a significant portion of its range) under the US Endangered Species Act (ESA 1973, as amended; herein “listed”). By 1990, 27% of recovery plans (documents that serve as guides for activities to help recover and conserve listed species) included captive breeding as a priority recovery (the process by which decline of listed species is arrested or reversed, or threats to its survival neutralized so that long-term survival in nature can be ensured) activity (Andrews and Kaufman 1994; Brown and Day 2002). In 2004, a survey conducted by the American Fisheries Society found that 42% of state and federal agencies relied on propagation for native fish recovery (Jackson et al. 2004). Artificial propagation programs designed for experimental reintroductions, translocations, and refuge populations (genetically managed captive populations designed to preserve wild endangered populations by acting as manmade refugium) often include intensive conservation research and monitoring of the potential genetic and ecological impacts on natural conspecifics, as opposed to recreational fisheries stocking that prioritizes high production targets. When planned and implemented responsibly, artificial propagation programs can sustain declining populations and may become critical when active reintroductions are necessary to reestablish genetically diverse populations (Smith et al. 2014).

Globally, more than 300 fish species are reared in captivity for supplementation or conservation, more than 290 of which are freshwater species (Welcomme 1992), with billions of individuals released into the wild annually (Brown and Day 2002). Intensive hatchery methods are frequently used to raise large numbers of fish for supporting depleted or at risk populations. In the United States, the U.S. Fish and Wildlife Service (USFWS) National Fish Hatchery System maintains over 70 national fish hatcheries in 35 states that propagate >100 aquatic species (USFWS 2016). State-run facilities number more than 500, and an estimated 2,000 private fish producers operate nationwide (Flagg and Mobrand 2010).

Seventy-one percent of recruitment failures among freshwater fish reintroduction efforts are associated with hatchery fish (Cochran-Biederman et al. 2014). Repatriation of hatchery-propagated Bonytail Chub (Gila elegans) began in 1981, but has yet to yield measureable adult recruitment (Minckley et al. 2003; Kappenman et al. 2012). Schooley and Marsh (2007) argue that after more than 30 y of population augmentation, Razorback Sucker (Xyrauchen texanus) in the Colorado River drainage may actually be further from recovery than when they were first listed in 1991. Kootenai White Sturgeon (Acipenser transmontanus) and Pallid Sturgeon (Scaphirhynchus albus) may become extinct in the wild before hatchery progeny assimilate into the natural spawning population (Israel et al. 2011). Still, relatively little has been published on how rearing methods affect growth, survival, and prognosis of listed nongame fishes such as Catostomids (O'Neill et al. 2011).

We developed detailed case studies of propagation methods and recovery progress for three species of endangered Catostomids to inform a new effort for two species endemic to the Klamath River Basin on the California–Oregon border that are not currently propagated: Shortnose Sucker (Chasmistes brevirostris) and Lost River Sucker (Deltistes luxatus). We review the risks and trade-offs associated with propagation, synthesize published strategies of propagation and husbandry, and evaluate the efficacies of each program in realizing species recovery. From these case studies, we discuss the most promising directions for developing a propagation strategy and suite of methods for these species.

Catostomidae are freshwater fishes found throughout warmwater rivers, coldwater streams, wetlands, and lakes of all sizes. At least 76 species are recognized and 75 are endemic to North America (Harris et al. 2014). Harris et al. (2014) reported that 35% are Endangered, Threatened, or of Special Concern (at risk of being listed as Threatened or Endangered), and the American Fisheries Society recognizes 49% as imperiled (Jelks et al. 2008; Table 1). Catostomids exhibit high variation in life history strategies and include both short- and long-lived species. We refer to of the genus Chasmistes (three species) and Deltistes (one species) collectively as western lake suckers. They are among the longest-lived Catostomids and are important constituents of freshwater fish assemblages in the large lakes of the arid Great Basin (Belk et al. 2011). All have experienced or are currently experiencing >15 y of negligible recruitment. June Sucker (Chasmistes liorus mictus) is endemic to Utah Lake, Utah, and Cui-ui (Chasmistes cujus) is confined to Pyramid Lake, Nevada. Shortnose Sucker and Lost River Sucker occur sympatrically in several large, shallow lakes of the Upper Klamath River Basin on the California–Oregon border (Figure 1). The Razorback Sucker is a related species endemic to the Upper Colorado River Basin; we include it in this review because it exhibits similar life history traits and has a long history of being propagated. See Text S1 (Supplemental Material) for full species case study accounts that include specific facility information and known rearing conditions, management action summaries, and accounts of historical use.

Shortnose and Lost River Suckers were historically abundant and fished by native peoples. Lost River Sucker, in particular, were caught in great numbers each spring by The Klamath Tribes of Oregon (Jordan and Evermann 1902). Localized extirpations were recorded as early as the 1920s (NRC 2004), but federal listing as Endangered under the ESA did not occur until 1988 (USFWS 1988) following marked declines of Upper Klamath Lake populations in the 1960s (USFWS 2013). Adult populations of both species have steadily declined since 2001, with very few individuals surviving the juvenile life stage and recruiting into the spawning population (USFWS 2013).

Cui-ui is the largest extant species of the genus Chasmistes. Females reportedly exceed 700 mm and males 662 mm total length (TL). They can live more than 40 y, reach maturity between 6 and 12 y, and are obligate stream-spawners. Spawning runs begin in late winter and early spring when Cui-ui stage at the mouth of the Truckee River and await appropriate water temperature and flows to initiate their migration. Spawning occurs over gravel substrates and females use their anal fin to bury eggs. Hatching occurs 7–14 d later, depending on water temperature. Approximately 5–10 d posthatch, larvae undertake a primarily nocturnal migration downstream to Pyramid Lake, and at 6 d posthatch, they commence feeding on chironomids and zooplankton (Bres 1978). This usually begins in April, with the adult spawning migration, and concludes in June or July, with larval outmigration.

Cui-ui stocking has been deemed largely successful in preventing extinction and bolstering recruitment of previously absent age classes, and to date, further population crashes seem to be absent (Scoppettone and Vinyard 1991; Schooley and Marsh 2007). Age-structured stock-recruitment models can be instrumental in understanding the dynamics of sucker populations by elucidating demographic parameters critical to understanding and refining the efficacy of augmentation programs (USFWS 1992). Cui-ui remains the only western lake sucker with a published population viability analysis model that can be used for management (Emlen et al. 1993; Scoppettone and Rissler 2007).

Razorback Sucker is one of the largest suckers in North America, capable of reaching 1 m in length and weighing up to 6 kg. Perhaps the most distinguishing characteristic is the predorsal keel, a large protrusion anterior to the dorsal fin that develops in mature fish. A riverine desert fish, Razorback Suckers can live >40 y and reach maturity between 2 and 4 y in the wild (Table 2). Historically, Razorback Suckers were found in abundance throughout larger streams of the Upper and Lower Colorado River Basins from Sonora, Mexico, to Wyoming, and they served as a major food source for native peoples. Since the 1940s, razorbacks have grown increasingly scarce, save for relic populations in the Green River, Lake Mead, and Lake Mohave that require augmentation through stocking. The remaining middle Green River Basin, Utah population was approximately 1,000 fish in 1988, but between 1989 and 1996 only 10 fish were found.

Spawning runs are triggered by water temperature and high-flow events that typically occur between April and June. Fish migrate en masse to shallow cobble bars where males form breeding territories and group spawning occurs. Eggs hatch approximately 1 wk after fertilization (Bestgen 2008), whereupon larvae drift from spawning areas into backwaters and floodplain wetlands where they remain until maturity. Razorback Suckers are highly adapted to persist in harsh, unpredictable environments that fluctuate between severe flood and drought conditions.

Development of propagation techniques for Razorback Suckers began in 1974 with the collection of 40 wild adults from Lake Mohave that became the first broodstock at Willow Beach National Fish Hatchery, Arizona (Toney 1974). Similar collections were made for Dexter National Fish Hatchery in New Mexico, and for many years stocking throughout central Arizona rivers and streams relied solely on the progeny of paired matings from these original 40 individuals (Schooley and Marsh 2007). These early efforts launched more than 30 y of intensive rearing-technique development and the release of >15 million fish in 544 release events at 200 locations. However, long-term survival rates remains unknown, and no new populations have been established (Schooley and Marsh 2007). Given the low success of these scattered population restoration larval stocking efforts (85% of releases were larvae) and the need to replenish precipitously declining wild adult fish, two different stocking approaches have been developed.

Captive brood propagation is used to produce all fish for release into the mainstem Colorado River, including Lake Havasu and other waters in Arizona, Utah, Colorado, and New Mexico, such as the Colorado, Green, and Gunnison rivers. The Lake Mohave repatriation program (Mueller 1995), however, exclusively stocks wild-born progeny. Adults spawn naturally and larvae are captured and reared in off-channel ponds protected from predation by nonnatives until maturity. Genetic monitoring shows that diversity within the population has been preserved using this method, although the wild population has been replaced with a repatriated population (Dowling et al. 2013). Culture practices vary widely and include different rearing environments, densities, feeding regimes, types of feed, and grading or sorting practices. Growth rates are currently reported to vary from 0.2 to 1.8 mm per day, with the highest rates being reported from natural or seminatural pond environments. Current propagation practices focus on maximizing growth in captivity to increase survival upon release.

Unfortunately, Razorback Sucker stocking programs have shown limited success in rebuilding populations (Schooley and Marsh 2007). Enhanced release strategies are a focus in part because early stocking efforts consisting of mass releases of young fish were largely unsuccessful due to suspected predation. Releases of larger sized fish show increases in subsequent-year survival (Marsh and Schooley 2007). Currently, lower basin repatriates are held in nonnative-free environments until they reach at least 300 mm TL, which takes at least 1 y to achieve. Schooley and Marsh (2007) suggested that had earlier stocking programs used today's knowledge of size-based release survival and held fish until they reached 350 mm TL, we would observe more than a 100-fold increase in survivors today. Survival rates of stocked fish >350 mm in Lake Mohave are nearly double compared to fish stocked at 300 mm, strengthening interest and effort in accommodating additional grow-out time. Detailed accounts of stocking strategies are available in Minckley et al. (2003) and Mueller (2006).

Predation continues to restrict natural recruitment throughout the Colorado River Basin, and conflicting ideals between native fish management and proponents of recreational fisheries continues to stymie nonnative fish removal. As a result, the prognosis for razorback remains extremely poor, irrespective of the advancements made in artificial propagation, and no self-sustaining populations have been established.

June Sucker populations were first observed to be in decline by 1970 (Heckmann et al. 1969), and by 1980 comprised a mere 0.3% of fish captured during standard Utah Lake surveys. It seems be the most imperiled lake Sucker, with an estimated wild population of approximately 300 fish and no documented natural recruitment (USFWS 1999). Spawning occurs in the lower 4.9 mi (7.9 km) of the Provo River during late spring and early summer, historically peaking during their namesake in June, but now much earlier (April and May) due to flow regulation. Adhesive eggs are deposited on gravel substrates and hatch within 4–10 d depending on water temperature. Larvae drift downstream where they begin to feed on zooplankton immediately. Hybridization with Utah Sucker is well documented, such that Miller and Smith (1981) declared that a pure strain no longer exists and that all remaining individuals are of mixed descent. More recent research (Cole et al. 2008) indicates that genetic exchange is an ancestral phenomenon, and until future research findings document otherwise, June Sucker used for propagation are targeted based on morphological differentiation in accordance with federal policy (UDWR 2004).

Releases of propagated progeny began in 1987, and limited successful recruitment into wild populations has been observed. The establishment of multiple independent, self-sustaining captive populations averted the imminent risk of extinction, but the prolonged absence of natural recruitment fails to spawn hope for the recovery of a self-sustaining population in Utah Lake. Billman and Belk (2009) cite successful cage culture in supporting recovery of Bonytail Chub and potential for exposure to more natural conditions than intensive fish culture affords, which may increase survival and recruitment upon repatriation. Allowing hatchery progeny an acclimation period in a controlled wild habitat might improve stocking outcomes, and using cages as seminatural grow-out habitats would be a cost-effective alternative to intensive hatchery culture.

Billman et al. (2011) found the ideal length of stocked fish to be 375 mm TL to avoid significant risk of avian and piscivorous predation. Stocking early (late spring or early summer) in the season avoids stress from elevated water temperatures and reduces vulnerability to stress-induced predation. The authors recommend that full-fledged augmentation programs begin with an initial exploration of optimal release sizes and timing and use a variety of rearing techniques before full program implementation. Rasmussen et al. (2009) also found a similar “optimal” length of stocking (roughly 350 mm TL), and more importantly a significant difference in survival related to captive rearing location (hatchery vs. reservoir).

The process of reviewing and synthesizing western lake Sucker propagation methods yielded several important insights for application to Klamath suckers that would not otherwise have been apparent. Many programs are increasingly emphasizing captive rearing of vulnerable life stages (Brown and Day 2002), particularly where early mortality due to predation or lack of habitat is a concern. In a recent general review, stocking variables were found to be less influential predictors of program failure than environmental characteristics; namely, water quality and quantity may actually hold more bearing on the success of a program than species biology (Cochran-Biederman et al. 2014).

Augmentation programs focusing on larval rearing and releases have shown negligible success (Marsh et al. 2015). In recent years, efforts have shifted from mass larval stockings to withholding larvae for rearing in captive, protected environments to increase survival during this critical life stage transition. Razorback Suckers are currently released at 250–300 mm TL, but researchers and managers have repeatedly published that a 350 mm release minimum is highly likely to increase survival (Schooley and Marsh 2007). There is interest in rearing fish to 400–500 mm, barring space limitations (Marsh et al. 2015). June Sucker are released at 200 mm, although >375 mm was identified as the ideal release length in 2011 (Billman et al. 2011). Nonnative predation is not suspected to be a primary limiting factor for Klamath suckers; however, rearing and releasing age-2 and/or older Klamath suckers will take precedence over younger, smaller fish until extensive and suitable larval rearing habitats have been restored.

Restoration of priority juvenile rearing habitats, in conjunction with increased grow-out, may maximize the potential for survival of wild larval and juvenile as well as released fish. Off-channel pond rearing of Razorback Sucker has shown promise for seminatural grow-out while also increasing available habitat for critical life stages (Mueller 2006). Testing the efficacy of a similar strategy for rearing young Klamath suckers could address multiple data gaps, including if and when wild fish recruit to these newly available habitats and variation in survival of wild vs. stocked young fish. High rearing densities, nonnatural feed types and foraging behaviors, lack of environmental variability and stimulation or enrichment, and predator unawareness associated with intensive fish culture facilities all contribute to increasing concerns for the ecological viability of propagated fishes (Naslund and Johnsson 2014). Habitat complexity, the ability to segregate or seek refugia from stressful schooling or grouping when needed, exposure to specific and nonspecific congeners, and real food are not difficult provisions to incorporate into propagation strategies. However, consistent annual production goals and space limitations often preclude the inclusion of such factors in many hatchery programs.

Reservoirs hold potential as seminatural refuges that can be used for a variety of propagation activities. Small reservoirs or ponds can be stocked and used as grow-out habitats for varying lengths of time. They may be used to hold additional refuge or broodstock groups in reserve. June Sucker and Razorback Sucker stocked in ponds for grow-out or as refuge populations have been documented to spawn naturally. “Volunteer” spawning by Bonytail Chub in hatcheries was documented beginning in 1981 (progeny are culled due to unknown parentage), and historical establishment of reservoir populations during the 1950s bodes well for pond or reservoir holding.

Several experimental rearing strategies are emerging with the potential to reduce the degree of domestication selection. Portable streamside rearing facilities developed to support Lake Sturgeon (Acipenser fulvenscens) recovery provide a cost-effective means to address relevant concerns about imprinting, spawning location fidelity, and genetic conservation for this remnant population (Holgren et al. 2007; Crossman et al. 2009). Efforts to rear June Sucker in situ have produced variable results, and although the potential for improvement is high this method is no longer used. Net pens are an attractive alternative that alleviates certain imprinting (e.g., no removal from natal waters) and genetic conservation concerns (e.g., no artificial propagation of brood), can be used on a transitory basis as needed, and do not require development of permanent structures. However, without adequate environmental conditions in situ this option requires refinement.

The need for scientifically defensible objectives for propagation programs is ubiquitous. Published evaluations and analyses of existing programs suggest that a paradigm shift is needed if managers hope to meet conservation-based goals, including reducing ecological and genetic risks while simultaneously implementing habitat restoration (Scott et al. 2005). Iterative, independent programmatic reviews are critical elements in evolving the use of propagation in native fish conservation (George et al. 2009). To increase effectiveness, emphasis needs to center around increasing survival of the earliest and most vulnerable life stages in relation to suspected limiting ecological factors (Ireland et al. 2002; Mueller and Wydoski 2004), while also addressing domestication selection risks as a result of extended time in captivity.

Captive populations must be carefully managed to maintain genetic diversity and representation of wild populations in successive generations (Lindberg et al. 2013) because heritably low fitness of hatchery fish is a concern, particularly in already depleted populations (Dowling et al. 2005; George et al. 2009; Christie et al. 2014a; Evans et al. 2014). Large-scale operations that rely on the high fecundity of a few females are likely to reduce natural genetic diversity within only a few captive-bred generations (Osborne et al. 2006; George et al. 2009; Israel et al. 2011; Christie et al. 2013). Effective population size reduction is a risk; offspring produced from a limited broodstock may swamp naturally produced fish (Britten and Brussard 1996). Therefore, continuous genetic monitoring is important to conservation propagation (Koike et al. 2008; Dowling et al. 2012). Complicated evolutionary dynamics and systematic challenges inhibit our ability to effectively address many genetic risks of artificially propagating catostomids (Cook 2001; Cole et al. 2008). Moreover, contention exists among researchers regarding the efficacy of conserving species complexes and the role of hybridization in recovery (Dowling and Secor 1997; Fiumera et al. 2004).

Wild population augmentation or supplementation efforts are further complicated by phenotypic variation that can arise in captive vs. wild populations (Belk et al. 2008). Relaxed natural and sexual selection pressures are compounded by increased domestication selection, resulting in reduced survival in the wild (Evans et al. 2014). In addition, developmental phenotypic plasticity may drive trait divergence between wild and captive individuals (Belk et al. 2008; Rasmussen et al. 2009; Evans et al. 2014). In a comprehensive review of the relative reproductive success of early-generation hatchery salmon, captive-origin fish were found to possess half the reproductive success of their wild-origin counterparts (Christie et al. 2014a). All species exhibited reduced fitness as a result of hatchery rearing, and the Christie et al. (2014a) caution that further investigations of the reasons for such reductions and long-term fitness of wild populations are needed. Fitness values can be calculated across multiple years to assess among year dynamics and variation, reduce multiyear estimate confidence intervals, and support more precise population-level estimates of relative reproductive fitness (Christie et al. 2012, 2014a, 2014b).

Poorly planned augmentation can result in systems that exceed natural carrying capacities, thereby reducing survival of both wild and repatriated individuals during critical life stages. Excessively large releases of select genotypes may swamp already compromised wild populations, resulting in misrepresentation of natural genetic diversity within a single generation (Fraser 2008; Milot et al. 2012; Christie et al. 2013). Large numbers of hatchery-origin Kootenai White Sturgeon progeny are currently being released to test the carrying capacity and any detrimental effects on other ESA-listed species (BPA 2013). Similarly, release strategy and ecological interactions have been related to density- and size-dependent mortality of many species of propagated fishes (Justice et al. 2009). Therefore, release plans need a means to assess success (i.e., tagging), and the ability to differentiate between species-level responses to restoration and supplementation and proactively evaluate declining species. For long-lived species such as sturgeon and suckers that require longer grow-out periods, this poses additional development needs with respect to program planning, construction of additional rearing spaces, and the potential inclusion of enrichment of conditioning.

Hatchery environments designed for production often lack resemblance to natural environments, with fish held at densities up to 100 times greater than those found in the wild (Flagg and Mobrand 2010). Hatchery progeny are less fit than their wild counterparts (Archer and Crowl 2014), and it has been suggested that life skills training, including a predator avoidance program, may increase survival of repatriated fish. In June Sucker, learned associations are strong for 2 d after training, but they disappear by 10 d without reinforcement (Archer and Crowl 2014). Rearing fish in more natural environments such as outdoor ponds before release can increase repatriate survival rates (Maynard et al. 1996), but also increases facility and financial resource obligations. Rasmussen et al. (2009) suggest incorporating truncated grow-out or acclimation periods in more natural environments immediately preceding release, which could be implemented in conjunction with controlled predator exposure and opportunities for natural foraging (Brown et al. 2013).

Setting quantitative thresholds and goals for production, survival, and release comprise the most straightforward portion of evaluating the impacts on these activities on the environment (Brown and Day 2002). Perhaps the most simplistic and relatable measure of success of augmentation is survival and recruitment to the spawning population. The next measure, less emphasized in most programs, examines robustness of and successful reproduction by progeny.

We initiated this review as a means to summarize captive rearing methodologies for western lake suckers and to direct the development of rearing criteria for Shortnose and Lost River Suckers. We quickly realized a comprehensive review of propagation strategies was warranted and more beneficial. Reviewing and synthesizing management plans specific to western Catostomids, within the context of current scientific guidelines provided the opportunity to incorporate available research from other long-lived species to supplement unavailable data specific to Catostomids. Future work will broaden the basis for a secondary review of all North American Catostomid species being propagated (e.g., Robust Redhorse Moxostoma robustum). We strongly encourage conservation practitioners to conduct similar review processes before embarking on new propagation efforts or other controversial conservation measures. Opponents of propagation often argue that it diverts resources from more cost-effective ecosystem restoration and habitat conservation measures. For this reason, we recommend that propagation not be planned as a direct or permanent fix, but rather a single component of population rebuilding.

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Text S1. State and federal facility background information and known captive rearing conditions for June sucker Chasmistes liorus, razorback sucker Xyrauchen texanus, and cui-ui Chasmistes cujus propagation compiled from agency and academic literature sources through 2015.

Found at DOI: DOI: http://dx.doi.org/10.3996/022016-JFWM-011.S1 (74 KB DOCX).

Reference S1. [BPA] Bonneville Power Administration. 2013. Kootenai River white sturgeon and burbot hatcheries program. Preliminary Environmental Assessment. Available: https://www.bpa.gov/efw/Analysis/NEPADocuments/nepa/Kootenai_Aquaculture_Program/KOOTENAI_AQUA_3B_PRELIM_FINAL_PEA_2-15-2013.pdf (February 2017).

Found at DOI: http://dx.doi.org/10.3996/022016-JFWM-011.S2 (5141 KB PDF).

Reference S2. Mueller GA. 2006. Ecology of bonytail and razorback sucker and the role of off-channel habitats in their recovery. U.S. Geological Survey Scientific Investigations Report 2006-5065, Reston, Virginia. Available: http://www.fwspubs.org/doi/suppl/10.3996/092012-JFWM-084/suppl_file/10.3996_092012-jfwm-084.s3.pdf?code=ufws-site (February 2017).

Found at DOI: http://dx.doi.org/10.3996/022016-JFWM-011.S3 (8618 KB PDF).

Reference S3. O'Neill MW, Ward DL, Stewart WT. 2011. Growth of razorback sucker (Xyrauchen texanus) at Bubbling Ponds Fish Hatchery. Arizona Game and Fish Department, Phoenix, Arizona. Available: http://www.azgfd.gov/w_c/documents/c10_growth_rasu_nov11.pdf (February 2017).

Found at DOI: http://dx.doi.org/10.3996/022016-JFWM-011.S4 (8618 KB PDF).

Reference S4. [USFWS] U.S. Fish and Wildlife Service. 1992. Cui-ui Chasmistes cujus Recovery Plan. Second revision. Portland, Oregon. Available: https://www.fws.gov/lahontannfhc/fish/cuiui/documents/cuiui_recovery_plan_2nd_revision_1992.pdf (February 2017).

Found at DOI: http://dx.doi.org/10.3996/022016-JFWM-011.S5 (3951 KB PDF).

Reference S5. [USFWS] U.S. Fish and Wildlife Service. 1999 June sucker Chasmistes liorus recovery plan. U.S. Fish and Wildlife Service, Denver, Colorado. Available: http://www.junesuckerrecovery.org/pdfs/990625.pdf (February 2017).

Found at DOI: http://dx.doi.org/10.3996/022016-JFWM-011.S6 (5362 KB PDF).

Reference S6. [USFWS] U.S. Fish and Wildlife Service. 2013. Revised recovery plan for the Lost River Sucker (Deltistes luxatus) and shortnose sucker (Chasmistes brevirostris). Available: https://www.fws.gov/klamathfallsfwo/suckers/sucker_news/FinalRevLRS-SNSRecvPln/FINAL%20Revised%20LRS%20SNS%20Recovery%20Plan.pdf. (February 2017).

Found at DOI: http://dx.doi.org/10.3996/022016-JFWM-011.S7 (1662 KB PDF).

Reference S7. [USFWS] U.S. Fish and Wildlife Service. 2016. National Fish Hatchery System. Available: https://www.fws.gov/Fisheries/Fish-and-Aquatic-Conservation-Facilities.pdf (February 2017).

Found at DOI: http://dx.doi.org/10.3996/022016-JFWM-011.S8 (50 KB PDF).

Reference S8. [UDWR] Utah Division of Wildlife Resources. 2004. Management plan for June sucker in captivity. Salt Lake City, Utah. Available: http://www.junesuckerrecovery.org/test/pdfs/Accomplishment%20Report%202003.pdf (February 2017).

Found at DOI: http://dx.doi.org/10.3996/022016-JFWM-011.S9 (745 KB PDF).

This manuscript was greatly improved by input from anonymous reviewers and the Associate Editor. Doug Routledge offered plentiful June Sucker expertise. Darrick Weissenfluh and Torrey Tyler provided helpful insight on earlier versions of this manuscript.

Funding for this work was provided by the U.S. Bureau of Reclamation Interagency agreement R13PG20201.

Any use of trade, product, or firm names is for descriptive purposes only and does not imply endorsement by the U.S. Government.