Urban Stream Syndrome and Contaminant Uptake in Salamanders of Central Texas

The conversion of natural intact ecosystems to other land uses, such as agriculture, industry, and urban development, is one of the greatest anthropogenic forces exerted on natural resources worldwide (Seto et al. 2012). As native habitats such as bottomland hardwood forests or tall grass prairies are cleared for other purposes, a series of changes occurs to the physical, chemical, and biological features of the habitat that alter a myriad of natural ecological functions. These alterations to ecological functions are permanent and irreversible, in most cases, without carefully planned restoration designed to return the areas back to natural landscapes (Booth et al. 2016). Urban development, for example, typically involves a substantial increase in impervious cover (IC) as large areas are paved with concrete or asphalt, which alters the hydrologic function and physical properties of affected watersheds (Walsh et al. 2005). Urbanization is a major contributor of contaminant loading to water bodies throughout the United States (Booth and Jackson 1997; Chadwick et al. 2006). Contaminants commonly found in urban streams include fertilizers, pesticides, petroleum hydrocarbons, solid waste, constituents associated with municipal wastewater effluents, and suspended solids (Kolpin et al. 2002; Chalmers et al. 2007). These contaminants degrade water quality, causing a wide variety of problems ranging from increased nutrient loading to elevated concentrations of insecticides or trace metals. Trace metals can be acutely toxic to aquatic organisms, resulting in a loss of biodiversity (Brown et al. 2009) and ecosystem function (Riva-Murray et al. 2010). Urban threats to water quality can originate from point sources (i.e., end of pipe), nonpoint sources (e.g., stormwater runoff), and catastrophic events (e.g., diesel fuel spill). They include contaminants related to municipal and industrial wastewater effluents, hazardous chemicals from leaking underground storage tanks, oil spills from pipelines or refineries, bacterial pathogens from septic tanks, and sediments from construction activities.

The Edwards Aquifer is a karst aquifer that drains over 10,000 km² in Central Texas, serves as a main water supply for over 2.3 million people, and is home to 23 species listed as threatened or endangered under the U.S. Endangered Species Act (ESA 1973, as amended; NAS 2016). In Central Texas, there are several endemic taxa that are geographically restricted (Longley 1981; Bowles and Arsuffi 1993; Chippindale et al. 1993); some are endemic to single spring systems (Miller et al. 2009) and are thus highly vulnerable to natural and anthropogenic disturbances. Metropolitan areas in the Edwards Aquifer region have experienced some of the highest population growth rates in the United States. (U.S. Census Bureau 2006). Areas with high population density have been shown to have pulses of increased discharge from urban flows (stormwater and wastewater) into the recharge zones, which degrades the water quality of the aquifer (Jagucki et al. 2011) and alters the health of spring-fed stream habitats through increased contaminant loading. These changes to water quality are of concern for the long-term health of the Edwards Aquifer and the long-term survival of the endemic species that evolved in pristine waters with stable temperatures and chemical properties.

Although threats to local aquatic systems are increasing because of extensive development in areas that overlie the aquifer, scientific literature on the bioaccumulation and effects of contaminants on local stream communities specific to Central Texas is limited. Nonetheless, the sensitivity of aquatic life to urbanization, even at low levels, and the association between urban development and loss of aquatic biodiversity is well established and has been documented at many trophic levels in urbanized streams (Paul and Meyer 2001; Walsh et al. 2005). In particular, the negative relationship of aquatic invertebrates (Moore and Palmer 2005; Cuffney et al. 2011) and salamanders (Price et al. 2006) to their environment in urbanized areas throughout North America has been well documented. Several different species of stream salamanders showed decreased abundance with increasing urbanization in watersheds located in Georgia (Orser and Shure 1972), North Carolina (Price et al. 2012), Maryland and Virginia (Grant et al. 2009), and Central Texas (Bendik et al. 2014).

As with other amphibians, salamanders have experienced global declines (Hayes et al. 2010) and are believed to be particularly vulnerable to chemical stressors because of their inherently close association with water throughout their life cycle and their ability to absorb chemicals through their gills, semipermeable skin, and unshelled eggs (Boyer and Grue 1995). Early life stages of Eurycea are more sensitive for the above reasons, and therefore, most at risk from the effects of environmental pollution. Mercury, a known neurotoxin, has been shown to adversely affect behavior and feeding performance of the two-lined salamander Eurycea bislineata (Burke et al. 2010). Salamander food items such as amphipods and other crustaceans are also sensitive to contaminants (Burton and Ingersoll 1994), so a decline in prey abundance could presumably affect Eurycea spp. survival, growth, or reproductive success. Bodies of water that receive urban runoff typically exhibit lower invertebrate species diversity and abundance, with a disproportionate number of pollution-tolerant organisms (Coles et al. 2010). Therefore, Eurycea spp. might be both directly affected by the toxicity of impaired waters and indirectly affected by the subsequent loss of prey diversity and reduced forage success as dietary organisms succumb to the acute and chronic effects of xenobiotics. Chronic exposures may result in bioaccumulation of certain contaminants in Eurycea spp. tissues. The relationship between bioaccumulation of contaminants and varying levels of IC has not been examined in Texas Eurycea spp. before this study.

Within the genus Eurycea there are seven species protected under the Endangered Species Act (ESA 1973), along with many endemic taxa that occur in the Edwards Aquifer. This study examined relationships between urbanization and contaminants in 20 catchments in the northern portion of the Edwards Aquifer and how they relate to aquatic invertebrate abundance, Eurycea spp. contaminant loads (i.e., body burden), as well as implications for conservation and recovery of these rare endemic species. Here we assess bioaccumulation of contaminants in populations of Central Texas Eurycea spp., providing empirical data on how these populations may respond to future anthropogenic changes to the environment. We examined the concentration of bioaccumulative and persistent pollutants (organic and inorganic chemicals) directly in Eurycea spp. and in spring water to determine their relationship with changing land uses and urbanization, and assessed the diversity and abundance of aquatic invertebrates present at the sites.

We conducted this study at springs located throughout the Edwards Aquifer region of Central Texas (Figure 1; Data S1, Supplemental Material). We selected field sites for Eurycea spp. collections on the basis of ease of access, the abundance of Eurycea at a particular site, and the levels of anthropogenic land use in the area. We created catchments using the National Hydrography Dataset Plus. We collected percent IC and land-use data from the National Land Cover Database 2011 data sets (Xian et al. 2011; Homer et al. 2015). We created percent IC by clipping the IC layer for a specific catchment, then exporting the data into an excel file and taking the weighted average of the scores within that catchment. Using the results of this analysis, we assigned catchments a percent IC score (i.e., 0–100%), with low percentages indicating little or no IC and high percentages representing more IC within the catchment. Using the percent IC scores, we portioned sites into either developed (> 10% IC) or nondeveloped sites (≤ 10% IC; Bowles et al. 2006). We took the average age of urban development in an area from published literature when available (Bendik et al. 2014); otherwise we collected parcel data from each county and then clipped them in ArcGIS 10.2 and averaged them for each catchment.

We collected field data primarily from April to July in 2013 and 2014, although we made exceptions when flooding prevented access to certain sites, in which case we collected samples at a later date (i.e., PC Spring [invertebrates], Hog Hollow [water], and Cowan [water and salamanders]). We collected salamanders opportunistically each year on the basis of available populations that allowed the removal of individual salamanders from selected sites during field visits. We caught salamanders using a small handheld aquarium dip net after turning over rocks, cobble, and other cover in the streambed. We collected basic water-quality parameters such as dissolved oxygen, pH, temperature, and conductivity using a Hydrotech Compact DS5 multiparameter probe instrument.

We collected aquatic invertebrates at spring sites using a Surber sampler. We targeted single riffles with cobble and gravel substrates for sample collection, as they represent ideal habitat and were most likely to yield diverse samples of the resident invertebrate community (Texas Commission on Environmental Quality 2014, Appendix F). We collected and preserved all samples on site, then sorted and identified them in the lab. We made identifications to the genus level, except specimens from the Chironomid family, which we took to subfamily or tribe. We used metrics such as total taxa (total number of unique specimens), percent Chironomidae (vs. total), and ratio of dominant taxa (e.g., a preponderance of pollution-tolerant taxa indicates a community out of balance because of perturbation) to determine the level of disturbance at sampled sites. Using the metrics and scoring criteria established by Texas Commission on Environmental Quality surface water quality standards (vol. II) for Surber samples (TCEQ 2014), we calculated an aquatic life use score for each site. The Texas Commission on Environmental Quality protocol then defines qualitative ranks (or site descriptions) on the basis of the following point score ranges: a score of < 21 was described as “limited” in terms of diversity (i.e., poor condition), 21–30 “intermediate,” 31–40 “high,” and > 40 “exceptional” (i.e., pristine or near-pristine condition).

Because of the small size of the salamanders and the amount of tissue needed to conduct the chemical analysis, we composited individual samples by site. The number of individuals selected for each composite sample was dependent upon the status of the salamanders present at a particular site and capture success. Each composite sample consisted of only one species. We stored salamanders used to create the composite sample in a freezer and sent in an ice chest packed with ice for processing at the U.S. Geological Survey Columbia Environmental Research Center. Composite whole-body tissue samples comprised three to fve homogenized individual salamanders. We individually weighed, minced, composited, and homogenized whole salamanders to obtain sufficient sample mass to meet the analytical requirements. We subsampled each salamander composite for processing and analysis. Methods for the processing and analysis of the salamander samples followed established procedures for fish tissues as described by Gale et al. (2009). We analyzed samples for contaminant residues, including 19 trace metals, 29 organochlorine compounds (OC), 125 polychlorinated biphenyl (PCB) congeners and total PCBs, and 9 polybrominated diphenyl ethers (PBDEs or flame retardants) and reported them in nanograms per gram. We define total body burden here as the sum in nanograms per gram of all chemicals (minus metals) that we detected in salamander tissues.

To measure contaminants present in the water column, we used two types of passive water-quality samplers. The first type, known as a semipermeable membrane device (pg/L), collects the following fat-soluble, hydrophobic compounds: 31 OC pesticides, total PCBs, 5 PBDEs, and 33 polycyclic aromatic hydrocarbons (Alvarez 2010). The second device, known as a polar organic chemical integrative sampler (ng/L), samples water-soluble, hydrophilic compounds (Alvarez 2010); we deployed that sampler alongside the SPMD. The polar organic chemical integrative samplers collect 46 types of pollutants associated with a variety of contaminant sources, such as plasticizers, surfactants, fungicides, antimicrobials, flame retardants, disinfectants, caffeine, fragrances, and pharmaceuticals, many of which can be associated with sewage and municipal wastewater (Table S1, Supplemental Material). Methods for processing and analysis of the passive samplers followed standard procedures as previously described (Alvarez et al. 2008; Alvarez et al. 2012), with specific details given in this manuscript. We deployed passive samplers at 18 sites during a 45- to 55-d period at each site over the 2-y study period. In 2013, we sampled 7 sites and in 2014, 15 sites, with some duplicates from the previous year due to the lack of site access (Tables S2 and S3, Supplemental Material).

To examine relationships and significant differences among IC, average year developed, and response variables, we conducted linear regression or single-factor analysis of variance (ANOVA) tests (Aho 2014). Response variables were the levels of contaminants detected in either salamander tissue or the passive samplers. For contaminants collected in the water column using passive samplers, and in salamander tissues, we analyzed both the number of unique xenobiotics detected in a sample (count) and the concentration detected (sum of contaminants detected either in picograms per liter or nanograms per gram respectively). We used Cook's distance to identify potential outliers in linear regression analyses (Li 1985). When we detected a relationship for the complete data set but did not detect a relationship when we removed an identified outlier, then we reported findings for the latter, not the former (i.e., we implied no relationship). When necessary, we natural-log-transformed the response variable to meet assumptions of homoscedasticity or normality. Because some values were zeros, we added 1.0 to all values before log transformations. For the trace-metal data, we used nonmetric multidimensional scaling in two dimensions to examine similarities among sites and associations of metals within salamander whole-body composite samples from specific sites (Selago et al. 2011). We used the R package vegan with the code “metaMDS” for the analysis. We determined the significance level for all analyses at an alpha of P ≤ 0.05.

Impervious cover scores ranged from 0 to 43% for the data set (Table 1). Although some studies suggest that environmental impacts occur at IC levels below 10%, for this study we defined these areas as nondeveloped. There were 7 sites in the nondeveloped range (≤ 10%) and 13 sites in the developed range (> 10%) on the basis of percent IC. The Spicewood site had the highest IC score of 43%, followed by Brushy Creek, with 35%. The Spicewood site also had the highest urban developed land at 93%, followed by Barton Springs with 83% (MRLC 2011). Land uses varied within catchments where sample sites were located and are presented in Table S4 (Supplemental Material). Basic water chemistry collected at each site is presented in Table S5 (Supplemental Material).

We collected aquatic invertebrate samples at 18 of 20 sites because of the lack of riffle habitat at some spring sites. Aquatic life-use scores ranged from limited to exceptional (Table 1). The Troll and Spicewood sites scored as limited. A total of seven sites were in the intermediate range. Seven sites scored in the high range. Tributary 6 and Brushy Creek both scored as exceptional. There were some clear trends associated with the aquatic life use data when examined using linear regression, with percent Chironomidae having a statistically significant positive relationship and percent dominant taxa being marginally nonsignificant (P = 0.056; Figure 2). There were noticeable trends associated with other calculated metrics; we observed both positive and negative responses in relation to IC scores, although they were not statistically significant.

We collected 14 composite salamander tissue samples representing three species from 10 sites (Table S6, Supplemental Material). Results from one-way ANOVAs revealed significant differences between developed (> 10% IC) and nondeveloped (≤ 10% IC) sites at the catchment level for OCs and PBDEs for count and PBDEs for concentration (Table 2). When we examined the relationships between the average age of urban development and contaminants in salamander tissues using linear regression, there were significant positive relationships for both counts and concentrations of OCs, PBDEs, and total body burdens with age of urban development, whereas only concentration of PCBs was significant. (Figure 3). There were no significant relationships for the trace-metals data set.

Although toxicity data for salamanders are generally lacking and their sensitivity to trace metals is largely unknown, selenium residues in salamanders collected from the Barrow site (8 μg/g dry weight) exceeded thresholds designed to protect larval fish (4 μg/g dry weight) and piscivorous wildlife (3 μg/g dry weight; Lemly 1996, 2002). Analysis using nonmetric multidimensional scaling showed specific trace metals accumulated in composite salamander tissue samples associated with certain sites (stress = 0.11; Figure 4). Manganese was associated with three sites on the positive side of the nonmetric multidimensional scaling plot (i.e., Blizzard, House, and Cowan). Although other sources of Mn exist, each of these sites has a golf course close to the springs that may serve as a source of Mn, which is a common component of commercially available chemical fertilizers and fungicides used for turf management.

Single-factor ANOVA for water-quality data from the semipermeable membrane devices and polar organic chemical integrative sampler samples showed a significant difference between contaminant counts at sites with > 10% IC vs. sites with ≤ 10% IC. The analysis was significant for polycyclic aromatic hydrocarbons, OCs, total number of contaminants, and total contaminants without associated wastewater contaminants collected by the polar organic chemical integrative sampler in the count (Table 3). None of the contaminants detected was above any state or federal thresholds.

Within the catchments we studied there was a positive correlation between the level of IC present and the contaminants detected in salamander tissues and surface water. There were significant differences observed in water quality and contaminants detected in composite salamander tissue samples between urban developed and nondeveloped sites. As IC increased in the catchments, both the concentration and count of chemical contaminants present within salamander tissues and the water column increased. We also observed the influence of IC in the composition of benthic aquatic invertebrate communities, which followed a classic response to urbanization (i.e., decreasing aquatic invertebrate diversity with increasing IC). Although there was large variation between the sites, the effects of IC and other artifacts of urban stream syndrome were present (Walsh et al. 2005) and the positive trend associated with the relationship between IC and contaminants was present in all analyses, although not always statistically significant.

In addition, the relationship between the age of urban development and total body burden of contaminants in salamander tissues was significantly positively correlated. This relationship could be explained by two different factors. First, the longer an area has been developed, the greater is the potential for disturbance to occur from anthropogenic activities. Second, best management practices used either by choice or through regulation may be serving to reduce contaminant loading in recently built areas compared with older, less regulated neighborhoods.

Our study confirms the presence of a variety of contaminants within salamander habitats, with springs located closer to intense urban activity being more likely to have complex mixtures of contaminants present at higher concentrations. Although our study analyzed only a small subset of thousands of potentially harmful substances that might be present in the environment, and would not include many pollutants that are rapidly metabolized or excreted by salamanders or that rapidly degrade in the environment but are nonetheless capable of causing harm to biota, we conclude that exposure to xenobiotics results in both the accumulation of contaminants in salamander tissues and a reduction in the diversity of aquatic invertebrates that occupy springs. We chose the contaminants selected for this study because of their ubiquitous presence in the environment and the established analytical methods used to determine presence and concentrations. However, the variables we assessed in this manuscript are just a few of many effects of urbanization that could act individually or in combination to negatively affect salamanders in the rapidly changing environments of urban landscapes. The changes associated with urbanization often have drastic negative effects on salamanders at the population level (Bank et al. 2006; Price et al. 2012a). Locally, Bendik et al. (2014) showed a negative effect of development on densities of Eurycea tonkawae from 17 sites over a 4-y period. In addition, their trend analysis showed that salamander declines were correlated with increasing development over a 15-y period.

Although we detected contaminants at relatively low concentrations in the spring systems we examined, many pollutants are capable of eliciting a biological response at low levels (e.g., endocrine disruptors, hormonally active compounds) and other factors can combine to create a combination of disturbances to spring systems, the sum of which is greater than the individual stressors. For example, stronger pulses of stormwater due to channelization and IC in a catchment cause not only an increase in the transport and deposition of point- and nonpoint-source pollution within watersheds, but also cause 1) a rapid increase in discharge in creeks and rivers associated with spring systems (Booth and Jackson 1997), 2) changes to the functionality of riparian corridors (Booth and Jackson 1997) that potentially alter trophic dynamics (Wallace et al. 1997; Chadwick et al 2006), 3) changes to the community structure of aquatic invertebrates (Cuffney et al. 2010) used as prey by salamanders, and 4) increased sedimentation (Finkenbine et al. 2000). The contaminants and other ecological processes associated with urbanization may reduce salamander habitat quality to the point that individual reproductive success is diminished, leading to population declines over generations. Since site-specific factors can influence the level of stress urbanization exerts at any given site, the rate of decline in biodiversity will vary by spring and watershed. The unique hydrology of some springs may afford protection when flows are consistently high enough to dilute potential contaminants and provide a buffer against biological effects seen at comparable springs with lower base flows.

At all sites sampled for this study, salamander tissues accumulated locally available persistent chemical contaminants that were present in their respective catchments. Although the U.S. Fish and Wildlife Service has posited the creation of 300-m buffers around springs to protect critical habitat for federally endangered or threatened salamanders (USFWS 2013b), the results of this study suggest that a more catchment-wide approach is warranted, one that affords greater protection on a landscape scale because of the unknown flow paths of a karst system. Given the complex hydrology of the Edwards Aquifer, it is not surprising that water quality at a given spring can be influenced by anthropogenic activities miles away (Hunt et al. 2005). Although buffer zones and other best-management practices provide some benefit to water quality, they may fall short of preventing the decline of rare endemic species like Eurycea salamanders. We believe that the practice of establishing protected areas, such as nature preserves that remain undeveloped, coupled with directing urban growth away from environmentally sensitive areas, will better protect salamanders of the Edwards Aquifer. If these conservation measures are pursued along with additional research and monitoring, the level of sufficient protection should become evident when populations that were once in decline may begin to stabilize or increase in size. Conversely, the status quo, with no changes to urban growth, conservation, or watershed management, may result in the continued decline of these imperiled species that are unique to springs of the Edwards Aquifer.

Please note: The Journal of Fish and Wildlife Management is not responsible for the content or functionality of any supplemental material. Queries should be directed to the corresponding author for the article.

Data S1. This data access file contains land-use calculations on impervious cover, contaminant results regarding the water and salamander Eurycea spp. tissues, and invertebrate data.

Found at DOI: https://doi.org/10.3996/032018-JFWM-017.S1 (113 KB XLSX).

Table S1. Chemicals targeted for analysis in water samples collected from the northern Edwards Plateau in 2013 and 2014 using a semipermeable membrane device and polar organic chemical integrative sampler. MDL = method detection limit; MQL = method quantitation limit; OC = organochlorine; PAHs = polycyclic aromatic hydrocarbons; PCBs = polychlorinated biphenyls; PBDEs = polybrominated diphenyl ethers.

Found at DOI: https://doi.org/10.3996/032018-JFWM-017.S2 (19 KB DOCX).

Table S2. Contaminants (count) detected in the water column using semipermeable membrane devices and polar organic chemical integrative samplers deployed at northern Edwards Plateau spring sites in 2013 and 2014. Numbers represent unique contaminants that we individually detected at the site. We sampled some sites each year. PBDE = polybrominated diphenyl ethers; PCB = polychlorinated biphenyl; PAH = polycyclic aromatic hydrocarbons; Trib = tributary.

Found at DOI: https://doi.org/10.3996/032018-JFWM-017.S3 (17 KB DOCX).

Table S3. Contaminants (concentration) detected in the water column using semipermeable membrane devices (pg/L) and polar organic chemical integrative samplers (wastewater; ng/L) deployed at northern Edwards Plateau spring sites in 2013 and 2014. Numbers represent unique contaminants that we individually detected at the site. We sampled some sites each year. PBDE = polybrominated diphenyl ethers; PCB = polychlorinated biphenyl; PAH = polycyclic aromatic hydrocarbons; Trib = tributary.

Found at DOI: https://doi.org/10.3996/032018-JFWM-017.S4 (23 KB DOCX).

Table S4. Land-use analysis for each spring site along the northern Edwards Plateau taken from the multiresolution land characteristics consortium. We completed spatial designations for each site using the National Hydrography Dataset (NHD) Plus at the catchment level. Values shown represent the percent land use or cover type present within the catchment for the spring analyzed (the sum of which equals 100%); Abbreviations: 2011IC = impervious cover from 2011; OW = open water; DOS = developed open space; DL = developed low intensity; DM = developed medium intensity; DH = developed high intensity; BL = barren land; DF = deciduous forest; EF = evergreen forest; MF = mixed forest; SS = shrub/scrub; GH = grassland/herbaceous; PH = pasture/ hay; CC = cultivated crops; WW = woody wetlands; EHW = emergent herbaceous wetlands.

Found at DOI: https://doi.org/10.3996/032018-JFWM-017.S5 (24 KB DOCX).

Table S5. Water-quality data collected from each spring site along the northern Edwards Plateau in central Texas from 2013 to 2014.

Found at DOI: https://doi.org/10.3996/032018-JFWM-017.S6 (24 KB DOCX).

Table S6. Results of analyses conducted on Eurycea whole-body tissue samples collected from the northern Edwards Plateau springs sites in 2013 and 2014. We present data in two forms, concentration in nanograms per gram and the total count of each type of contaminant, with totals provided for all pollutants as “total body burden” (i.e., combined concentrations of all contaminants in nanograms per gram) and “total body burden count” (sum of all individual contaminants detected). JPS = Jollyville Plateau salamander Eurycea tonkawae; GTS = Georgetown salamander Eurycea naufragia; SS = Salado salamander Eurycea chisholmensis; OC = organocholorine compounds; PBDE = polybrominated diphenyl ethers; PCB = polychlorinated biphenyl.

Found at DOI: https://doi.org/10.3996/032018-JFWM-017.S7 (27 KB DOCX).

Reference S1. Alvarez DA. 2010. Guidelines for the use of the semipermeable membrane device (SPMD) and the polar organic chemical integrative sampler (POCIS) in environmental monitoring studies. Page 28 in U.S. Geological Survey, Techniques and Methods 1-D4.

Found at DOI: https://doi.org/10.3996/032018-JFWM-017.S8 (5.97 MB PDF); also available at https://pubs.usgs.gov/tm/tm1d4/pdf/tm1d4.pdf

Reference S2. Burton GA, Ingersoll CG. 1994. Evaluating the toxicity of sediments. The assessment and remediation of contaminated sediments assessment guidance document. EPA/905-B94/002, U.S. Environmental Protection Agency, Chicago, Illinois.

Found at DOI: https://doi.org/10.3996/032018-JFWM-017.S9 (3.74 MB PDF); also available at https://semspub.epa.gov/work/HQ/189664.pdf

Reference S3. Jagucki ML, Musgrove M, Lindgren RJ, Fahlquist L, Eberts SM. 2011. Assessing the vulnerability of public-supply wells to contamination—Edwards aquifer near San Antonio, Texas. Page 6 in U.S. Geological Survey Fact Sheet FS 2011-3142.

Found at DOI: https://doi.org/10.3996/032018-JFWM-017.S10 (5.18 MB PDF); also available at https://pubs.er.usgs.gov/publication/fs20113142

We thank all of the individuals that helped with this study. We thank Diego Araujo, Nate Bendik from the City of Austin, Blake Sissel, Adam Zerrenner, and Dr. Benjamin Pierce of Southwestern University. Funding for this study was provided by the U.S. Fish and Wildlife Service through the Science Support Partnership. We appreciate the numerous edits and flexibility by the reviewers and Associate Editor.

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