Research Paper

INTRODUCTION

Recovery of threatened and endangered species is imperative for efforts to conserve biodiversity. Approximately 20% of all wildlife species in the United States of America are at risk of extinction and 3 billion breeding birds have been lost over the past 50 yr in the USA and Canada (Stein et al. 2018, Rosenberg et al. 2019, North American Bird Conservation Initiative 2022). Conservation efforts for species facing precipitous declines face many challenges, including limited funding and hurdles with implementation of recovery actions. Indeed, these vital conservation efforts are often expensive and time-consuming and lack clear long-term goals and management objectives (Wilcove et al. 1998, Ferraro and Pattanayak 2006).

Captive breeding and conservation translocation programs are an increasingly popular tool used to bolster population sizes and preserve genetic variation of some of the rarest and most endangered species in the world (Seddon et al. 2007). Conservation translocations involve movement of any animal (whether captively raised or wild born) from one location to another with conservation objectives in mind (International Union for Conservation of Nature (IUCN) 2013). Conservation translocation programs tend to be costly, and success can be difficult to measure and is often low or unreported (Wolf et al. 1998, Fischer and Lindenmayer 2000, Bubac et al. 2019, Resende et al. 2020). The limited information available indicates 11–75% of conservation translocation programs that measured success reported positive outcomes, including achievement of project goals and survival or reproduction of animals after release (Griffith et al. 1989, Beck et al. 1994, Wolf et al. 1998, Seddon 1999, Stamps and Swaisgood 2007, Resende et al. 2020). The likelihood of success in conservation translocation programs has not changed over time (Fisher and Lindenmayer 2000, Bubac et al. 2019). Of more concern is that most translocation programs do not include post-release monitoring to assess success rates (Wolf et al. 1996, Bubac et al. 2019, Berger-Tal et al. 2020), despite repeated pleas and policy guidelines that stress the critical importance of post-release monitoring to assess the success or failure of these costly programs (Griffith et al. 1989, Wolf et al. 1998, IUCN 2013).

Endangered light-footed Ridgway’s rail (Rallus obsoletus levipes) populations have been supplemented by a conservation translocation program for more than 20 yr. A group of five organizations (state and federal agencies, zoos, and NGOs) have worked together to breed rails in captivity at three captive breeding facilities and annually release captive-raised rails (hereafter, captive-released rails) into coastal wetlands within their current U.S. range. Nearly 700 captive-raised rails have been released since the program began in 2001, but we know little about fates of these captive-released rails and the extent that they contributed to breeding populations. The number of breeding pairs in the wild have remained relatively stable (or even increased slightly) over the past 20 yr based on annual counts (Zembal and Hoffman 2022), but extents to which the translocation program has contributed to these trends are not known. In fact, we know very little about post-release survival of captive-released rails, which hinders abilities to implement possible improvements in the program to potentially boost post-release rail survival and thereby improve success of the program.

Our objectives were to address two questions: (1) what factors influence survival of captive-released juvenile light-footed Ridgway’s rails, and (2) does survival differ between captive-released and wild-caught juvenile rails? We first evaluated factors related to captive breeding environments and time in captivity on post-release survival of captive-reared rails. We then compared survival of captive-released juvenile rails and wild-caught juvenile rails. We used a suite of environmental variables related to marsh morphology and topography, predator abundance, movement, and environmental conditions to examine factors that influence juvenile rail survival.

METHODS

Study species and system

Light-footed Ridgway’s rails current U.S. range extends along southern California’s coastline, where 85% of vegetated wetlands have been developed or heavily degraded by urbanization (Stein et al. 2014, Brophy et al. 2019). The remaining wetlands available to light-footed Ridgway’s rails are surrounded by an urban matrix that is not used by rails. Reduction and fragmentation of available habitat contributed to low population sizes and limited genetic variability, leading to light-footed Ridgway’s rail’s federal listing as an endangered species in 1970 (U.S. Fish and Wildlife Service 1985).

Light-footed Ridgway’s rails inhabit tall, dense stands of California cordgrass (Spartina foliosa) in low-elevation marsh zones and forage along tidal channels and mudflats during low tides (Zembal et al. 1989, Eddleman and Conway 2020). Rails move up the elevational gradient of the marsh plain as tide levels increase to mid-elevation marsh where the dominant plant is pickleweed (Salicornia pacifica) and move up further into high marsh zones where vegetation is less dense and grows lower during the highest tides (Lewis and Garrison 1983, Massey et al. 1984, Zembal et al. 1989, Foin and Brenchley-Jackson 1991, Zedler 1993). Rail predators include diurnal raptors, owls, raccoons, domestic cats, and foxes (Eddleman and Conway 2020).

Light-footed Ridgway’s rails begin mate acquisition as early as January, seasonal peaks in breeding vocalizations occur from February through April (Zembal and Massey 1987), and peak hatch date is 20 April (Eddleman and Conway 2020). Adults care for young for about 5–6 wk post-hatching (Zembal and Fancher 1988), and flight feathers of juveniles are fully grown about 10 wk post-hatching (K. Sawyer, personal observation. A second peak in vocalizations occurs in September and October, corresponding with frequent observations of aggressive interactions among juveniles (Zembal and Massey 1987).

We conducted research at three marshes in San Diego County, California: Tijuana Slough National Wildlife Refuge, Batiquitos Lagoon Ecological Reserve, and San Elijo Lagoon Ecological Reserve (Fig. 1). These sites provide some of the largest remaining stands of cordgrass and tidal salt marsh within the U.S. range of this rare marsh bird. We chose these three marshes because they have some of the largest known populations of light-footed Ridgway’s rails within the USA (ca. 37% of the total U.S. population, or 157–196 pairs, based on local counts of breeding pairs during the years our study occurred; Table 1, Zembal and Hoffman 2022) and unoccupied suitable habitat to support additional rails (i.e., releases of captive-released rails). Selection of study sites known to have higher numbers of rails may create some bias in our results due to competition with resident wild rails; however, decisions regarding where to release captive rails were not in our hands, and marshes with lower densities of rails within the range of the species may not provide sufficient habitat for releases of additional rails. This study also relied upon our ability to capture wild juvenile rails at the same marshes where we released captive-raised rails and in similar quantities, a difficult objective at marshes with low rail populations.

Rail capture and transmitter deployment

Rails have been raised in captivity at three facilities in southern California: Living Coast Discovery Center (since 2001), SeaWorld (since 2001), and San Diego Zoo Safari Park (since 2005). Each facility cares for 1–2 adult pairs, and each pair produces 0–3 broods per year with 4–19 juveniles produced per pair each year. Rails are bred in nesting pens (58 m²) and moved to flight pens (366 m²) 50–70 d post-hatching. Breeding pens and flight pens consisted of some native vegetation (Salicornia pacifica, Spartina foliosa, Helianthus californicus, and Atriplex canescens) and a small creek or other water source. Juveniles remain in flight pens for 2–4 wk before being released at a marsh within their current U.S. range. Release sites are determined each year by an interagency working group that includes managers, researchers, and agency biologists.

Prior to release into the wild, we attached a 6-g solar-powered GPS transmitter with mortality mode (PinPoint; Lotek Wireless, Inc., Newmarket, Ontario, Canada) to each captive rail. We programmed transmitters to collect 2–3 GPS locations per day from June (when transmitter deployment began) through September when more daylight charged solar-powered transmitters more frequently but reduced this to one GPS location per day during shorter days from October through March. We used 2.5-mm tubular Spectra® ribbon (Bally Ribbon Mills, Bally, Pennsylvania, USA) to attach transmitters to rails via a backpack design (Sutherland et al. 2004).

We captured wild juvenile rails at the same three marshes where captive juvenile rails were released. We used carpet traps paired with audio lures (i.e., speakers broadcasting rail vocalizations) to capture light-footed Ridgway’s rails (Harrity and Conway 2020). We assembled carpet traps with 9-kg test copper-colored copolymer fishing line (KastKing, Eposeidon Outdoor Adventure Ltd., Garden City, New York, USA) with slip-loops tied to galvanized steel wire mesh. We attached several rows of slip-loops (6–7.5 cm in diameter) to rectangular mats of galvanized steel wire mesh to create carpet traps. We adapted methods of carpet trap placement from Harrity and Conway (2020) for use in tidal salt marshes. We attached the same transmitters (described above) to wild juvenile rails that we captured. We only attached transmitters to wild juvenile rails that were fledged (fully grown primary feathers). Capture and handling time was typically 30–60 min, and we attached transmitters only to rails when transmitters were ≤3% of a rail’s total body weight. We determined cause of death for rails when we were able to recover transmitters and carcasses within 72 h of transmitters switching into mortality mode. We classified fatalities as avian predation when feathers were plucked and intact, or as mammal predation when bones were crushed, feathers were broken, and/or bite marks were present on transmitters.

Statistical analyses

We used nest-survival models of Program Mark (implemented in RMark; Laake 2022) to estimate daily survival probability of light-footed Ridgway’s rails (White and Burnham 1999, Mong and Sandercock 2007, Lewis et al. 2017). Nest survival models are suitable for telemetry data where individuals are tagged on different days, individuals are monitored at irregular intervals, and exact date of death is not always certain (Mong and Sandercock 2007, Devineau et al. 2014).

We built encounter histories for models with five pieces of information: (1) day of transmitter attachment (i); (2) last day the individual was known alive (j); (3) last day an individual was detected or a mortality was detected (k); (4) fate of an individual, where 0 = survived and 1 = died; and (5) number of individuals with the same encounter history (n). We restricted our analysis to only timeframes between release of juveniles (June–December) through 31 March of the year following transmitter attachment. We used 31 March of the year following release as the end of a bird’s status as a juvenile and the start of their first breeding season given that rails often breed in their first spring following hatch. We ran two sets of models: (1) all possible combinations of covariates for pre-release conditions (i.e., captive-released survival), and (2) all possible combinations of covariates for both wild-caught and captive-released rails. We used Akaike’s Information Criterion corrected for small sample sizes (AIC_c; Burnham and Anderson 1998) to compare fit of candidate models.

Captive-released rail covariates

We selected covariates for use in survival analyses pertaining to pre-release conditions that would only impact captive-released rails. We considered covariates including: (1) breeding facilities where rails were raised (Living Coast Discovery Center or SeaWorld; only two facilities had pairs that produced young during our study), (2) breeding pairs that produced a rail (five different breeding pairs produced young during our study), (3) age (days post-hatch) of rails at release, (4) mass (to the nearest g) of rails at release, (5) number of days spent in flight pens prior to release, and (6) day of year a rail was released. We also considered interactions between mass and age, mass and release date, mass and days in pen, as well as between release date and age, and release date and days in pen.

Wild-caught and captive-released rail covariates

We separately tested for differences in survival based on number of days alive between captive-released and wild-caught juvenile rails with two-way independent sample Welch’s t-tests. We also assessed a suite of covariates pertaining to post-release conditions, including metrics for predator abundance (see below), tidal stage (relative to high tide), habitat conditions, home range size, and time since release (for captive rails) or time since capture and release (for wild rails) using nest-survival models in Program Mark.

Predator abundance

We used local raptor abundance data from eBird to quantify time-specific and location-specific raptor abundance. We used the ebirdst package (Strimas-Mackey et al. 2022) in Program R 4.3.1 (R Core Team 2023) to access eBird Status and Trends data (Fink et al. 2022). We obtained relative abundance values for raptors within 2.96 km² cells that covered our study sites. We accessed eBird Status and Trends data for nine raptor species with relative abundances greater than 0 within our study sites and that we identified as potential rail predators: (1) red-tailed hawk (Buteo jamaicensis), (2) northern harrier (Circus cyaneus), (3) peregrine falcon (Falco peregrinus), (4) white-tailed kite (Elanus leucurus), (5) red-shouldered hawk (Buteo lineatus), (6) Cooper’s hawk (Accipiter cooperii), (7) barn owl (Tyto alba), (8) great horned owl (Bubo virginianus), and (9) short-eared owl (Asio flammeus). We extrapolated weekly abundance estimates to a daily scale and used those values as an explanatory variable to evaluate whether raptor abundance affected daily survival of light-footed Ridgway’s rails.

Avian predation was the primary cause of death for a closely related subspecies, California Ridgway’s rails (Rallus obsoletus obsoletus) (Overton et al. 2014, Casazza et al. 2016). Thus, we included raptor abundance as a possible explanatory variable in our suite of candidate models to assess effects of higher predator abundance on daily survival of light-footed Ridgway’s rails. We included an interaction between origin (captive-released or wild-caught) and raptor abundance to examine whether raptor abundance affected survival of captive-released and wild-caught rails differently.

High tide

We included water level at high tide for each day of the study to evaluate potential effects of high water levels and inundation of marshes on availability of cover for rails to seek refuge from predators. We obtained tide data from the La Jolla, California tide monitoring station at the National Oceanic and Atmospheric Administration (NOAA) Tide and Currents webpage (http://tidesandcurrent.noaa.gov). We extracted the highest water level reported for each day of the study to assess whether higher tides affected daily survival of rails. We included an interaction term for origin of a rail (captive or wild) and tide height to account for the possibility of captive-released rails responding to high tides less optimally than wild-caught rails.

Marsh morphology and topography

We extracted the following habitat-related covariates for each GPS location collected: average elevation, average distance to roads, proportion Spartina foliosa, proportion mudflat, proportion Salicornia pacifica, and proportion high marsh. We calculated average elevation for each rail based on all recorded locations of an individual. We obtained elevation from a LiDAR-derived digital elevation model with 5-cm accuracy obtained from the California Coastal Conservancy LiDAR Project through NOAA (Office for Coastal Management 2023). We assigned each GPS location to one of the following habitat types based on a local vegetation layer that identified dominant vegetation within a polygon: (1) Spartina foliosa, (2) Salicornia pacifica, (3) Mudflat, or (4) High Marsh (San Diego Geographic Information Source 2015). We calculated the proportion of points in each category and used those proportions as covariates. We calculated distance to roads for each location of every rail and then averaged across all points for each rail based on a local roads GIS layer (San Diego Geographic Information Source 2023).

Home range size

We calculated home ranges for all rails to account for effects that increased space use may have on rail survival. We used adaptive localized convex hulls (Getz et al. 2007) to estimate home range sizes by using package adehabitatHR in Program R (Calenge 2006, R Core Team 2023). Convex hulls are appropriate given the oblong, linear shape of marshes where circular home ranges produced by autocorrelated kernel density estimators contained developed areas outside of marshes (i.e., areas not used by rails).

Time since release/capture

We included an interaction term for number of days since release and origin (captive-released or wild-caught) of a rail to test whether captive-released rails acclimated to their new environments over time. We also included two interactions (high tide * days since release/capture, and raptor abundance * days since release/capture) to test whether these factors affected survival differently across time.

RESULTS

We attached transmitters to 46 juvenile captive-released rails and 42 juvenile wild-caught rails from June to December 2020–2022 at three study sites (Table 1). We recovered 68 transmitters and determined cause of death for 50 individuals; 48 were identified as having succumbed to avian predators and two to mammalian predators.

Captive-released rail survival

We deployed transmitters on 58% of 80 captive-released rails during our 3-yr study. Most rails (n = 39) were released from June through August, but seven rails were released from September through December. Only three of 46 captive rails released with transmitters during 2020–2022 survived to the next breeding season (through 31 March). The average number of days survived for the 43 captive-released juveniles that did not survive until the subsequent breeding season was 49 d (SE = 8 d). None of the seven juveniles released September through December survived longer than 8 d, with an average of 4 d alive.

Release date was negatively associated (β = -0.022, 95% CI [-0.029, -0.014]) with daily survival for the 46 captive-released juvenile rails based on our top model (Table 2, Fig. 2). All models within 2 AIC of the top model contained release date, and these nine models account for 92% of model weights (Table 2). All variables in our top models except release date had confidence intervals that overlapped zero. We present results from only our top model with no model averaging because beta estimates were within 0.001 units across all models with ≤2 AIC_c, and model averaging our results did not affect inference (Richards 2005, Richards et al. 2010). Daily survival probability began decreasing when releases occurred in mid-August and continued to decline with later releases. Daily survival probability for rails released at the end of June was 0.993 (SE = 0.002) compared with 0.972 (SE = 0.004) for birds released in early September (Fig. 2). To put this into perspective, if all rails were released at the end of June (n = 20; highest number of rails released in a year during our study), then three rails would survive to the next breeding season given a constant daily survival probability of 0.993. However, if all rails were released at the beginning of September, no rails would survive to the next breeding season given a constant daily survival probability of 0.972.

Wild-caught and captive-released rail survival

Wild-caught juveniles survived longer than captive-released juveniles (t = -2.6, P = 0.01); six wild-caught rails survived until 31 March, and the average number of days survived for the 36 wild-caught juveniles that did not survive until the subsequent breeding season was 92 d (SE = 11). We discuss below all parameters from our top model with estimates that had confidence intervals that did not overlap zero.

For our analysis that examined factors that influenced survival for all 88 juveniles (captive-released and wild-caught), our top model included the following variables: sex of a rail, mean elevation of all recorded locations, and three interactions among post-release variables: days since release/capture and water level at high tide, days since release/capture and origin of a rail, and raptor abundance and origin of a rail (Table 3). Daily survival probability was 0.994 (SE = 0.002) and 0.979 (SE = 0.007) for wild-caught and captive-released rails, respectively. Wild-caught juvenile rails had a 23.5% probability of surviving to the next breeding season, whereas captive-released juvenile rails had only a 0.3% probability of surviving to the next breeding season. However, a simple extrapolation of survival of captive-released rails across our entire study period (i.e., until the next breeding season) is misleading because captive rail survival is low initially but steadily increases to a point where it was similar to wild-caught juveniles (if a captive-released rail survived beyond 100 d after release). Captive rails released at the end of June, however, had an average daily survival probability of 0.993 (SE = 0.002), nearly equivalent to wild rail average daily survival.

Male rails had lower survival (0.984, SE = 0.004) than females (0.992, SE = 0.003) with confidence intervals that did not overlap zero (Table 4; t = 3.33, P = 0.001). The effect of daily raptor abundance on survival differed between wild-caught and captive-released rails; higher daily raptor abundance at a marsh decreased survival for captive-released rails, but wild-caught rail survival was not associated with daily raptor abundance (Table 4). Relationships with raptor abundance had wide confidence intervals (β = -10.61, 95% CI [-21.02, -0.02]), but wide confidence intervals are not surprising given low precision of the metric we used to estimate raptor abundance.

Survival was negatively associated with elevation (β = -1.00, 95%CI [-1.63, -0.37]), particularly at elevations above 6.5 m (Table 4, Fig. 3). Captive-released and wild-caught rails occupied similar mean elevations (5.1 m vs 4.9 m; t = 1.62, P = 0.11), but captive-released rails used the highest portions of tidal wetlands, whereas wild-caught rails did not; range in elevations was 4.1–7.6 m and 4.1–5.6 m, respectively.

Survival of captive-released rails increased as time after release increased, whereas survival of wild-caught rails did not change as time since capture increased (Fig. 4). Survival probability of captive-released rails was 0.978 (95% CI [0.912, 0.985] immediately after release and increased with each day in the wild. After 172 d in the wild, captive-released rail survival probability reached 0.994 (95% CI [0.982, 0.998]), matching wild-caught rail survival, which did not vary through time (Fig. 4). Survival immediately after release showed little change as tidal height increased, but survival was positively associated with higher tides (but highly variable) as days since release/capture increased (Table 4; High Tide * Days since release interaction).

DISCUSSION

We found that captive-released rails survived poorly compared with their wild-caught counterparts, but several factors that were associated with survival of captive-released rails could be altered by managers to increase survival probability of future captive rail releases. This information is vital for managers and decision makers when determining how to improve and assess the success of a translocation program.

Impacts on survival of captive-released and wild-caught rails

Captive-released rails released later in a year may face more challenging conditions than those released earlier (i.e., before September). California Ridgway’s rails, a closely related subspecies, have lower survival in winter, especially during high tides (Albertson 1995, Overton et al. 2014). Spartina foliosa, the dominant refugium for rails, senesces over winter, which reduces cover and contributes to lower rail survival during winter (Overton et al. 2014). This reduced overhead cover coincides with the highest high tides of the year and increased raptor abundance (Zetler and Flick 1985). Thus, captive rails released in September–December may struggle to respond optimally to reduced availability of cover combined with higher water levels and increased predator abundance. Heightened mortality risk during winter is likely intensified by post-release behavioral modification, where behavior and decisions made by captive-released animals in a novel environment are sub-optimal initially but improve as experience is gained and time since release increases (Berger-Tal et al. 2014, Berger-Tal and Saltz 2014).

Our results suggest that captive-released rails do indeed fare better with time. Daily survival probability was positively associated with time after release for captive-released juvenile rails (but not for wild-caught juvenile rails), implying that captive-released rails are particularly vulnerable immediately after release as they adjust to their new environment. This increase in daily survival probability after release may be further evidence of post-release behavior modification; captive-released rails are learning after an initial period of sub-optimal behavior post-release. The relationship between survival and days since release/capture for the highest water levels may indicate that rails are adjusting to higher water levels as they get older (i.e., wild-caught rails) or as they spend more time in the wild (i.e., captive-released rails). Alternatively, the captive rearing environment may artificially inflate survival, allowing some individuals to survive until transmitter age (i.e., fledgling age) that lack comparable behavior and predator avoidance resulting in high mortality immediately after release. Future research that holds much promise to improve success includes studies designed to address the potential for post-release behavioral modification in this species and how it may affect survival and predator avoidance. For example, habitat selection or movement analyses that address whether selection or movement of rails changes with more time in the wild and if changes in behavior through time influence survival would be helpful to inform modifications to improve success of future releases.

Lower survival for rails that use higher portions of the tidal marsh system (i.e., high marsh zone) has implications for future management of this coastal species. Sea-level rise is expected to drastically alter coastal wetlands across southern California; even under low sea-level rise scenarios, high marsh habitat is expected to be lost by the end of the century (Thorne et al. 2018, Rosencranz et al. 2019). As sea levels rise, coastal marshes in southern California are constrained by urban development, and marshlands have little to no space to migrate upland to escape rising water levels (Thorne et al. 2018). This will only exacerbate lower survival when rails use higher portions of a marsh, as these high marsh zones may no longer have vegetation, and rails will be met with busy roadways and suburban neighborhoods.

Transmitters can have a negative effect on bird behavior and survival (Barron et al. 2010, Severson et al. 2019), and harness attachments and large reflective surfaces (i.e., solar panels) are associated with lower estimated annual survival for some birds (Burger et al. 1991, Severson et al. 2019). Survival probabilities of captive-released and wild light-footed Ridgway’s rails without solar-powered transmitter attachments may, therefore, be higher than our study indicates.

Management implications

Given the results of our study, managers who wish to increase survival of captive-released rails could release as many captive-raised rails as possible in June and July to maximize survival initially and give captive-released rails time to modify their behaviors and optimize their decision making in their new environments. This single management approach would increase the number of rails that survive to the next breeding season and contribute to breeding populations.

Several management approaches that may increase post-release survival and warrant evaluation include pre-release predator avoidance training, acclimation to a release site prior to release (soft releases), or exposure to simulated tidal action while in captivity. Indeed, the majority of successful species translocation projects included pre-release training or acclimation (50% and 75%, respectively; Beck et al. 1994). Soft releases have also been successful with other rail translocations (Wanless et al. 2002). Exposing captive animals to food sources mimicking those found in the wild, to natural foraging conditions, and providing supplemental feeding after release have all been found to increase survival in the first year after release, in particular with Gruiformes (Roberts and Luther 2023). Future research that continues to monitor rails post-release will help to further improve the success of releases. Monitoring changes in habitat selection or response to predators with time spent in the wild would be especially useful. Research designed to compare new management or release strategies to existing strategies would help to elucidate the best approaches for improving survival of these rails. Given the particularly low survival of rails released later in the season, juveniles from late-hatched broods could be used to experiment with soft release methods and pre-release conditioning methods to evaluate whether such methods could be used to improve post-release survival.

Our findings indicate a clear need to increase survival of captive-released rails to improve the efficacy of this conservation translocation program. Additional research is needed to better understand behaviors, habitat selection, movements, and survival of these captive rails and especially how it changes through time once released.

Efforts to increase persistence of light-footed Ridgway’s rails face a multitude of complicated management issues. Further efforts to document factors that influence the post-release survival of this species are essential for making informed management decisions now and into the future. Those efforts would also help inform methods for other related species for which captive-released birds are released into the wild to bolster the species’ persistence. A captive breeding and translocation program has the potential to aid populations and maintain genetic diversity of the species but continued post-release monitoring is essential for improving the translocation program.

LITERATURE CITED

Albertson, J. 1995. Ecology of the California clapper rail in south San Francisco Bay. Thesis, San Francisco State University, San Francisco, California, USA.

Barron, D. G., J. D. Brawn, and P. J. Weatherhead. 2010. Meta-analysis of transmitter effects on avian behaviour and ecology. Methods in Ecology and Evolution 1:180-187. https://doi.org/10.1111/j.2041-210X.2010.00013.x

Beck, B. B., L. G. Rapaport, M. R. Stanley Price, and A. C. Wilson. 1994. Reintroduction of captive-born animals. Pages 265-286 in P. J. S Olney, G. M. Mace, and A. T. C. Feistner, editors. Creative conservation: interactive management of wild and captive animals. Springer Netherlands, Dordrecht, The Netherlands. https://doi.org/10.1007/978-94-011-0721-1_13

Berger-Tal, O., and D. Saltz. 2014. Using the movement patterns of reintroduced animals to improve reintroduction success. Current Zoology 60:515-526. https://doi.org/10.1093/czoolo/60.4.515 https://doi.org/10.1093/czoolo/60.4.515

Berger-Tal, O., J. Nathan, E. Meron, and D. Saltz. 2014. The exploration–exploitation dilemma: a multidisciplinary framework. PLoS ONE 9:e95693. https://doi.org/10.1371/journal.pone.0095693

Berger-Tal, O., D. T. Blumstein, and R. R. Swaisgood. 2020. Conservation translocations: a review of common difficulties and promising directions. Animal Conservation 23:121-131. https://doi.org/10.1111/acv.12534

Brophy, L. S., C. M. Greene, V. C. Hare, B. Holycross, A. Lanier, W. N. Heady, K. O’Connor, H. Imaki, T. Haddad, and R. Dana. 2019. Insights into estuary habitat loss in the western United States using a new method for mapping maximum extent of tidal wetlands. PLoS ONE 14:e0218558. https://doi.org/10.1371/journal.pone.0218558

Bubac, C. M., A. C. Johnson, J. A. Fox, and C. I. Cullingham. 2019. Conservation translocations and post-release monitoring: identifying trends in failures, biases, and challenges from around the world. Biological Conservation 238:108239. https://doi.org/10.1016/j.biocon.2019.108239

Burger, L. W., M. R. Ryan, D. P. Jones, and A. P. Wywialowski. 1991. Radio transmitters bias estimation of movements and survival. The Journal of Wildlife Management 55:693-697. https://doi.org/10.2307/3809520

Burnham, K. P., and D. R. Anderson. 1998. Model selection and inference: a practical information-theoretical approach. Springer-Verlag, New York, New York, USA. https://doi.org/10.1007/978-1-4757-2917-7

Calenge, C. 2006. The package “adehabitat” for the R software: a tool for the analysis of space and habitat use by animals. Ecological Modelling 197:516-519. https://doi.org/10.1016/j.ecolmodel.2006.03.017

Casazza, M. L, C. T. Overton, T.D. Bui, J. Y. Takekawa, A. M. Merritt, and J. M. Hull. 2016. Depredation of the California Ridgway’s rail: causes and distribution. Proceedings of the Vertebrate Pest Conference 27:226-235. https://doi.org/10.5070/V427110553

Devineau, O., W. L. Kendall, P. F. Doherty, T. Shenk, G. C. White, P. Lukacs, and K. P. Burnham. 2014. Increased flexibility for modeling telemetry and nest-survival data using the multistate framework. Journal of Wildlife Management 78:224-230. https://doi.org/10.1002/jwmg.660

Eddleman, W. R., and C. J. Conway. 2020. Ridgway’s Rail (Rallus obsoletus), version 1.0. In P. G. Rodewald, editor. Birds of the world. Cornell Lab of Ornithology, Ithaca, New York, USA. https://doi.org/10.2173/bow.ridrai1.01

Ferraro, P.J., and S. K. Pattanayak. 2006. Money for nothing? A call for empirical evaluation of biodiversity conservation investments. PLoS Biology 4:e105. https://doi.org/10.1371/journal.pbio.0040105

Fink, D., T. Auer, A. Johnston, M. Strimas-Mackey, S. Ligocki, O. Robinson, W. Hochachka, L. Jaromczyk, A. Rodewald, C. Wood, I. Davies, and A. Spencer. 2022. eBird status and trends, data version 2021. Cornell Lab of Ornithology, Ithaca, New York, USA. https://doi.org/10.2173/ebirdst.2021

Fischer, J., and D. B. Lindenmayer. 2000. An assessment of the published results of animal relocations. Biological Conservation 96:1-11. https://doi.org/10.1016/S0006-3207(00)00048-3

Foin, T. C., and J. L. Brenchley-Jackson. 1991. Simulation model evaluation of potential recovery of endangered light-footed clapper rail populations. Biological Conservation 58:123-148. https://doi.org/10.1016/0006-3207(91)90116-Q

Getz, W. M., S. Fortmann-Roe, P. C. Cross, A. J. Lyons, S. J. Ryan, and C. C. Wilmers. 2007. LoCoH: nonparameteric kernel methods for constructing home ranges and utilization distributions. PLoS ONE 2:e207. https://doi.org/10.1371/journal.pone.0000207

Griffith, B., J. Scott, J. Carpenter, and C. Reed. 1989. Translocation as a species conservation tool: status and strategy. Science 245:478-480. https://doi.org/10.1126/science.245.4917.477

Harrity, E. J., and C. J. Conway. 2020. Noose carpets: a novel method to capture rails. Wildlife Society Bulletin 44:15-22. https://doi.org/10.1002/wsb.1068

International Union for Conservation of Nature (IUCN). 2013. Guidelines for reintroductions and other conservation translocations. IUCN Species Survival Commission, Gland, Switzerland.

Laake, J. 2022. RMark: R code for MARK analysis. https://cran.r-project.org/package=RMark

Lewis, J. C., and R. L. Garrison. 1983. Habitat suitability index models: clapper rail. United States Fish and Wildlife Service, Washington, D.C., USA.

Lewis, T. L., D. Esler, B. D. Uher-Koch, R. D. Dickson, E. M. Anderson, J. R. Evenson, J. W. Hupp, and P. L. Flint. 2017. Attaching transmitters to waterbirds using one versus two subcutaneous anchors: retention and survival trade-offs. Wildlife Society Bulletin 41:691-700. https://doi.org/10.1002/wsb.833

Massey, B. W., R. Zembal and P. D. Jorgensen. 1984. Nesting habitat of the light-footed clapper rail in southern California. Journal of Field Ornithology 55:67-80.

Mong, T. W., and B. K. Sandercock. 2007. Optimizing radio retention and minimizing radio impacts in a field study of upland sandpipers. Journal of Wildlife Management 71:971-980. https://doi.org/10.2193/2005-775

North American Bird Conservation Initiative. 2022. The state of the birds, United States of America, 2022. https://www.stateofthebirds.org/

Office for Coastal Management. 2023. 2009–2011 CA coastal conservancy coastal lidar project. 15 April 2023. https://www.fisheries.noaa.gov/inport/item/48166

Overton, C., M. L. Casazza, J. Y. Takekawa, D. R. Strong, and M. Holyoak. 2014. Tidal and seasonal effects on survival rates of the endangered California clapper rail: does invasive Spartina facilitate greater survival in a dynamic environment? Biological Invasions 16:1897-1914. https://doi.org/10.1007/s10530-013-0634-5

R Core Team. 2023. R: a language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. https://www.R-project.org/

Resende, P. S., A. B. Viana-Junior, R. J. Young, and C. S. D. Azevedo. 2020. A global review of animal translocation programs. Animal Biodiversity and Conservation 221-232. https://doi.org/10.32800/abc.2020.43.0221

Richards, S. A. 2005. Testing ecological theory using the information-theoretic approach: examples and cautionary results. Ecology 86:2805-2814. https://doi.org/10.1890/05-0074

Richards, S. A., M. J. Whittingham, and P. A. Stephens. 2010. Model selection and model averaging in behavioural ecology: the utility of the IT-AIC framework. Behavioural Ecology and Sociobiology 65:77-89. https://doi.org/10.1007/s00265-010-1035-8

Roberts, J. L., and D. Luther. 2023. An explanatory analysis of behavior-based and other management techniques to improve avian conservation translocations. Biological Conservation 279:1099941. https://doi.org/10.1016/j.biocon.2023.109941

Rosenberg, K. V., A. M. Dokter, P. J. Blancher, J. R. Sauer, A. C. Smith, P. A. Smith, J. C. Stanton, A. Panjabi, L. Helft, M. Parr, and P. P. Marra. 2019. Decline of the North American avifauna. Science 366:120-124. https://doi.org/10.1126/science.aaw1313

Rosencranz, J. A., K. M. Thorne, K. J. Buffington, C. T. Overton, J. Y. Takekawa, M. L. Casazza, J. McBroom, J. K. Wood, N. Nur, R. L. Zembal, G. M. MacDonald, and R. F. Ambrose. 2019. Rising tides: assessing habitat vulnerability for an endangered salt marsh-dependent species with sea-level rise. Wetlands 39:1203-1218. https://doi.org/10.1007/s13157-018-1112-8

San Diego Geographic Information Source. 2015. Western San Diego County vegetation. 15 April 2023. http://www.sangis.org/download/index.html

San Diego Geographic Information Source. 2023. San Diego County roads and trails. 15 April 2023. http://www.sangis.org/download/index.html

Seddon, P. J. 1999. Persistence without intervention: assessing success in wildlife reintroductions. Trends in Ecology and Evolution 14:503. https://doi.org/10.1016/S0169-5347(99)01720-6

Seddon P. J., D. P. Armstrong, and R. F. Maloney. 2007. Developing the science of reintroduction biology. Conservation Biology 21:303-312. https://doi.org/10.1111/j.1523-1739.2006.00627.x

Severson, J. P., P. S. Coates, B. G. Prochazka, M. A. Ricca, M. L. Casazza, and D. J. Delehanty. 2019. Global positioning system tracking devices can decrease greater sage-grouse survival. The Condor 121:1-15. https://doi.org/10.1093/condor/duz032

Stamps, J. A., and R. R. Swaisgood. 2007. Someplace like home: experience, habitat selection and conservation biology. Applied Animal Behaviour Science 102:392-409. https://doi.org/10.1016/j.applanim.2006.05.038

Stein, E. D., K. Cayce, M. Salomon, D. L. Bram, D. De Mello, R. Grossinger, and S. Dark. 2014. Wetlands of the southern California coast—historical extent and change over time. Southern California Coastal Water Research Project Technical Report 826, San Francisco Estuary Institute, Costa Mesa, California, USA.

Stein, B. A., N. Edelson, L. Anderson, J. Kanter, and J. Stemler. 2018. Reversing America’s wildlife crisis: securing the future of our fish and wildlife. National Wildlife Federation, Washington, D.C., USA.

Strimas-Mackey, M., S. Ligocki, T. Auer, and D. Fink. 2022. ebirdst: tools for loading, plotting, mapping and analysis of eBird status and trends data products. R package version 1.2021.0. https://cornelllabofornithology.github.io/ebirdst/

Sutherland, W. J., I. Newton, and R. Green. 2004. Bird ecology and conservation: a handbook of techniques, volume 1. Oxford University Press, New York, New York, USA. https://doi.org/10.1093/acprof:oso/9780198520863.001.0001

Thorne, K., G. MacDonald, G. Guntenspergen, R. Ambrose, K. Buffington, B. Dugger, C. Freeman, C. Janousek, L. Brown, J. Rosencranz, J. Holmquist, J. Smol, K. Hargan, and J. Takekawa. 2018. U.S. Pacific coastal wetland resilience and vulnerability to sea-level rise. Science Advances 4:eaao3270. https://doi.org/10.1126/sciadv.aao3270

U.S. Fish and Wildlife Service. 1985. Recovery plan for the light-footed clapper rail. U.S. Fish and Wildlife Service, Portland, Oregon, USA.

Wanless, R. M., J. Cunningham, P. A. R. Hockey, J. Wanless, R. W. White, and R. Wiseman. 2002. The success of a soft-release reintroduction of the flightless Aldabra rail (Dryolimnas [cuvieri] aldabranus) on Aldabra Atoll, Seychelles. Biological Conservation 107:203-210. https://doi.org/10.1016/S0006-3207(02)00067-8

White, G. C., and K. P. Burnham. 1999. Program MARK: survival estimation from populations of marked animals. Bird study 46.sup1:S120-S139. https://doi.org/10.1080/00063659909477239

Wilcove, D. S., D. Rothstein, J. Dubow, A. Phillips, and E. Losos. 1998. Quantifying threats to imperiled species in the United States. BioScience 48:607-615. https://doi.org/10.2307/1313420

Wolf, C. M., B. Griffith, C. Reed, and S. A. Temple. 1996. Avian and mammalian translocations: update and reanalysis of 1987 survey data. Conservation Biology 10:1142-1154. https://doi.org/10.1046/j.1523-1739.1996.10041142.x

Wolf, C. M., T. Garland, and B. Griffith. 1998. Predictors of avian and mammalian translocation success: reanalysis with phylogenetically independent contrasts. Biological Conservation 86:243-255. https://doi.org/10.1016/S0006-3207(97)00179-1

Zedler, J. B. 1993. Canopy architecture of natural and planted cordgrass marshes: selecting habitat evaluation criteria. Ecological Applications 3:123-138. https://doi.org/10.2307/1941796

Zembal, R., and J. M. Fancher. 1988. Foraging behavior and foods of the light-footed clapper rail. Condor 90:959-962. https://doi.org/10.2307/1368864

Zembal, R., and S. M. Hoffman. 2022. Light-footed Ridgway’s (clapper) rail in California, 2022. Report to U.S. Fish and Wildlife Service and California Department of Fish and Wildlife. https://doi.org/10.13140/RG.2.2.26330.31684

Zembal, R., and B. W. Massey. 1987. Seasonality of vocalizations by light-footed clapper rails. Journal of Field Ornithology 58:41-48.

Zembal, R., B. W. Massey and J. M. Fancher. 1989. Movements and activity patterns of the light-footed clapper rail. Journal of Wildlife Management 53:39-42. https://doi.org/10.2307/3801302

Zetler, B. D., and R. E. Flick. 1985. Predicted extreme high tides for mixed-tide regimes. Journal of Physical Oceanography 15:357-359. https://doi.org/10.1175/1520-0485(1985)015%3C0357:PEHTFM%3E2.0.CO;2