Conservation of Cyprinodon nevadensis pectoralis and three endemic aquatic invertebrates in an artificial desert spring refuge located in Ash Meadows National Wildlife Refuge, Nevada

by

Darrick S. Weissenfluh, B.S.

A Thesis

In

BIOLOGY

Submitted to the Graduate Faculty of Texas Tech University in Partial Fulfillment of the Requirements for the Degree of

MASTER OF SCIENCES

Approved

Dr. Gene R. Wilde Chair

Dr. Nancy E. McIntyre

Dr. Richard E. Strauss

Ralph Ferguson Dean of the Graduate School

December, 2010

Copyright 2010, Darrick S. Weissenfluh

Texas Tech University, Darrick S. Weissenfluh, December 2010

ACKNOWLEDGEMENTS I would like to thank my advisor, Dr. Gene R. Wilde, for accepting me as a graduate student and for working with me remotely. I also wish to thank him for

challenging me to communicate more effectively through writing. Dr. Nancy E.

McIntyre and Dr. Richard E. Strauss also provided recommendations and critiques of my

work and were valuable committee members.

Many people unselfishly assisted me with collecting data in what seemed to be

the most extreme Mojave Desert field conditions: Jeff Goldstein, Sam Skalak, Erin

Bradshaw, April Bradshaw, Cristi Baldino, Mark James, Carl Lundblad, and Paula

Booth. Sam Skalak and Cristi Baldino also provided valuable comments and suggestions

concerning my paper and my study. Additionally, Paula Booth and Marie Weissenfluh

interpreted poor writing on numerous data sheets while assisting with data entry. Kathie

Taylor spent countless hours assisting me with GIS maps and database development, for

which I am indebted.

The design of my research site, School Springs, was largely the creation of Rob

Andress and the Ash Meadows Recovery Implementation Team (AMRIT). On numerous

occasions Rob and the AMRIT went out of their way to assist with questions and data

needs related to my research. Sharon McKelvey, Cristi Baldino, and Heather Hundt

worked with Cynthia Martinez to make my position at Ash Meadows National Wildlife

Refuge possible and supported me throughout my collateral duty as a U.S. Fish and

Wildlife Service Student Career Employment Program biologist and graduate student.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Thanks to their support and encouragement I have had the opportunity to assist the

Refuge in obtaining important scientific information while continuing to pursue my

career.

Finally, I am indebted to my wife Marie Weissenfluh who unselfishly allowed me

to pursue this degree and dealt with all my stress and distraction during the process. My daughter, Kaitlyn, is owed countless hours for allowing “Daddy” to spend too much time

away from her. I look forward to spending more time with both of them. My parents

Steve and Diana, brother Shawn, and sister Holly also supported and encouraged me to

continue my education and for that I am grateful.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

TABLE OF CONTENTS

ACKNOWLEDGEMENTS ...... ii

ABSTRACT……………………………………………………………………………..vi

LIST OF TABLES ...... viii

LIST OF FIGURES ...... x

I. INTRODUCTION ...... 1

Literature Cited ...... 6

II. HABITAT ASSOCIATION AND DISTRIBUTION OF ENDANGERED WARM SPRINGS PUPFISH, CYPRINODON NEVADENSIS PECTORALIS, IN SCHOOL SPRINGS REFUGE ...... 9

Introduction ...... 9

Study Area ...... 12

Methods...... 16 Study Design ...... 16 Determination of Life Stages Using Digital Images ...... 18 C. n. pectoralis Habitat Association ...... 20 Physical and Chemical Stream Variables in School Springs Refuge ...... 21

Results ...... 22 Non-native Aquatic Fish and Crayfish Eradication Results ...... 22 Status of C. n. pectoralis in School Springs Refuge ...... 22 C. n. pectoralis Habitat Association in School Springs Refuge ...... 24

Discussion ...... 24 School Springs Refuge Renovation: Was it a Success? ...... 24 Does School Springs Refuge Reduce the Threats to C. n. pectoralis Conservation? ...... 28 Determining C. n. pectoralis Length from Digital Images ...... 30 Recommendations for the Management of School Springs Refuge ...... 31

Literature Cited ...... 34

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III. ENDEMIC AQUATIC INVERTEBRATE DISTRIBUTION AND ASSOCIATION WITH PHYSICAL AND CHEMICAL STREAM PROPERTIES IN SCHOOL SPRINGS REFUGE ...... 51

Introduction ...... 51

Study Area ...... 54

Methods...... 55 Aquatic Invertebrate Distribution in School Springs Refuge ...... 55 Endemic Aquatic Invertebrate Association with Physical and Chemical Stream Properties in School Springs Refuge ...... 57

Results ...... 58 Non-native Aquatic Invertebrate Eradication in School Springs Refuge ...... 58 Endemic Aquatic Invertebrate Translocation ...... 59 Persistence of Endemic Aquatic Invertebrates in School Springs Refuge ...... 59 Distribution and Dispersal of Endemic Aquatic Invertebrates in School Springs Refuge ...... 61 Endemic Aquatic Invertebrate Distribution and Their Association with Chemical and Physical Stream Properties in School Springs Refuge ...... 63

Discussion ...... 64 Evaluating the Translocation Success of Endemic Aquatic Invertebrates in School Springs Refuge ...... 64 Evaluating the Success and Design of School Springs Refuge ...... 68 Management Recommendations ...... 71

Literature Cited ...... 74

A. SCHOOL SPRINGS REFUGE CHANNEL CHARACTERISTICS FOR EACH REACH ...... 100

B. SUMMARY STATISTICS FOR PHYSICAL AND CHEMICAL STREAM CHARACTERISTICS IN EACH REACH FROM SCHOOL SPRINGS REFUGE. ALL SAMPLES FROM 6 MARCH 2009 TO 3 MARCH 2010 WERE COMBINED ...... 101

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ABSTRACT The Warm Springs Complex (WSC) is one of four management units within Ash

Meadows National Wildlife Refuge. It contains six low-discharge warm spring systems

with individual flows ranging from 1.13 x 10-4 to 1.98 x 10-4 cubic meters per second and

spring-source water temperatures ranging from 28o to 33.5oC year round. School

Springs is one component of the WSC and its spring source is the warmest. This spring

has undergone dramatic anthropogenic transformation since at least the 1930s. In 1969

the Bureau of Land Management (BLM) increased pool habitat in School Springs in an

effort to preserve the endangered Warm Springs pupfish, Cyprinodon nevadensis

pectoralis. Four concrete ponds were constructed at School Springs in 1983 to further

increase available habitat to C. n. pectoralis. During the summer of 2008, the School

Springs refuge was completely renovated: the large concrete ponds were removed and a

“naturalized” channel consisting of pools, runs, riffles, and a wash was created. There

were three primary objectives of this renovation: (1) eradicate three aquatic non-native

species including western mosquitofish Gambusia affinis, red swamp crayfish

Procambarus clarkii, and red-rimmed melania Melanoides tuberculatus; (2) improve

amount of suitable habitat for the endangered Warm Springs pupfish and three aquatic invertebrates (P. pisteri, S. c. calida, A. relictus), that are endemic to the WSC; and (3)

test hypotheses concerning endemic fish and invertebrate habitat use and distribution

inherent in the design of the refuge.

Based on my study, two of the aquatic non-native species, G. affinis and P.

clarkii, were successfully eradicated, but M. tuberculatus was not. My results also vi

Texas Tech University, Darrick S. Weissenfluh, December 2010 indicate C. n. pectoralis use pool habitat more frequently than any other habitat type,

regardless of life stage; however, they were captured in all habitat types and the fish may

be distributed throughout the system from the spring source to the wash. Habitat type

was a better predictor of C. n. pectoralis presence than water volume, regardless of the

season, which further supports the importance of creating pool habitat for conservation of

C. n. pectoralis in Ash Meadows.

Endemic aquatic invertebrates were translocated into the upper 20 m of School

Springs refuge and, as of September 2010, continue to persist. The median-gland Nevada

springsnail Pyrgulopsis pisteri is narrowly distributed in the upper 20 m of School

Springs and, therefore, has not dispersed downstream of the translocation site. Both the

Devils Hole warm springs riffle beetle Stenelmis calida calida and the Warm Springs

naucorid Ambrysus relictus are seasonally distributed throughout the stream channel, but

are restricted to the upper 40 m of stream channel during the winter. P. pisteri, S. c.

calida, and A. relictus presence was not associated with substrate type, but their presence

was associated with pool and riffle habitat types; however, the pool habitat in this case

was the spring source. On occasional night visits to School Springs refuge in the summer

of 2009, I observed numerous S. c. calida and A. relictus. These observations suggest

night surveys may be appropriate for monitoring of A. relictus and S. c. calida

populations in School Springs and elsewhere.

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LIST OF TABLES 2.1. Summary of C. n. pectoralis life stages captured (n), adult/juvenile (A/J) ratio, adult density (m-2) and the adult population...... 39 2.2. Number and total length (mm) of C. n. pectoralis captured by habitat type in School Springs refuge...... 40 2.3. Frequency distribution of C. n. pectoralis life stages by reach, estimated area (m2) and estimated water volume (m3) ...... 41 3.1. The total number of each invertebrate family or genus (italized), arranged by abundance, sampled in School Springs refuge, Ash Meadows National Wildlife Refuge, Nevada...... 78 3.2. Contingency table (2 x 2) showing presence-absence frequencies of Pyrgulopsis pisteri in dip net samples from the upper 20 meters of School Springs refuge versus the remainder of the spring. χ2 = 130.27, df = 1, P < 0.01...... 79 3.3. Contingency table (2 x 2) showing presence-absence frequencies of Stenelmis calida calida in dip net samples from the upper 20 meters of School Springs refuge versus the remainder of the spring. χ2 = 26.34, df = 1, P < 0.01...... 80 3.4. Contingency table (2 x 2) showing presence-absence frequencies of Ambrysus relictus in dip net samples from the upper 20 meters of School Springs refuge versus the remainder of the spring. χ2 = 12.29, df = 1, P < 0.01...... 81 3.5. Contingency table (4 x 2) showing presence-absence frequencies of Pyrgulopsis pisteri in dip net samples from School Springs refuge collected in each substrate type. χ2 = 0.16, df = 3, P > 0.05...... 82 3.6. Contingency table (4 x 2) showing presence-absence frequencies of Pyrgulopsis pisteri in dip net samples from School Springs refuge collected in each habitat type. χ2 = 14.65, df = 3, P < 0.01...... 83 3.7. Contingency table (4 x 2) showing presence-absence frequencies of Stenelmis calida calida in dip net samples from School Springs refuge collected in each substrate type. χ2 = 1.02, df = 3, P > 0.05...... 84 3.8. Contingency table (4 x 2) showing presence-absence frequencies of Stenelmis calida calida in dip net samples from School Springs refuge collected in each habitat type. χ2 = 8.06, df = 3, P < 0.05...... 85 3.9. Contingency table (4 x 2) showing presence-absence frequencies of Ambrysus relictus in dip net samples from School Springs refuge collected in each substrate type. χ2 = 5.84, df = 3, P > 0.05...... 86

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3.10. Contingency table (4 x 2) showing presence-absence frequencies of Ambrysus relictus in dip net samples from School Springs refuge collected in each habitat type. χ2 = 14.82, df = 3, P < 0.01...... 87 3.11. Spearman rank correlations, corrected for ties, detailing how Pyrgulopsis pisteri, Stenelmis calida calida, and Ambrysus relictus abundance is correlated with chemical and physical variables in School Springs refuge. Significant P -values are denoted by * P < 0.05 and ** P < 0.01...... 88 4.1. Habitat type, slope, surface area, reach length, mean reach width, and mean reach depth for each reach in School Springs refuge...... 100 4.2. Summary statistics for conductivity, DO, algae density, vegetation density, salinity, TDS, velocity, water depth, water temperature, and pH in each reach of School Springs refuge...... 101

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LIST OF FIGURES 2.1. Warm Springs Complex and School Springs refuge study area within Ash Meadows National Wildlife Refuge, Amargosa Valley, Nevada...... 42 2.2. School Springs refuge, Ash Meadows, Nevada, 1969-2010. A. 1969- The spring terminates in a dug-out pool. B. 1983-Four concrete pools were constructed to improve habitat for C. n. pectoralis at School Springs. C. 2008-School Springs was rehabilitated, which included diversifying the habitats for Pyrgulopsis pisteri, Stenelmis calida, Ambrysus relictus, and C. n. pectoralis. D. 2010-Algae was abundant in the largest pool, reach 19. Photograph credit: U.S. Fish and Wildlife Service...... 43 2.3. School Springs refuge habitat as-built depicting reach segments, as well as pool, run, and riffle habitat types. Note: The wash is not included on this map because its length varies seasonally...... 44 2.4. Examples of School Springs refuge habitats: pool (A), riffle (B), run (C), and wash (D)...... 45 2.5. Comparison of the same digital fish images in original JPEG format (A) and images converted in Image Tool 11 to TIFF format (B). Lengths were determined from TIFF formatted images only...... 46 2.6. Total number of C. n. pectoralis individuals captured each survey in School Springs refuge between March 2009 and March 2010...... 47 2.7. Length-frequency histogram of C. n. pectoralis captured in School Springs refuge from March 2009 to March 2010. Total lengths displayed on the x-axis are median values for each bin...... 48 2.8. Catch curves depicting estimated total annual survival rate (Ŝ) and estimated instantaneous total mortality rate (Ž) based on C. n. pectoralis length frequency captured between March 2009 and March 2010. Surveys were grouped so that Season 1: December 21, 2009 – March 19, 2010, Season 2: March 20, 2009 – June 20, 2009, Season 3: June 21, 2009 – September 22, 2009, Season 4: September 23, 2009 – December 20, 2009. Total lengths (mm) are binned identically to Figure 2.6...... 49 2.9. Capture frequency of C. n. pectoralis by habitat type in School Springs refuge during the course of my study. Reach 19 was sampled more frequently than other reaches...... 50 3.1. Warm Springs Complex and School Springs refuge study area within Ash Meadows National Wildlife Refuge, Nevada...... 89

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3.2. School Springs refuge, Ash Meadows, Nevada, 1969-2010. A. 1969- The spring terminates in a dug-out pool. B. 1983-Four concrete pools were constructed to improve habitat for C. n. pectoralis at School Springs. C. 2008-School Springs was rehabilitated, which included diversifying the habitats for Pyrgulopsis pisteri, Stenelmis calida calida, Ambrysus relictus, and C. n. pectoralis. D. 2010- Algae was abundant in the largest pool, reach 19. Photograph credit: U.S. Fish and Wildlife Service...... 90 3.3. School Springs refuge habitat as-built depicting reach segments, as well as pool, run, and riffle habitat types. Note: The wash is not included on this map because its length varies seasonally...... 91 3.4. Examples of School Springs refuge habitats: pool (A), riffle (B), run (C), and wash (D)...... 92 3.5. Ambrysus relictus length-frequency histogram of all individuals captured during dip net sampling in School Springs refuge from February 2009 to April 2010...... 93 3.6. Locations in School Springs refuge where Pyrgulopsis pisteri was collected...... 94 3.7. Locations in School Springs refuge where Stenelmis calida calida was collected...... 95 3.8. Locations in School Springs refuge where Ambrysus relictus was collected...... 96 3.9. Scatterplot analyses depicting Pyrgulopsis pisteri abundance with chemical and physical stream properties in School Springs refuge...... 97 3.10. Scatterplot analyses depicting Stenelmis calida calida abundance with chemical and physical stream properties in School Springs refuge...... 98 3.11. Scatterplot analyses depicting A. relictus abundance with chemical and physical stream properties in School Springs refuge...... 99

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CHAPTER I INTRODUCTION There are an estimated 1.8 million described living species worldwide (Chapman

2009), although there may be as many as 30 million species (May 1992). Despite this considerable biodiversity, numerous species throughout the world are threatened or endangered with extinction by natural and anthropogenic factors. Among the latter factors, habitat loss and invasive species present the greatest threats to biodiversity

(Wilcove et al. 1998). In the past, there have been five mass extinctions due to large stochastic events such as asteroid impacts and climate change (Raup 1986). We now are in the midst of a sixth mass extinction, which is largely due to anthropogenic causes (Pimm et al. 1995; Jones 2009).

Biogeography is the field of science concerned with the distribution of living species. The distribution of a species across a landscape may affect its risk of extinction (Dobson et al. 1997; Taylor and Warren 2001; Thomas et al. 2004), so wildlife managers responsible for conserving imperiled species need to understand how the survival of populations is influenced by their isolation. Within biogeography, the theory of island biogeography attempts to explain how island size and isolation affect colonization and extinction rates (MacArthur and Wilson 1967).

The size of islands (or habitat patches) and distance between islands necessary to prevent species extinctions (Simberloff 1976; Simberloff and Abele 1976) has important implications for land managers concerned with preserving species with limited or fragmented distributions. 1

Texas Tech University, Darrick S. Weissenfluh, December 2010

In the 1970s a debate emerged among ecologists, based on island biogeography, as to whether a single large preserve or several small preserves (SLOSS) were more important for conservation. The result of the SLOSS debate was the formulation and testing of a number of hypotheses to determine whether patterns of colonization and extinction were discernable at the landscape level. The single large hypothesis is supported by the species-area relationship (Watson 1859; Arrhenius 1921; Preston

1962), which shows that species richness generally increases with area. In contrast, establishing multiple populations is more likely to prevent extinction, so spreading of risk supports several small-reserves (den Boer 1968). Both arguments are supported by numerous studies and both have led to suggestions for the design of natural reserves and refuges to reduce a species risk of extinction (Diamond 1975; Simberloff and Abele 1982).

A refuge can serve many purposes at many scales, but the primary purpose of a refuge is to reduce a species’ risk of extinction. The importance of refuges for conserving species in fragmented habitats, such as islands, has been documented in multiple studies (e.g. Pister 1974; Williams 1991; Thomas and Jones 1993; Karim and Main 2009). However, the design and placement of refuges for maximum effectiveness is less well studied. Recommendations for the design of natural reserves and refuges led to a paradigm shift in which habitat fragmentation and species survival have come to the forefront of conservation biology (Wilcox and

Murphy 1985).

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Williams (1991) defined a refuge as a place managed for one or several species, rather than for an entire biota, and which may comprise natural or artificial habitats. I subscribe to Williams (1991) definition of a refuge; however, I distinguish natural refuges from artificial refuges. For example, the first National Wildlife Refuge

(NWR), Pelican Island, was created in Florida in 1903 when land was set aside to prevent the extinction of brown pelicans Pelacanus occidentalis and other native birds by reducing the impacts of commercial hunting. Pelican Island NWR is an example of a natural refuge. In contrast, a concrete tank was constructed at the

Hoover Dam refugium in Arizona, in which a population of Devils Hole pupfish

Cyprinodon diabolis was harbored to reduce its risk of extinction (Sharpe et al. 1973).

This is an example of an artificial refuge.

Artificial refuges have been created to conserve a variety of aquatic species. For example, artificial reefs have been constructed since the 1970s as refuges for mussels and other marine invertebrates in the Mediterranean Sea (Bombace 1989). In freshwaters, artificial pools were created to aid in preserving the threatened Railroad

Valley springfish Crenichthys nevadae in the outflow of Chimney Hot Springs,

Nevada (Williams and Williams 1989) and artificial riffles were created in a Kansas river to conserve the threatened Neosho madtom Noturus placidus (Fuselier and Edds

1995). Although the construction of artificial refuges in the United States is gaining prevalence as a conservation tool, few case studies provide practical design recommendations.

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Moyle and Sato (1991) discussed the design of natural fish preserves and emphasized the importance of proper habitat, size, and stability, among other criteria; however, only two studies have suggested design considerations for artificial fish refuges in the southwest United States (e.g., Williams 1991; Winemiller and

Anderson 1997). There is no case study suggesting design considerations for artificial refuges for the conservation of freshwater aquatic invertebrates. Factors that may influence artificial refuge design include habitat requirements, population objectives, and genetic diversity. In a study of 24 semi-natural refuge populations of desert pupfish Cyprinodon macularis and six refuge populations of Quitobaquito pupfish Cyprinodon eremus, Koike et al. (2008) determined refuges could sustain genetic diversity equivalent to that occurring in wild populations of these species.

These results are encouraging, but studies evaluating the short- and long-term success of artificial refuges are needed to ensure desirable conservation objectives are met.

Many of the artificial refuges in the southwest United States were created to preserve a single species and they have had varied success (Minckley 1995; Pister

1990; Williams 1991; Wilcox and Martin 2006). In southern Nevada, managers of imperiled fishes have created numerous artificial refuges to prevent extinctions and sustain populations; currently there are at least 19 artificial refuges containing federally-listed fishes (Hobbs et al. 2007). Recovery plans, including, but not limited to, the Ash Meadows recovery plan (USFWS 1990), clearly identify artificial refuges as important for the recovery of several federally-listed pupfishes Cyprinodon.

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Indeed, refuges already have played a key role in preventing the extinction of a number of Cyprinodon species (Pister 1990).

Frequently, refuge populations are established by translocations of native species, which offer unique research opportunities when monitoring is performed (Minckley and Brooks 1980; Minckley 1995; Peacock et al. 2010). Success of a translocation and establishment of a refuge population is hard to define; however, Minckley (1995) suggests short-term success may be achieved when multiple generations of a species survives at one or more sites, although he cautions that reporting success prematurely may negatively impact a translocation program. With these considerations, and others, in mind, I evaluate the success of one endangered fish and three endemic aquatic invertebrates in Ash Meadows National Wildlife Refuge, Nevada, after an existing artificial refuge, School Springs, was renovated.

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LITERATURE CITED

Arrhenius, O. 1921. Species and area. Journal of Ecology 9:95-99. Bombace, G. 1989. Artificial reefs in the Mediterranean Sea. Bulletin of Marine Science 442:1023-1032. Chapman, A. D. 2009. Numbers of living species in Australia and the world. Report for the Australian Biological Resources Study, Canberra, Australia. 2nd edition. den Boer, P. J. 1968. Spreading of risk and stabilization of animal numbers. Acta Biotheory 18:165-194. Diamond, J. M. 1975. Assembly of species communities. Pages 342-444 in M. L. Cody and J. M. Diamond, editors. Ecology and Evolution of Communities. Belknap Press, Cambridge, MA. Dobson, A. P., A. D. Bradshaw, and A. J. M. Baker. 1997. Hopes for the future: restoration ecology and conservation biology. Science 25:515-522. Fuselier, L., and D. Edds. 1995. An artificial riffle as restored habitat for the threatened Neosho madtom. North American Journal of Fisheries Management 15:499-503.

Hobbs, B., J. Heinrich, J. Hutchings, M. Burrell, and J. C. Sjöberg. 2007. Native aquatic species program 2006 annual report: southern region. Nevada Department of Wildlife, unpublished report, Las Vegas. Jones, A. R. 2009. The next mass extinction: human evolution or human eradication. Journal of Cosmology 2:316-333. Karim, A., and M. B. Main. 2009. Habitat Fragmentation and conservation strategies for a rare forest habitat in the Florida Keys archipelago. Urban Ecosystems 12:359-370. Koike, H., A. A. Echelon, D. Lofts, and R. A. Van Den Busch. 2008. Microsatellite DNA analysis of success in conserving genetic diversity after 33 years of refuge management for the desert pupfish complex. Animal Conservation 11:321-329. MacArthur, R. H., and E. O. Wilson. 1967. The Theory of Island Biogeography. Princeton University Press, Princeton, NJ. May, R. M. 1992. How many species inhabit the earth? Scientific American (October):18-24. Minckley, W. L. 1995. Translocation as a tool for conserving imperiled fishes: experiences in western United States. Biological Conservation 72:297-309. 6

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Minckley, W. L. and J. E. Brooks. 1980. Transplantations of native Arizona fishes: records through 1980. Journal of Arizona-Nevada Academy of Science (1985) 20:73-89. Moyle, P. B., and G. M. Sato. 1991. On the design of preserves to protect native fishes. Pages 155-169 in W. L. Minckley and J. E. Deacon, editors. Battle against extinction: native fish management in the American West. University of Arizona Press, Tucson. Peacock, M. M., M. L. Robinson, T. Walters, H. A. Mathewson, R. Perkins. 2010. The evolutionary significant unit concept and the role of translocated populations in preserving the genetic legacy of Lahontan cutthroat trout. Transactions of the American Fisheries Society 139:382-395. Pimm, S. L., G. J. Russell, J. L. Gentleman, and T. M. Brooks. 1995. The future of biodiversity. Science 269:347-350. Pister, E. P. 1974. Desert fishes and their habitats. Transactions of the American Fisheries Society 103:531-540. Pister, E. P. 1990. Desert fishes: an interdisciplinary approach to endangered species conservation in North America. Journal of Fish Biology 37(Supplement A):183- 187. Preston, F. W. 1962. The canonical distribution of commonness and rarity of species. Ecology 43:185-215. Raup, D. M. 1986. Biological extinction in earth history. Science 231:1528-1533. Sharpe, F. P., H. R. Guenther, and J. E. Deacon. 1973. Endangered desert pupfish at Hoover Dam. Reclamation Era 59:24-29.

Simberloff, D. S. 1976. Experimental zoogeography of islands: effects of island size. Ecology 57:629-648. Simberloff, D. S., and L. G. Abele. 1976. Island biogeography theory and conservation practice. Science 191:285-286. Simberloff, D. S., and L. G. Abele. 1982. Refuge design and island biogeographic theory: effects of fragmentation. The American Naturalist 120:41-50. Taylor, C. M., and M. L. Warren. 2001. Dynamics in species composition of stream fish assemblages: environmental variability and nested subsets. Ecology 82:2320- 2330.

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Thomas, C. D., and T. M. Jones. 1993. Partial recovery of a skipper butterfly (Hesperia comma) from population refuges: lessons for conservation in a fragmented landscape. Journal of Animal Ecology 62:472-481. Thomas, C. D, C. Alison, R. E. Green, M. Bakkenes, L. J. Teaumont, Y. C. Collingham, B. F. N. Erasmus, M. R. de Siqueira, A. Grainger, L. Hannah, L. Hughes, B. Huntley, A. S. van Jaarsveld, G. F. Midgley, L. Miles, M. A. Ortega- Huerta, A. T. Peterson, O. L. Phillips, and S. E. Williams. 2004. Extinction risk from climate change. Nature 427:145-148. Watson, H. C. 1859. Cybele Britannica. Vol. 4. London. 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. Wilcox, B. A., and D. D. Murphy. 1985. Conservation strategy: the effects of fragmentation on extinction. The American Naturalist 125:879-887. Wilcox, J. L., and A. P. Martin. 2006. The devil’s in the details: genetic and phenotypic divergence between artificial and native populations of the endangered pupfish (Cyprinodon diabolis). Animal Conservation 9:316-321. Williams, C. D., and J. E. Williams. 1989. Refuge management for the threatened railroad valley springfish in Nevada. North American Journal of Fisheries Management 9:465-470. Williams, J. E. 1991. Preserves and refuges for native western fishes: history and management. Pages 171-189 in W. L Minckley and J. E. Deacon, editors. Battle against extinction: native fish management in the American West. The University of Arizona Press, Tucson. Winemiller, K. O., and A. A. Anderson. 1997. Response of endangered desert fish populations to a constructed refuge. Restoration Ecology 5:204-213. USFWS (U.S. Fish and Wildlife Service). 1990. Recovery plan for the endangered and threatened species of Ash Meadows, Nevada. Portland, Oregon.

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CHAPTER II

HABITAT ASSOCIATION AND DISTRIBUTION OF ENDANGERED WARM SPRINGS PUPFISH, CYPRINODON NEVADENSIS PECTORALIS, IN SCHOOL SPRINGS REFUGE INTRODUCTION

In the greater Death Valley system, which includes Death Valley National Park in

California and Ash Meadows National Wildlife Refuge (Ash Meadows) in Nevada, there

are 12 described pupfish species (genus Cyprinodon) and subspecies. These Cyprinodon species inhabit a variety of habitats in the greater Death Valley system, including spring

and wetland habitats (Miller 1948; Soltz and Naiman 1978; Soltz and Hirshfield 1981).

One subspecies, Tecopa pupfish, C. nevandensis calidae, is now believed extinct (Soltz

and Naiman 1978; Miller et al. 1989).

Individual populations of Cyprinodon also have been extirpated in the greater Death

Valley system (Pister 1974) and continue to be at risk due to introductions of non-native

species, spring and groundwater pumping, and climate change, among other factors.

Non-native species, such as western mosquitofish, Gambusia affinis, and red swamp

crayfish, Procambarus clarkii, have been recorded in desert springs of the greater Death

Valley system, since at least the 1930’s (Miller 1948) and may have led to the extinction

of the Ash Meadows killifish, Empetrichthys merriami (Soltz and Naiman 1978), which

inhabited the same springs as the Ash Meadows Amargosa pupfish C. n. mionectes.

Similarly, spring-source pumping, has led to the extirpation of at least one Cyprinodon

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Texas Tech University, Darrick S. Weissenfluh, December 2010 population (Pister 1974), and groundwater pumping is a growing threat as humans

consume groundwater in quantities exceeding recharge rates in areas such as Las Vegas,

NV, (Deacon et al. 2007, Bushman et al. 2010). Climate change also is expected to

adversely impact Cyprinodon species (Martin 2010), given reduced precipitation

forecasts for Death Valley region (Milly et al. 2005; Seager et al. 2007). The creation of

artificial refuges is one way managers are coping with known and unknown threats facing

desert fishes in the race to preserve their existence.

Artificial refuges are important to imperiled fish recovery and are identified as a

recovery strategy in recovery plans such as the Ash Meadows recovery plan (USFWS

1990). However, refuges created for imperiled fishes in the western United States have

had mixed success (Williams 1991). This is especially true for Cyprinodon species in the

greater Death Valley system.

In particular, refuges created for the conservation of the Devils Hole pupfish, C.

diabolis, have had little success. The first refuge site, Purgatory Spring in Ash Meadows,

was a well-seep used for a translocated population of C. diabolis; however, the

population was eradicated because fish exhibited a significant change in morphology

compared to individuals remaining in Devils Hole, its only native habitat (Liu and Soltz

1983 cited in Williams 1991). A second artificial refuge, the Hoover Dam refuge, was

created for preservation of C. diabolis at the Willow Beach Fish Hatchery (Williams

1991). Williams (1977) concluded the Hoover Dam refuge also was unsuccessful

because C. diabolis exhibited altered morphology when compared with fish in Devils

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Hole. Within Ash Meadows, two artificial refuges, Point-of-Rocks refuge and the

Amargosa pupfish research station, were constructed to sustain a refuge population of C.

diabolis by mimicking the ecology of Devils Hole (Sharpe et al. 1983); however, these

refuges failed to perpetuate C. diabolis for unknown reasons. As the above examples

illustrate, monitoring of artificial refuge population is very important; however,

determining why they fail can be difficult.

Williams (1991) and Minckley (1995) both emphasized the importance that

monitoring refuges can play in successfully conserving species and recommended the

publication of monitoring efforts. Although artificial and semi-natural refuges

increasingly are common (Williams 1991), most studies of artificial fish refuges have

focused on genetics and morphology of the inhabitants (Lema and Nevitt 2006; Wilcox

and Martin 2006; Martin 2010). It is intuitive that an artificial refuge would be constructed to meet the habitat needs of the species it was created for by mimicking its natural habitat, yet few studies have evaluated this notion.

The U.S. Fish and Wildlife Service renovated School Springs refuge in June 2008 to

improve habitat for C. n. pectoralis, establish populations of endemic aquatic

invertebrates including the median-gland Nevada springsnail, Pyrgulopsis pisteri, Devils

Hole warm springs riffle beetle, Stenelmis calida calida, and warm springs naucorid,

Ambrysus relictus, and eradicate aquatic non-native species. The purpose of this study is to evaluate the design of School Springs refuge because it was designed to mimic habitats

in which C. n. pectoralis occur naturally and determine the eradication success of two

11

Texas Tech University, Darrick S. Weissenfluh, December 2010 non-native aquatic species. In regards to the former, I tested one hypothesis implicit in

the refuge design:

Ho: C. n. pectoralis, regardless of life stage, use no specific habitat in School Springs

refuge.

STUDY AREA

Ash Meadows is a rare desert oasis (Fraser and Martinez 2002) and has more than 50

perennial seeps and springs. It is located approximately 128 kilometers northwest of Las

Vegas, Nevada. The Ash Meadows National Wildlife Refuge was established in 1984 to

conserve threatened and endangered species. This refuge contains the largest

concentration of endemic species (N = 26) in the United States (Stevens and Bailowitz

2008; Crews and Stevens 2009).

The Warm Springs Complex (WSC) is one of four management units within Ash

Meadows (Figure 2.1). The only fish native to this complex is the Warm Springs

pupfish, C. n. pectoralis, which is federally listed as endangered. The WSC consists of

six low-discharge warm springs with individual flows ranging from 1.13 x 10-4 to 1.98 x

10-4 cm-s and spring-source water temperatures ranging from 28o to 34oC. A seventh

spring, Mexican Spring, dried in 1973 due to evapotranspiration, resulting in the

extirpation of a C. n. pectoralis population (Soltz 1974; Kodric-Brown and Brown 2007).

School Springs has the lowest flow and highest source water temperature among

WSC springs. Miller (1948) first collected and described C. n. pectoralis in 1939 at

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Texas Tech University, Darrick S. Weissenfluh, December 2010

School Springs, then referred to as Lovell Spring, and described the habitat as a concrete pool approximately 6.1 m wide, 7.6 m long, and 2.4 m deep with silt on the bottom

occurring about 50 yards below the spring source. This was the only location at which C. n. pectoralis was known to occur at the time. In 1966, two additional populations were

discovered at Scruggs Springs and Indian Springs, where the habitat was described as a

ditch with silt, Chara sp., and small stone substrate (Miller and Deacon 1973). Besides

physical alteration of springs and their hydrology, Miller and Deacon (1973) observed

western mosquitofish, Gambusia affinis, a non-native species, present at Scruggs Springs

and Indian Springs in 1966. Based on these reports, human alteration of C. n. pectoralis

habitats was well underway by the 1930s.

In 1969, the Bureau of Land Management (BLM), which managed School Springs at

that time, created an earthen pond at School Springs (Figure 2.2). The pond was created

to increase water volume and available habitat for the soon-to-be (1970) federally listed

C. n. pectoralis. Three springs discharged water into the pond, which was approximately

2.5 m x 1 m in area and 0.25 m in depth.

The BLM further increased water volume and available habitat for C. n. pectoralis at

School Springs in 1983 by constructing four concrete ponds downstream from the spring

sources and approximately 20 m of earthen outflow. The outflow soon became densely

vegetated with American rush, Schoenoplectus americanus. All four ponds constructed

in 1983 had a centrally located standpipe, which drained water to an observation pond

downstream.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Sometime after 1983, three aquatic non-native species, G. affinis, P. clarkii, and red-

rimmed melania, Melanoides tuberculatus, invaded School Springs refuge. It is unclear

exactly how and when they arrived at School Springs or whether they adversely impacted

C. n. pectoralis, because there was no systematic monitoring of the refuge. Nevertheless,

these non-native species likely competed for limited resources with C. n. pectoralis,

given the limited habitat and food resources.

From April-May 2008, nearly the entire system, except for the two spring-sources,

was temporarily desiccated to renovate School Springs refuge. The desiccation also was an opportunity to eradicate G. affinis, P. clarkii, and M. tuberculatus. I only report on G. affinis and P. clarkii and in this chapter, as they were susceptible to capture in my traps.

Refer to Chapter 3 for the status of M. tuberculatus.

School Springs refuge was extensively renovated in June 2008 to eradicate aquatic

non-native species, improve habitat for C. n. pectoralis, and establish populations of

endemic aquatic invertebrates (see Chapter 3). Prior to the renovation, an extensive

effort was made to salvage all C. n. pectoralis from School Springs. In total, 634 adult

and juvenile C. n. pectoralis were salvaged in April 2008 and held in captivity (e.g.

aquaria) until June 2008 while the refuge was renovated. Two weeks after renovation

was completed, algae were present throughout the stream. C. n. pectoralis were then

reintroduced into the refuge. In total, 761 C. n. pectoralis adults and juveniles were

translocated. This number was 127 greater than the number salvaged as a result of successful reproduction during the School Springs refuge renovation.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

In 1969, School Springs consisted of three springs; however, only two were present prior to the renovation in 2008. It is unclear what happened to the third spring. Only one of the two remaining springs, the northern spring, continued to flow after renovation in

2008. Discharge of this spring was 4.42 x 10-3 cm/s prior to renovation and monitoring

has indicated it has remained steady. In this spring, water temperature has also remained

stable at 34oC, which is similar to that reported by Sumner and Sargent (1940).

Discharge of the second spring was only 1.26 x 10-3 cm/s prior to renovation and has

since ceased flowing. This may have resulted from M. tuberculatus eradication attempts,

which included excavation of soil around the spring source.

The renovated School Springs refuge channel is approximately 101 meters long.

Within the renovated channel, the average width is 0.69 m and the average depth is 0.45

m (Appendix A). Much of the renovated refuge consists of an artificial base-substrate

constructed of concrete and mortar, except a middle portion of the stream channel, which

was constructed without a concrete base, having only the sides mortared. Below the

renovated channel is an ephemeral wash, which was unaltered and consists of natural

sand, silt, and rock. Its length varies seasonally, flowing above ground more than 100

meters during the winter and less than 40 meters during the summer, because of water

loss due to evapotranspiration. Concrete and mortar were used to limit the growth of

emergent vegetation in the renovated channel, especially cattails, Typha domingensis, and

S. americanus, which alter hydrology and promote exotic species populations

(Scoppettone et al. 2005). Because C. n. pectoralis is thermophilic, an additional reason

15

Texas Tech University, Darrick S. Weissenfluh, December 2010 for limiting the growth of vegetation was to maintain warm waters as far downstream as possible.

It is unclear which habitats and substrates were present historically in School Springs refuge; however, because of its elevation and low flow, the system was probably characterized by dense emergent vegetation with little variability in habitat or substrate.

Cyprinodon species utilize a variety of habitats, but are typically in greatest abundance in

spring-source and marsh habitats (Soltz and Naiman 1978). Although the renovated

channel consists of concrete and mortar, habitats within the stream were constructed to

include numerous pools, runs, and riffles of various sizes and depths to meet the

perceived habitat requirements of C. n. pectoralis and endemic aquatic invertebrates.

Substrate such as sand, silt, and rocks of various sizes also were added to the stream

channel based on hypothesized habitat preferences of endemic aquatic invertebrates (see

chapter 3). Prior to renovation in 2008, pool habitat was confined to the concrete pools at the terminal end of the refuge, run habitat only occurred in the upper 20 meters of the refuge, and riffle habitat was absent. Similarly, silt substrate was the only substrate present.

METHODS

STUDY DESIGN

I conducted 28 surveys between March 2009 and March 2010. Prior to sampling, the

School Springs stream channel was divided into 32 reaches, each distinguishable by

sequential position (Figure 2.3). Among the 32 reaches there are four distinguishable 16

Texas Tech University, Darrick S. Weissenfluh, December 2010 habitat strata within School Springs refuge including pools, runs, riffles, and a wash

(Figure 2.4). Every two weeks, I randomly selected eight sample sites within each

habitat stratum using Hawth’s Tools (Beyer 2003) within ArcGIS 9.2 (ESRI, Inc. 2006),

resulting in 32 samples per survey and 896 different random sampling locations throughout the duration of my study. Sample sites were surveyed from the lower reach of the channel upstream towards the spring source to reduce the probability of capturing the same fish or moving invertebrates throughout the refuge.

Because the largest pool, reach 19, is much larger than any of the other pools, three of the eight randomly selected pool samples each survey always occurred in that pool. Also,

because the spring source, reach 1, was thought to be important habitat for C. n.

pectoralis prior to renovation efforts, one pool sample each survey always occurred in the

spring pool. The wash, reach 32, varied in length seasonally, so if randomly selected

sites turned out to be dry on the day of sampling, alternative sites with enough water to

sample were randomly chosen. The spring source, reach 3, and the channel dependent on

its flow, reach 4, were not sampled during the course of my study, because the spring stopped flowing after renovation in 2008.

On each survey, one live trap capable of capturing larval, juvenile, and adult C. n.

pectoralis was placed in each of the 32 sample sites. Due to the variation in water depth

(0.5 to 47 cm) in the School Springs refuge, a standard trap, capable of capturing fish in

all water depths, was used. Live traps were constructed from window-screen mesh (1

mm2) wrapped around an oval steel wire frame (approximately 7 cm x 7 cm x 27.9 cm).

17

Texas Tech University, Darrick S. Weissenfluh, December 2010

Funnel entrances, approximately 1.9 cm in diameter, were located at either end. Each

trap was baited with dry cat food, which is known to be effective bait for C. n. pectoralis.

Traps were set for 30 minutes at 5- to 10-minute intervals beginning with the trap farthest

downstream. After I removed traps from the channel, captured fish were placed into

plastic buckets that contained a solution of water and Stress Coat® to minimize handling

stress. Captured C. n. pectoralis then were placed in white trays (25.4 cm x 20.3 cm x

4.4 cm) at a maximum density of eight fish per tray, where they were photographed to

later determine their length and age. Approximately 10% of those fish also were

physically measured when it was convenient to do so, without risking stress and handling

mortalities. Fish were immediately released to the location where they were captured

after they were photographed and measured.

DETERMINATION OF LIFE STAGES USING DIGITAL IMAGES

Three digital images were taken of each group of fish placed in the white tray using a

Fujifilm F650 6.0 megapixel digital camera. The camera was connected to a tripod and

positioned approximately 0.30 m above the center of the tray. A ruler was placed on the

bottom of the tray as a reference scale.

Digital images were used to estimate lengths and classify fish into age classes using

the software ImageTool 3.0 (UTHSCSA, 2002). Mettee and Beckham (1978)

distinguished larval, juvenile, and adult sheepshead minnow, C. variegatus, based on

length. I used their length categories for C. n. pectoralis: larval fish < 14.4 mm total

length (TL), juvenile fish ≥ 14.4-mm TL but ≤ 20.1-mm TL, and adult fish > 20.1-mm

18

Texas Tech University, Darrick S. Weissenfluh, December 2010

TL. I used ImageTool 3.0 to convert digital images from JPEG to TIFF format prior to

making length measurements. As a side effect of ImageTool 3.0, fish in TIFF format

images were tinted blue, which made fin margins, especially caudal fins, more

conspicuous. An example of a raw image (JPEG) and a converted image (TIFF) are

shown in Figure 2.5.

Approximately 10% of fish during each survey were physically measured. At the

onset of sampling, I used a paired t-test analysis to determine if hand measurements were

more accurate than digital image measurements. Based on my analysis, digital image

measurements did not differ significantly from hand measurements (t = 0.487, df = 9, P >

0.05). Therefore, I used only length measurements obtained from digital images in my

analyses.

I used length data to calculate adult-juvenile ratios, create length-frequency

histograms, and to calculate densities, estimate population size, and estimate annual

mortality of fish by age class in School Springs refuge. The ratio of adults to juveniles

was calculated for each survey by dividing the number of adults captured by the number

of juveniles captured in all 32 samples. Length-frequency histograms were transformed

using log(n) and were used to determine if there was evidence of recruitment. I

calculated the density of adult C. n. pectoralis m-2 survey each using the total number of

adults captured, with the assumption that each trap fished an area of 1 m2, which was

based on observations made in School Springs refuge. Densities were then multiplied by

the estimated surface area of water, 86.42 m2, in the renovated School Springs refuge to

19

Texas Tech University, Darrick S. Weissenfluh, December 2010 estimate the adult population. Annual mortality rates (Ŝ) and instantaneous mortality

rates (Ž) for adult C. n. pectoralis were estimated from length-frequency histograms following the Chapman-Robson method (Chapman and Robson 1960; Robson and

Chapman 1961).

I used Shapiro-Wilks normality test to determine if length measurements and abundance were normally distributed. Neither C. n. pectoralis length nor abundance

data were normally distributed. Therefore, I used a non-parametric Kruskal-Wallis

analysis of variance to determine if average fish length and abundance varied

significantly between surveys and seasons.

C. n. pectoralis HABITAT ASSOCIATION

To determine if C. n. pectoralis were associated with specific habitat types, I assumed

as a null model that 25% of fish would use each habitat (pool, run, riffle, and wash).

Pearson’s chi-square was calculated to determine which habitats C. n. pectoralis was

associated with more often than expected. I used STATISTIX 9 (Analyitical Software

2008) for these statistical tests and significance was determined at P ≤ 0.05.

In R (R Development Core Team 2009), a post hoc analysis using the generalized

linear model routine was completed to determine if area or water volume alone explained

why C. n. pectoralis presence was associated with pool habitat, more than any other

habitat type. The null hypotheses associated with these models are:

H01: C. n. pectoralis distribution is independent of habitat type;

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Texas Tech University, Darrick S. Weissenfluh, December 2010

2 H02: C. n. pectoralis distribution is independent of habitat area (m );

3 H03: C. n. pectoralis distribution is independent of water volume (m );

H04: C. n. pectoralis distribution is independent of habitat type, habitat area, and

water volume (m3).

I used the AICc (Akaike Information Criterion) statistic (Hurvich and Tsai 1989) to determine which model best predicted C. n. pectoralis presence in School Springs refuge.

The AICc is a better statistic when the number of samples divided by the number of parameters is fewer than 40 (Burnham and Anderson 1998), which was the case for my models. Therefore, I always chose the model with the lowest AICc as the best model.

PHYSICAL AND CHEMICAL STREAM VARIABLES IN SCHOOL SPRINGS

REFUGE

Stream slope, substrate, habitat, algae density, vegetation density, velocity, dissolved oxygen (DO), pH, conductivity, water temperature, and water depth were measured at each sample site. Minimum, maximum, mean, standard deviation, and sample size for each variable in each reach are presented in Appendix B. Pearson correlation coefficients were calculated to measure the strength of association between physical and chemical variables. Of all variables measured, only water temperature and pH were not associated with each other (P > 0.05); however, there was an association between their values and the distance from the spring source (P < 0.01). Because School Springs refuge issues from a single point source with a stable discharge water temperature, it was expected that

21

Texas Tech University, Darrick S. Weissenfluh, December 2010 these variables would be associated with the distance from spring source. At least two

studies (Soltz and Naiman 1978; Minckley et al. 1991) have determined Cyprinodon sp.

have a wide tolerance for physical and chemical variables and, therefore, suggest these

variables rarely are an obstacle for management of these species. For this reason and

because I observed C. n. pectoralis throughout School Springs refuge, I chose not to

evaluate the influence of physical and chemical variables on its distribution.

RESULTS

NON-NATIVE AQUATIC FISH AND CRAYFISH ERADICATION RESULTS

The temporary desiccation of School Springs refuge to eradicate G. affinis and P.

clarkii had mixed success. One G. affinis was collected with a dip net in July 2008,

indicating desiccation was not completely successful. However, no additional G. affinis

were collected during the course of my study, which suggests the specimen removed by

dip net may have been the last individual in School Springs refuge. Therefore, eradication of this species appears to be successful. P. clarkii was not observed or

collected after renovation or during the course of my study, indicating desiccation led to

its eradication.

STATUS OF C. n. pectoralis IN SCHOOL SPRINGS REFUGE

C. n. pectoralis was captured during every survey, but the abundance, density, and

adult-juvenile ratios varied throughout my study. On average, I captured 185 individuals

each survey, with the lowest (N = 82) number of individuals captured on 3 Feb 2010 and

22

Texas Tech University, Darrick S. Weissenfluh, December 2010 the greatest (N = 259) number captured on 18 March 2009 (Figure 2.6). Adult C. n.

pectoralis density averaged 4.9 m-2 throughout the duration of my study, but for 3 Feb

2010 and 18 March 2009 was 2.38 m-2 and 6.47 m-2, respectively. The estimated adult

population averaged 424 fish, but varied between 205 on 2 March 2010 and 570 on 21

July 2009. The adult-to-juvenile ratio averaged 7.32 but was lowest, 3.03, on 15

September 2009 and greatest, 14.50, on 27 May 2009 (Table 2.1).

During the course of my study, 4,228 adult, 709 juvenile, and 54 larval fish were

captured. C. n. pectoralis length varied significantly between surveys (U = 202.35; df =

26; P < 0.001) and seasons (U = 49.78; df = 3; P < 0.001). The majority of fish captured

were between 21- to 26-mm TL; however, the smallest fish captured was 9-mm TL and

the largest fish captured was 54-mm TL. Smaller fish were progressively less common below 21-mm TL, as were larger fish above 26-mm TL (Figure 2.7). Annual survival

rates (Ŝ) for adult C. n. pectoralis varied seasonally between 61% and 66%, whereas the

estimated instantaneous mortality rate (Ž) varied seasonally between 16% and 22%

(Figure 2.8). To my knowledge, this is the first time annual survival and instantaneous

mortality rates have been estimated for Cyprinodon species in the greater Death Valley

system.

I captured C. n. pectoralis in all four habitat types every survey. Of the 4,998 C. n.

pectoralis captured during sampling, 3,875 were captured in pools, 343 were captured in riffles, 624 were captured in runs, and 156 were captured in the wash (Table 2.2). The

greatest number of fish was captured in the largest pool, reach 19 (Figure 2.9). Adult fish

23

Texas Tech University, Darrick S. Weissenfluh, December 2010 were captured in all reaches, although 47.9% were captured in reach 19, the largest pool.

Juvenile C. n. pectoralis were captured in every reach except 16, 20, and 27. The

greatest percentage of juvenile fish (38.5%) was captured in reach 19. Larval C. n.

pectoralis were captured in 57% of reaches and in all habitat types (Table 2.3). The

greatest percentages of larval fish were captured in reaches 11 (37%) and 19 (25.9%),

both of which are pools.

C. n. pectoralis HABITAT ASSOCIATION IN SCHOOL SPRINGS REFUGE

During the course of my study, C. n. pectoralis presence was not independent of

habitat type, regardless of season (χ2 = 1531.12, df = 3, P < 0.01). Similarly, the presence

of larval (χ2 = 82.00, df = 3, P < 0.01), juvenile (χ2 = 882.03, df = 3, P < 0.01), and adult

(χ2 = 6684.53, df = 3, P < 0.01) fish was not independent of habitat type. Habitat type,

area, and water volume combined (Model 4; AICc = 1721.3);) were the best predictor of

presence overall; however, habitat type (Model 1; AICc = 1959.0) was a better predictor

of C. n. pectoralis presence than were area (Model 2; AICc = 2168.6) and water volume

(Model 3; AICc = 2192.7).

DISCUSSION

SCHOOL SPRINGS REFUGE RENOVATION: WAS IT A SUCCESS?

The ultimate objective behind the creation of an artificial refuge is to sustain an

imperiled population. Success, however, will vary depending on the objectives of the

refuge, as well as the timeframe established for achieving those objectives. Intuitively,

24

Texas Tech University, Darrick S. Weissenfluh, December 2010 success of a refuge in perpetuating a population could be identified in the short-term by documenting a species presence from survey to survey, recruitment of young fish into the population, and by maintaining a minimum population size. During the course of my study, I collected individuals in every survey and observed recruitment of C. n. pectoralis into the population. The average estimated population of adults was 424 fish. As of

September 2010, C. n. pectoralis was still present and abundant in School Springs refuge,

both of which are indicators of success, to date, at School Springs refuge.

In 1976, the C. n. pectoralis population in School Springs refuge was estimated to be

58 (43-86, 95% CI) in the main pool in late June, and although there were two other pools, the number of individuals captured was too small to estimate population size (Soltz

1976). In all three pools, Soltz (1976) captured a total of 75 fish, of which only two were

reported as being juveniles. This yields an adult-juvenile ratio of 37.5. Soltz used traps

which were effective at catching fish as small as 12 mm. For comparison, 84 fish was the

fewest number of individuals I collected and I averaged more than 185 fish.

Additionally, the adult-juvenile ratio was never greater than 21.78, although it averaged

7.63 during the two surveys I conducted in June 2009. The greater abundance and lower

adult-juvenile ratio in my study suggests a larger population and more recruitment than in

Soltz’s 1976 survey, which is additional evidence of success.

In a refuge constructed for the endangered Comanche Springs pupfish, Cyprinodon

elegans, and Pecos gambusia, Gambusia nobilis, in Texas, Winemiller and Anderson

(1997) reported C. elegans juveniles were rare. They speculated that juvenile recruitment

25

Texas Tech University, Darrick S. Weissenfluh, December 2010 was low because carrying capacity was reached in the refuge. Although there was

evidence for recruitment in my study, larval and juvenile C. n. pectoralis were not

abundant. Throughout the duration of my study, I captured only 54 larval fish. From my

study it is unclear if my results are due to gear bias or to low recruitment. Soltz and

Naiman (1978) suggest larval Cyprinodon sp. experience high mortality, as eggs and fry

are preyed upon by aquatic invertebrates such as snails; however, most Cyprinodon sp.

reach maturity only two to four months after hatching. High mortality and fast growth

rates may explain why larval fish were not very abundant in my samples, yet 21- to 26-

mm TL fish were most abundant.

Soltz (1974) reported C. n. mionectes populations peaked in warm-spring systems

during the spring and autumn and were lowest in winter. Based on my density estimates,

the peak population of C. n. pectoralis in School Springs refuge occurred in July 2009,

whereas the lowest density occurred in September 2009. These results may be explained

by the timing of the initial C. n. pectoralis reintroduction, which coincided with warmer

water temperatures conducive for reproduction. This may have resulted in an unusually large number of fish surviving through their first winter in School Springs refuge. Soltz and Naiman (1978) report population size generally increases in summer with warmer

temperatures. Although I did not observe large population oscillations, I did estimate C.

n. pectoralis population size was greatest during in summer months in School Springs

refuge.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Numerous studies have shown that Cyprinodon species occupy a variety of habitats including spring-sources, rivers, ponds, and wetlands (Miller 1948; Soltz and Naiman

1978; Minckley et al. 1991). In School Springs refuge, I determined all life stages of C. n. pectoralis used pool habitat more often than any other habitat type, regardless of season. Although larval fish of some stream fishes prefer shallow riffles (Schlosser

1982), I captured larval and juvenile C. n. pectoralis in pool habitats more than in any other habitat type. The largest pool, reach 19, accounted for the greatest abundance and density of fish throughout my study and abundance was greater in pool habitats overall, regardless of season. This led me to questions whether C. n. pectoralis presence in habitat was due to water volume. Therefore, I modeled C. n. pectoralis presence with habitat type, area, and water volume separately to determine if one predicted fish presence better than another. I found that habitat area (m2) and water volume (m3) did not predict presence as well as did habitat type, which suggests C. n. pectoralis prefer pool habitat for reasons other than area and water volume. Algae and detritus are the primary food sources of many Cyprinodon species (Soltz and Naiman 1978). Algal density was positively correlated with pool habitat in School Springs refuge, so algae is one possible explanation for why C. n. pectoralis was associated with pool habitat.

Based on my generalized linear model analysis, I recommend that future monitoring in

School Springs refuge focus on sampling pools.

Spatial distribution of C. n. pectoralis in School Springs refuge did not change seasonally; however, I neither captured nor observed C. n. pectoralis in the wash in numbers comparable to those in other habitats in School Springs refuge. Scoppettone 27

Texas Tech University, Darrick S. Weissenfluh, December 2010

(2009) observed large population oscillations of C. n. pectoralis in the lowest reach of the

nearby South Scruggs Spring, in which warmer temperatures occurred during the

summer. Currently, South Scruggs Spring outflow is confined to a ditch, best described

as a run that contains no pool or riffle habitats. It also contains large numbers of exotic

P. clarkii in most of the outflow, except the lowest reach where Scoppettone (2009)

observed the greatest seasonal variation of C. n. pectoralis abundance and density. The

contrasting use of the lowest reach by C. n. pectoralis in South Scruggs Spring and

School Spring refuge may be influenced by the presence of P. clarkii in South Scruggs

Spring, differences in overall habitat availability for C. n. pectoralis between the two

systems, and water temperature fluctuations.

DOES SCHOOL SPRINGS REFUGE REDUCE THE THREATS TO C. n.

pectoralis CONSERVATION?

Threats to the conservation of imperiled fishes include non-native species,

diminishing supplies of groundwater, and habitat loss. All of those factors may lead to

the loss of individual populations, as well as entire species. One way refuges increase the

likelihood of sustaining one or more imperiled species is by reducing threats to their

survival.

Prior to renovation of the School Springs refuge, one threat to C. n. pectoralis conservation was non-native species, specifically G. affinis and P. clarkii. Both of these

species can adversely affect fishes because they are prolific and utilize many of the same

resources as native fish (Kennedy et al. 2005; Nico et al. 2010). Many eradication

28

Texas Tech University, Darrick S. Weissenfluh, December 2010 strategies have been deployed to eradicate P. clarkii and G. affinis, including chemicals, but many have been successful (Holdich et al. 1999; Peay 2001).

The renovation of School Springs refuge provided an opportunity and means to

eradicate both G. affinis and P. clarkii. Although eradication of these species often is

unsuccessful, both have been eradicated from School Springs refuge. From April to May

2008 the entire School Springs system, except the two spring-sources, was desiccated by

diverting the flow of water from the springs into 5-cm diameter PVC pipe. Although the

two springs were not desiccated, pebble and cobble rock were placed in the springs to

reduce the water volume and, therefore, the ability of P. clarkii to burrow and survive.

Both P. clarkii and G. affinis are susceptible to capture in my traps; however, during the

course of my study, my effort amounted to 14,400 trap hours and neither P. clarkii nor G.

affinis individuals were captured. These results suggest desiccation can be effective at

eradicating P. clarkii and G. affinis, which is promising for other refuges where non-

native species occur and the use of piscicide is undesirable, illegal, or impractical.

Habitat loss and groundwater pumping also are threats to the conservation of C. n. pectoralis. At Ash Meadows, and other desert oases, surface waters from springs often

flow varying distances depending on the season, due to solar radiation and evapotranspiration, resulting in the loss of habitat either temporarily or permanently.

Water loss due to evapotranspiration can have dire consequences for species occupying

these habitats. For example, evapotranspiration caused the loss of a C. n. pectoralis in

the WSC, and the Owen’s Valley pupfish, C. radiosus, was almost lost (Soltz 1974;

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Pister 1993; Kodric-Brown and Brown 2007). Because of these occurrences, as well as

climate change predictions, School Springs refuge was constructed with a mortar base,

which limits the growth of emergent vegetation and reduces the loss of surface water to

the soil. The long-term impact of mortar in this system is unclear; however, providing

enough water to sustain a viable population is critical to the conservation of C. n.

pectoralis.

DETERMINING C. n. pectoralis LENGTH FROM DIGITAL IMAGES

Monitoring plant and wildlife species is time consuming, expensive, and difficult

(Caughlan and Oakley 2001; Field et al. 2007). Because digital cameras are omnipresent

and cost effective, the ability to estimate length from digital images in reducing

processing time and has been shown to be more accurate than hand measurements (Mott

et al. 2010, this study). Despite varying environmental conditions leading to inconsistent

digital image quality, I determined that obtaining accurate total-length measurements

from digital images of Cyprinodon in the field is possible. It is likely that digital-image

measurements considerably reduce handling stress because handling is reduced and

multiple fish can be photographed simultaneously, leading to more rapid release and,

possibly, reduced mortality due to a decrease in stress factors associated with handling.

Based on my analysis, I recommend that future monitoring requiring length measurements of C. n. pectoralis and other Cyprinodon species utilize digital images for

measurements, unless specific research objectives require an alternative method. Also, it

may be possible to take digital images of the fish dorsally, which would allow the

30

Texas Tech University, Darrick S. Weissenfluh, December 2010 collection of additional morphological characteristics, such as pectoral fin length and caudal fin length, in addition to total length. This could be especially valuable if there is a reason to detect changes in fish morphology.

RECOMMENDATIONS FOR THE MANAGEMENT OF SCHOOL SPRINGS

REFUGE

Although my study has documented the short-term success of School Springs refuge,

the ultimate goal of the refuge is to sustain the C. n. pectoralis population over the long

term. Minckley (1995) recognized that failure of translocated populations can only be

determined when monitoring has documented the disappearance in a narrow span of time.

If time or funding is limited, then I recommend sampling intensity in School Springs refuge be reduced to a subset of samples on a bi-weekly basis or once monthly. The

frequency and intensity of monitoring will always depend on the availability of resources

(time, money, etc.), but should focus on (1) survival, (2) establishment, (3) population

growth, and (4) research opportunities, although it is important to consider the efficiency

of sampling a species and its life expectancy, as well (Minckley 1995). Because C. n.

pectoralis is short-lived and only inhabits five isolated springs, I recommend continued

monitoring of their population in School Springs refuge. However, the objectives of

future monitoring, as well as the definition of success, must be clearly defined before

additional monitoring.

Based on my analyses and observations, I believe the following five questions should

be the focus of future monitoring at School Springs refuge: (1) are larval C. n. pectoralis

31

Texas Tech University, Darrick S. Weissenfluh, December 2010 fully represented in my samples and, if so, what is the greatest source of mortality for larval fish? (2) do larval C. n. pectoralis use specific habitat types? (3) what is the

carrying capacity of C. n. pectoralis in the refuge? (4) how does water temperature affect

C. n. pectoralis distribution within specific habitats (e.g. pools)?, and (5) how has C. n.

pectoralis genetic structure respond to renovation of School Springs refuge? Answering

these questions will allow managers to better assess the short- and long-term success of

the School Springs refuge and may provide insight into the recovery of C. n. pectoralis in

restored natural habitats as well.

As for the definition of success, I suggest monitoring occur at least once per month at

School Springs refuge and the following criteria are used: (1) C. n. pectoralis is present

(2) a minimum adult population size is estimated (the minimum population size needs to

be determined based on genetic viability of the population), and (3) based on length-

frequency analysis, there is evidence of recruitment from March-September. If any of

these criteria are not met, immediate action should be taken to remedy the failure before

the population is lost. The information I collected and previously discussed, including

life-history, as well as physical and chemical stream properties, may be useful to assess

whether changes occur in School Springs refuge and if they are adversely impacting C. n.

pectoralis.

I did not study the genetic structure of C. n. pectoralis; however, genetic variation is

important for the conservation of fishes (Minckley 1995; Storfer 1999; Wilcox and

Martin 2006; Peacock et al. 2010), especially for artificial refuge populations (Turner

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Texas Tech University, Darrick S. Weissenfluh, December 2010

1984; Koike et al. 2008; Martin 2010). In School Springs refuge, Martin (2010) estimated the effective population size of C. n. pectoralis to be 51 (10-263, 95% CI) fish, based on samples collected prior to the renovation of School Springs refuge in 1998 and

2008. Based on his findings, Martin (2010) suggested that assisted migration, whereby

refuge staff translocates individual fish of the same species among springs, may be

necessary to prevent further loss of genetic variation (Martin 2010). My study shows that

alteration of habitats may be successful in increasing population size, which also has been

shown to maintain genetic variability (Frankman 1996). Ideally, however, one would

monitor both the population of the fish themselves, as well as their genetic structure to

ensure the population remains viable. Therefore, additional monitoring of C. n. pectoralis

in School Springs refuge is warranted to determine if the constructed habitat has

benefited the fish genetically and will continue to sustain the population.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

LITERATURE CITED

Analytical software. 2008. STATISTIX 9, user’s manual version 9. Analytical Software, Tallahassee, Florida.

Bentley, C., Leonard, R. H., Nelson, C. F., and S. A. Bentley. 1999. Journal of the American Dental Association 130:809-816.

Beyer, H. 2003. Hawth’s Analysis Tools. URL: http://www.SpatialEcology.com.

Burnham, K. P., and D. R. Anderson. 1998. Model selection and inference: a practical information-theoretic approach. Springer-Verlag, New York.

Bushman, M., Nelson, S. T., Tingey, D., and D. Eggett. 2010. Regional groundwater flow in structurally-complex extended terranes: an evaluation of the sources of discharge at Ash Meadows, Nevada. Journal of Hydrology 386:118-129.

Caughlan, L., and K. L. Oakley. 2001. Cost considerations for long-term ecological monitoring. Ecological Indicators 1:123-134.

Chapman, D. G., and D. S. Robson. 1960. The analysis of a catch curve. Biometrics 16:354-368.

Crews, S. C., and L. E. Stevens. 2009. Spiders of Ash Meadows National Wildlife Refuge, Nevada. The Southwestern Naturalist 54:331-340.

Deacon, J. E., A. E. Williams, C. D. Williams, and J. E. Williams. 2007. Fueling population growth in Las Vegas, NV: how large scale groundwater withdrawal could burn regional biodiversity. Bioscience 57:688-698.

ESRI, Inc. 2006. ArcGIS 9.2. ESRI, Inc., Redlands, California.

Field, S. A., P. J. O’Connor, A. J. Tyre, and H. P. Possingham. 2007. Austral Ecology 32:485-491.

Frankman, R. 1996. Relationship of genetic variation to population size in wildlife. Conservation Biology 10:1500-1508.

Fraser, J., and C. Martinez. 2002. Restoring a desert oasis. Endangered Species Bulletin 27:18-19.

Hendricks, W. A., and K. W. Robey. 1936. The sampling distribution of the coefficient of variation. Annals of Mathematical Statistics 7:129-132.

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Hockwin, O., V. Dragomirescu, and H. Laser. 1982. Measurements of lens transparency or its disturbances by densitometric image analysis of scheimpflug photographs. Graefe’s Archive for Clinical and Experimental Ophthalmology 219:255-262.

Holdich, D. M. 1999. The negative effects of established crayfish introductions. Pages 31-47 in D. M. Holdich and F. Gherardi, editors. Crayfish in Europe as an alien species: how to make the best of a bad situation. A. A. Balkema Publishers, Rotterdam, Netherlands.

Hurvich, C. M., and C. Tsai. 1989. Regression and time series model selection in small samples. Biometrika 76:297-307.

Kennedy, T. A., J. C. Finlay, and S. E. Hobbie. 2005. Eradication of invasive Tamarix ramosissima along a desert stream increases native fish density. Ecological Applications 15:2072-2083.

Kodric-Brown, A. and J. H. Brown. 2007. Native fishes, exotic mammals, and the conservation of desert springs. Frontiers in Ecology and the Environment 5:549- 553. Koike, H., A. A. Echelon, D. Lofts, and R. A. Van Den Busch. 2008. Microsatellite DNA analysis of success in conserving genetic diversity after 33 years of refuge management for the desert pupfish complex. Animal Conservation 11:321-329. Lema, S. C., and G. A. Nevitt. 2006. Testing an ecophysiological mechanism of morphological plasticity in pupfish and its relevance to conservation efforts for endangered Devils Hole pupfish. The Journal of Experimental Biology 209:3499- 3509. Martin, A. P. 2010. The conservation genetics of Ash Meadows pupfish populations. I. The warm springs pupfish Cyprinodon nevadensis pectoralis. Conservation Genetics 11:1847-1857.

Mettee, M. F., and E. C. Beckham, III. 1978. Notes on the breeding behavior, embryology, and larval development of Cyprinodon variegatus Lacépède in aquaria. Tulane Studies in Zoology and Botany 20:137-148.

Miller, R. R. 1948. The cyprinodont fishes of the Death Valley system of eastern California and southwestern Nevada. Miscellaneous Publications of the Museum of Zoology, University of Michigan 68:1-155.

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Miller, R. R., and J. E. Deacon. 1973. New localities of the rare Warm Springs pupfish, Cyprinodon nevadensis pectoralis, from Ash Meadows, Nevada. Copeia 1973:137-140.

Miller, R. R., J. D. Williams, and J. E. Williams. 1989. Extinction in North American fishes during the past century. Fisheries 14:22-29, 31-38.

Milly, P. C. K. A. Dunne, and A. C. Vecchia. 2005. Global patterns of trends in streamflow and water availability in a changing climate. Nature 438:347-350.

Minckley, W. L. 1995. Translocation as a tool for conserving imperiled fishes: experiences in western United States. Biological Conservation 72:297-309. Minckley, W. L., G. K. Meffe, and D. L. Soltz. 1991. Conservation and management of short-lived fishes: the Cyprinodontoids. Pages 247-282 in W.L. Minckley and J.E. Deacon, editors. Battle against extinction: native fish management in the American West. The University of Arizona Press, Tuscon.

Mott, C. L., S. E. Albert, M. A. Steffen, and J. M Uzzardo. 2010. Assessment of digital image analyses for use in wildlife research. Wildlife Biology 16:93-100.

Nico, L., P. Fuller, and G. Jacobs. 2010. Gambusia affinis. U.S. Geological Survey Nonindigenous Aquatic Species Database, Gainsville, Florida. URL: http://nas.er.usgs.gov/queries/FactSheet.aspx?speciesID=846. Revision Date: 5/5/2001.

Peay, S. 2001. Eradication of alien crayfish populations. R & D Technical Report No. W1-037/TR1.

Peacock, M. M., J. L. Robinson, T. Walters, H. A. Mathewson, and R. Perkins. 2010. The evolutionary significant unit concept and the role of translocated populations in preserving the genetic legacy of Lahontan cutthroat trout. Transactions of the American Fisheries Society 139:382-395.

Pister, E. P. 1974. Desert fishes and their habitats. Transactions of the American Fisheries Society 1974:531-540.

Pister, E. P. 1993. Species in a bucket. Natural History 102:14-19.

R Development Core Team. 2009. R: a language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. ISBN 3- 900051-07-0, URL http://www.R-project.org.

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Robson, D. S., and D. G. Chapman. 1961. Catch curves and mortality rates. Transactions of the American Fisheries Society 90:181-189.

Schlosser, I. J. 1982. Fish community structure and function along two habitat gradients in a headwater stream. Ecological Monographs 52:395-414.

Scoppettone, G. G., P. H. Rissler, C. Gourley, and C. Martinez. 2005. Habitat restoration as a means of controlling non-native fish in a Mojave Desert oasis. Restoration Ecology 113:247-256.

Scoppettone, G. G., Rissler, P., Johnson, D., and M. Hereford. 2009. Relative abundance and distribution of fishes and crayfish at Ash Meadows, Nye County, Nevada. U.S. Geological Survey, unpublished report, Reno.

Seager, R., M. Ting, I. Held, Y. Kushnir, J. Lu, G. Vecchi, H. P. Huang, N. Harnik, A. Leetmaa, N. C. Lau, C. Li, J. Velez, and N. Naik. 2007. Model projections of an imminent transition to a more arid climate in southwestern North America. Science 316:1181-1184.

Sharpe, F. P. 1983. Status report of the desert pupfish sanctuary constructed by the Bureau of Reclamation below Hoover Dam, Clark County, Nevada. Proceedings of the Desert Fishes Council 4(1972):78-80.

Soltz, D. L. 1974. Variation in life history and social organization of some populations of Nevada pupfish, Cyprinodon nevadensis. Ph.D. dissertation, University of California, Los Angeles.

Soltz, D. L. 1976. Vegetation-water relationships at School Spring and the effects on the population dynamics of the Warm Springs pupfish (Cyprinodon nevadensis pectoralis). Submitted to Warm Springs Pupfish Recovery Team, Unpublished report. California State University, Los Angeles.

Soltz, D. L., and R. J. Naiman. 1978. The natural history of native fishes in the Death Valley system. Natural History Museum of Los Angeles County, Science Series 30:1-76.

Soltz, D. L., and M. F. Hirshfield. 1981. Genetic differentiation of pupfishes (genus Cyprinodon) in the American Southwest. Pages 291-333 in R. J. Naiman and D. L. Soltz, editors. Fishes of the North American Deserts. John Wiley & Sons, New York.

Stevens, L. E., and R.A. Bailowitz. 2008. Odonata of Ash Meadows National Wildlife Refuge, southern Nevada, USA. Arizona-Nevada Academy of Sciences 40:128- 135. 37

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Storfer, A. 1999. Gene flow and endangered species translocations: a topic revisited. Biological Conservation 87:173-780.

Sumner, F. B., and M.C. Sargent. 1940. Some observations on the physiology of Warm Spring fishes. Ecology 21:45-54.

Turner, B. J. 1984. Evolutionary genetics of artificial refugium populations of an endangered species, the desert pupfish. Copeia 2:364-369.

USFWS (U.S. Fish and Wildlife Service). 1990. Recovery plan for the endangered and threatened species of Ash Meadows, Nevada. Portland, Oregon.

UTHSCSA (University of Texas Health Science Center at San Antonio). 2002. ImageTool. Version 3.0. URL: http://ddsdx.uthscsa.edu/dig/itdesc.html

Wilcox, J. L., and A. P. Martin. 2006. The devil’s in the details: genetic and phenotypic divergence between artificial and native populations of the endangered pupfish (Cyprinodon diabolis). Animal Conservation 9:316-321.

Williams, J. E. 1977. Observations on the status of the Devils Hole pupfish in the Hoover Dam refugium. U.S. Bureau of Reclamation, unpublished report. Environmental Research Center REC-ERC-77-11, 1-15.

Williams, J. E. 1991. Preserves and refuges for native western fishes: history and management. Pages 171-189 in W. L Minckley and J. E. Deacon, editors. Battle against extinction: native fish management in the American West. The University of Arizona Press, Tucson. Winemiller, K. O., and A. A. Anderson. 1997. Response of endangered desert fish populations to a constructed refuge. Restoration Ecology 5:204-213.

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Table 2.1. Summary of C. n. pectoralis life stages captured (n), adult/juvenile (A/J) ratio, adult density (m-2) and the adult population.

Adult Estimated Survey Captured Adult Juvenile Density Adult Date (n) (n) (n) A/J Ratio (m-2) Population 3/6/2009 220 188 31 6.06 5.88 508 3/18/2009 259 207 48 4.31 6.47 559 4/1/2009 167 137 29 4.72 4.28 370 4/16/2009 201 186 15 12.40 5.81 502 4/29/2009 216 180 33 5.45 5.63 486 5/13/2009 176 160 16 10.00 5.00 432 5/27/2009 218 203 14 14.50 6.34 548 6/10/2009 195 168 24 7.00 5.25 454 6/24/2009 224 198 24 8.25 6.19 535 7/7/2009 216 194 18 10.78 6.06 524 7/21/2009 248 211 32 6.59 6.59 570 8/4/2009 239 200 35 5.71 6.25 540 8/18/2009 205 150 48 3.13 4.69 405 9/1/2009 200 165 30 5.50 5.16 446 9/15/2009 164 121 40 3.03 3.78 327 9/29/2009 94 74 17 4.35 2.31 200 10/13/2009 141 110 30 3.67 3.44 297 10/27/2009 126 98 28 3.50 3.06 265 11/10/2009 126 104 22 4.73 3.25 281 11/24/2009 166 126 40 3.15 3.94 340 12/9/2009 210 186 22 8.45 5.81 502 12/21/2009 156 144 12 12.00 4.50 389 1/6/2010 205 196 9 21.78 6.13 529 1/20/2010 250 208 38 5.47 6.50 562 2/3/2010 82 76 6 12.67 2.38 205 2/16/2010 111 94 17 5.53 2.94 254 3/2/2010 183 151 31 4.87 4.72 408 Average 185.11 156.85 26.26 7.32 4.90 424

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Table 2.2. Number and total length (mm) of C. n. pectoralis captured by habitat type in School Springs refuge.

Habitat N Mean SD Minimum Maximum Pool 3873 25.871 5.8995 9.4809 53.889 Riffle 344 25.145 5.5228 9.8964 41.559 Run 626 24.896 5.2949 11.721 42.579 Wash 154 26.620 5.1605 15.063 42.912

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Table 2.3. Frequency distribution of C. n. pectoralis life stages by reach, estimated area (m2) and estimated water volume (m3)

Larval Juvenile Adult Water (< 14.5 (14.5-20 (> 20.1 Reach Habitat Area Volume mm) mm) mm) Type (m2) (m3) N (%) N (%) N (%) 1 Pool 0.65 0.03 0.00 0.80 3.10 2 Riffle 3.16 0.10 0.00 1.40 0.90 5 Run 3.48 0.16 5.60 11.30 4.40 6 Riffle 0.56 0.01 0.00 0.10 0.30 7 Pool 0.28 0.03 0.00 0.10 0.10 8 Run 0.56 0.03 0.00 0.30 0.20 9 Pool 0.70 0.11 0.00 1.70 0.70 10 Riffle 1.67 0.08 1.90 2.40 1.70 11 Pool 4.18 1.02 37.00 12.70 7.50 12 Riffle 1.11 0.02 1.90 1.70 1.30 13 Pool 6.32 0.29 5.60 3.40 2.90 14 Riffle 0.46 0.01 0.00 0.30 0.40 15 Pool 1.11 0.14 1.90 2.70 1.80 16 Riffle 0.42 0.01 0.00 0.00 0.50 17 Pool 1.11 0.14 0.00 1.00 0.40 18 Riffle 0.56 0.00 3.70 1.80 0.40 19 Pool 2.97 1.81 25.90 38.10 47.90 20 Riffle 0.56 0.01 0.00 0.00 0.10 21 Run 0.93 0.03 3.70 1.50 1.10 22 Run 3.02 0.03 0.00 2.70 3.60 23 Riffle 0.70 0.00 0.00 0.00 0.20 24 Pool 1.11 0.68 1.90 0.80 0.50 25 Riffle 2.60 0.02 3.70 0.10 1.00 26 Pool 1.63 0.30 0.00 1.10 1.30 27 Run 1.02 0.02 0.00 0.00 0.20 28 Pool 2.51 0.38 0.00 1.50 1.40 29 Pool 5.11 1.56 3.70 8.50 10.40 30 Riffle 1.86 0.03 0.00 0.10 0.40 31 Run 4.95 0.50 1.90 1.80 2.30 32 Wash 31.12 0.47 1.90 2.00 3.10 Total 86.42 8.02 100 100 100

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Figure 2.1. Warm Springs Complex and School Springs refuge study area within Ash Meadows National Wildlife Refuge, Amargosa Valley, Nevada.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Figure 2.2. School Springs refuge, Ash Meadows, Nevada, 1969-2010. A. 1969-The spring terminates in a dug-out pool. B. 1983-Four concrete pools were constructed to improve habitat for C. n. pectoralis at School Springs. C. 2008-School Springs was rehabilitated, which included diversifying the habitats for Pyrgulopsis pisteri, Stenelmis calida, Ambrysus relictus, and C. n. pectoralis. D. 2010-Algae were abundant in the largest pool, reach 19. Photograph credit: U.S. Fish and Wildlife Service.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Figure 2.3. School Springs refuge habitat as-built depicting reach segments, as well as pool, run, and riffle habitat types. Note: The wash is not included on this map because its length varies seasonally.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Figure 2.4. Examples of School Springs refuge habitats: pool (A), riffle (B), run (C), and wash (D).

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Texas Tech University, Darrick S. Weissenfluh, December 2010

A

B

Figure 2.5. Comparison of the same digital fish images in original JPEG format (A) and images converted in Image Tool 11 to TIFF format (B). Lengths were determined from TIFF formatted images only. 46

Texas Tech University, Darrick S. Weissenfluh, December 2010

Figure 2.6. Total number of C. n. pectoralis individuals captured each survey in School Springs refuge between March 2009 and March 2010.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Figure 2.7. Length-frequency histogram of C. n. pectoralis captured in School Springs refuge from March 2009 to March 2010. Total lengths displayed on the x-axis are median values for each bin.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Season 1 Season 2

(Winter) (Spring)

(

Ŝ = 0.61 ( Ŝ = 0.66

Ž = 0.22 Ž = 0.18

Season 3 Season 4

(Summer) (Fall)

Ŝ = 0.63 Ŝ = 0.61

Ž = 0.16 Ž = 0.22

Figure 2.8. Catch curves depicting estimated total annual survival rate (Ŝ) and estimated instantaneous total mortality rate (Ž) based on C. n. pectoralis length frequency captured between March 2009 and March 2010. Surveys were grouped so that Season 1: December 21, 2009 – March 19, 2010, Season 2: March 20, 2009 – June 20, 2009, Season 3: June 21, 2009 – September 22, 2009, Season 4: September 23, 2009 – December 20, 2009. Total lengths (mm) are binned identically to Figure 2.6.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Figure 2.9. Capture frequency of C. n. pectoralis by habitat type in School Springs refuge during the course of my study. Reach 19 was sampled more frequently than other reaches.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

CHAPTER III

ENDEMIC AQUATIC INVERTEBRATE DISTRIBUTION AND ASSOCIATION WITH PHYSICAL AND CHEMICAL STREAM PROPERTIES IN SCHOOL SPRINGS REFUGE INTRODUCTION

Freshwater habitats are home to 7% of described species in the world (Balian et al.

2008) and are increasingly at risk of overexploitation by humans (Kreamer and Springer

2008; Darwall et al. 2009). Ricciardi and Rasmussen (1999) predict freshwater species face extinction rates of 4% per decade. It is no surprise, then, that water quality, surface water diversion, and groundwater depletion are major threats facing freshwater aquatic invertebrates (Mehlhop and Vaughn 1994).

Ash Meadows National Wildlife Refuge, Nevada, is home to the largest oasis in the

Mojave Desert and has been set aside to preserve threatened and endangered species

(USFWS 2009). In Ash Meadows, freshwater springs and wetlands have been exploited by groundwater withdrawal and agriculture (Sada and Vineyard 2002). Although agricultural practices have ceased since the Ash Meadows was established in 1984, groundwater withdrawal for municipal use in nearby Las Vegas, Nevada, continues to be a major threat for organisms inhabiting freshwater habitats in surrounding areas (Deacon et al. 2007; Unmack and Minckley 2008). This is of great concern given the number of unique organisms inhabiting Ash Meadows. Among these are 13 endemic aquatic invertebrates.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Ash Meadows is divided into four hydrological management basins (USFWS 2009).

The smallest of these is the Warm Springs Complex (WSC), which supports three endemic aquatic invertebrates, including the median-gland Nevada springsnail,

Pyrgulopsis pisteri, the Devils Hole Warm Springs riffle beetle, Stenelmis calida calida,

and the Ash Meadows Warm Springs naucorid, Ambrysus relictus. Elements of the WSC

have been modified since at least the 1930s (Miller 1948) and one spring, Mexican

Spring, dried in 1973 (Soltz 1974; Kodric-Brown and Brown 2007).

School Springs, a small spring system in Ash Meadows, was the only site renovated in the WSC, although all of the WSC spring systems have altered hydrology and are proposed for renovation (USFWS 2009). In the 1960s, School Springs was modified into an artificial refuge for the endangered Warm Springs pupfish, Cyprinodon nevadensis

pectoralis. School Springs refuge underwent further modification in the 1983 to improve habitat for C. n. pectoralis (see Chapter 2 for details). Sometime after 1983, three aquatic

non-native species became established in School Springs refuge, including red-rimmed

melania, Melanoides tuberculatus, red swamp crayfish, Procambarus clarkii, and

western mosquitofish, Gambusia affinis. I discuss the eradication of M. tuberculatus in

this chapter, as it was easily collected in dip net samples, whereas the eradication of P.

clarkii and G. affinis is discussed in Chapter 2.

School Springs refuge was renovated in 2008 to eradicate aquatic non-native species,

improve habitat for C. n. pectoralis (See Chapter 2), and establish populations of endemic

aquatic invertebrates including P. pisteri, S. c. calida, and A. relictus. It is unclear

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Texas Tech University, Darrick S. Weissenfluh, December 2010 whether P. pisteri, S. c. calida, or A. relictus historically occupied School Springs;

however, neither was present prior to renovation. After School Springs refuge was

renovated, the U.S. Fish and Wildlife Service translocated A. relictus, S. c. calida, and P. pisteri from two other WSC springs, Marsh and North Scruggs, into School Springs

refuge. This was done to establish an additional population of each species at School

Springs to reduce their risk of extinction and so they could be used to re-establish

populations following future restoration of other WSC springs where they had been

extirpated, such as North and South Indian springs.

The purpose of my research is to document the first-year status of P. pisteri, S. c.

calida, and A. relictus, determine if translocations of these species were successful, and determine whether the eradication of one non-native aquatic invertebrate, Melanoides

tuberculatus was successful. Also, because School Springs refuge was designed to

mimic habitats in which P. pisteri, S. c. calida, and A. relictus occur naturally, I tested

two hypotheses implicit in the refuge design:

Ho: P. pisteri, S. c. calida, and A. relictus are distributed at random throughout

School Springs refuge;

Ho: Distribution of P. pisteri, S. c. calida, and A. relictus are independent of

physical and chemical stream properties (slope, substrate, habitat, algae density,

vegetation density, velocity, DO, pH, conductivity, water temperature, and water

depth).

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STUDY AREA

Ash Meadows is a rare desert oasis (Fraser and Martinez 2002) and has more than 50

perennial seeps and springs. It is located approximately 128 kilometers northwest of Las

Vegas, Nevada. It was established to conserve threatened and endangered species in

1984 as a National Wildlife Refuge and it contains 26 endemic species of plants (N = 9),

fish N = 4, and invertebrates N = 13). Ash Meadows contains the largest concentration of

endemic species in the United States (Crews and Stevens 2009).

The WSC (Figure 3.1) consists of six low-discharge warm spring systems with individual flows ranging from 1.13 x 10-4 to 1.98 x 10-4 cm-s and spring-source water

temperatures ranging from 28o to 34oC. School Springs (Figure 3.2) has the lowest flow and highest source water temperature among WSC springs. All three invertebrates are typically found in run and riffle habitats on hard substrates such as pebble and cobble, but

P. pisteri and S. c. calida also may utilize wood and emergent vegetation. Because they

are thermophilic, all three species are typically found in greatest year-round abundance

near the spring source. It is unclear which habitats were present historically in School

Springs refuge; however, because of its elevation and low flow, the system was probably

characterized by dense emergent vegetation with little variability in habitats.

The renovated School Springs refuge channel is approximately 107 m long. Within

the renovated channel, the average width is 0.69 m and the average depth is 0.45 m.

Much of the renovated refuge consists of an artificial base-substrate constructed of

concrete and mortar, except a middle portion of the stream channel, which was

54

Texas Tech University, Darrick S. Weissenfluh, December 2010 constructed without a concrete base, having only the sides mortared. Below the renovated channel is an ephemeral wash, which was unaltered and consists of natural sand, silt, and rock. Its length varies seasonally, flowing above ground more than 100 m

during the winter and less than 40 m during the summer, because of water loss due to

evapotranspiration. Concrete and mortar was used to limit the growth of emergent

vegetation in the renovated channel, especially cattails, Typha domingensis, and S.

americanus, which alter hydrology and promote exotic species populations (Scoppettone

et al. 2005). An additional reason for limiting the growth of vegetation was to maintain

warm waters as far downstream as possible.

METHODS

AQUATIC INVERTEBRATE DISTRIBUTION IN SCHOOL SPRINGS REFUGE

The School Springs stream channel was divided into 32 reaches (Figure 3.3). There

are four distinguishable stream habitat types: pools, runs, riffles, and a wash (Figure 3.4).

Once a month, I selected ten sample sites at random from among the 32 reaches using

Hawth’s Tools (Beyer 2003) and ArcGIS 9.2 (ESRI, Inc. 2006). I selected an additional

five sample sites randomly within reaches 1, 2, and 5 (from 28 May 2009 through the

completion of this study), as these were the reintroduction sites for the endemic

invertebrates. A total of 220 samples were collected during 15 surveys conducted

between February 2009 and March 2010.

Invertebrate samples were collected by placing a 12.7 cm x 9.5 cm steel wire quadrat

in the stream channel with a dip net of the same dimensions placed on the downstream 55

Texas Tech University, Darrick S. Weissenfluh, December 2010 edge of the quadrat, resulting in a sampling area of 120.65 cm2. The substrate within the quadrat was agitated for five seconds to dislodge invertebrates and allow them to wash downstream into the dip net. In areas with pebble or cobble, these substrates were collected in the dip net. The contents of the dip net were then placed into a white plastic tray, where substrates and the dip net were rinsed with water from School Springs refuge to wash invertebrates into the tray in which they were identified with a hand lens. In the case of A. relictus, I used a ruler to estimate carapace length, because lengths correspond to juvenile, < 5 mm, and adult, ≥ 5 mm, instars. This sampling protocol is similar to that used by Parker et al. (2000) and was used to determine presence of invertebrates in

School Springs refuge. An effort was made to identify and enumerate all collected invertebrates in School Springs refuge, but the emphasis was to identify and enumerate

endemic invertebrates.

All endemic aquatic invertebrates were translocated into the upper 20 meters of

School Springs refuge in July 2008, six months before my study commenced. Because

P. pisteri, S. c. calida, and A. relictus are thermophilic, the upper 20 meters was believed

meet their thermal requirements year-round. I used Pearson’s chi-square to test for

deviations from the null hypothesis that 50% of each species would remain at the

translocation site and 50% would disperse. Statistix 9 software was used (Analytical

Software 2008) for all statistical tests and significance was determined when P ≤ 0.05.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

ENDEMIC AQUATIC INVERTEBRATE ASSOCIATION WITH PHYSICAL

AND CHEMICAL STREAM PROPERTIES IN SCHOOL SPRINGS REFUGE

Stream slope, substrate, habitat, algae density, vegetation density, velocity, dissolved oxygen (DO), pH, conductivity, water temperature, and water depth were measured at

each sample site to determine whether these physical and chemical variables influenced

A. relictus, S. c. calida, and P. pisteri distributions. Stream slope was estimated for each

reach as the change in elevation divided by the total length of the reach. Substrate

composition was determined based on sight and texture as silt (< 0.06 mm), sand (0.06-

to 2-mm diameter), granule (2- to 4-mm diameter), pebble (4- to 64-mm diameter),

cobble (64- to 256-mm diameter), or boulder (> 256 mm diameter). Habitat was

determined based on hydrology as pool, run, riffle, or wash. Algae density (%) and

vegetation density (%) were estimated visually as percent cover by centering a 0.5 m x

0.5 m quadrat over each sample site. Velocity (cm-s) was measured using a Marsh-

McBirney Flo-Mate 2000 flow meter. Water temperature, DO, and conductivity were

collected in the field using an YSI 85 meter (Yellow Springs Instruments, Yellow

Springs, OH). pH was measured in the field with a Hanna Combo meter (HI991405,

Hanna, UK). Both meters were calibrated the day of or the day before each survey.

Water temperature was measured with a digital thermometer immediately after each

sample. Water depth was measured to the nearest centimeter using a tape measure. At

the start and end of sampling, air temperature and wind speed were recorded. Minimum,

maximum, mean, standard deviation, and sample size for each variable and each reach

are presented in Appendix B.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

To determine if P. pisteri, S. c. calida, and A. relictus were collected independently of

substrate and habitat, I used Pearson’s chi-square to test the null hypotheses that 25%

would use each substrate and 25% would use each habitat. Because physical and

chemical variables were not normally distributed and log transformations were not useful

in normalizing my data, I used non-parametric Wilcoxon rank-sum tests to determine

whether there was a significant difference among stream reaches and habitat types. I

used Spearman rank correlation to determine if physical and chemical variables were

correlated with the distance from spring source and to explore whether P. pisteri, S. c.

calida, and A. relictus abundance was associated with those same variables. Statistix 9

software was used (Analytical Software 2008) for all statistical tests and significance was

determined at P ≤ 0.05.

RESULTS

NON-NATIVE AQUATIC INVERTEBRATE ERADICATION IN SCHOOL

SPRINGS REFUGE

The temporary desiccation of School Springs refuge to eradicate M. tuberculatus was unsuccessful, as 256 individuals were collected during the course of my sampling. Three

were first collected on 4 February 2009 in reach 1, the spring-source; however, during my

survey on 31 March 2009 they were captured in reach 11, and as far downstream as reach

19 on 3 March 2010. M. tuberculatus was the third most abundant invertebrate in my samples, and as of September 2010, they continue to persist from reach 1 downstream to

reach 19.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

ENDEMIC AQUATIC INVERTEBRATE TRANSLOCATION

Endemic aquatic invertebrates were translocated to School Springs refuge in July

2008 from two source locations, Marsh and North Scruggs springs. To reduce handling

and increase capture efficiency, porous tuffa rock was added to the source locations in

June 2008. Two weeks later in July 2008, the tuffa rock was collected from the source

locations and translocated into School Springs refuge. An effort was made to enumerate

all adult endemic invertebrates prior to placing tuffa rock into School Springs refuge. A

total of 1648 P. pisteri and 24 A. relictus were translocated. Additionally, 1345 riffle

beetles were translocated, including Microcylloepus similis and the endemic S. c. calida,

although they were enumerated collectively due to time constraints and to reduce

handling. Other native invertebrates translocated into School Springs refuge were not

identified or enumerated extensively, but included Hyallela sp., Argia sp., Dugesia sp.,

Tryonia sp., Helicopsyche sp., Elmidae larvae, Ostracoda, Gomphidae, and Baetidae. At

least one of these, Hyallela sp., is known to be preyed on by A. relictus (Parker et al.

2000); however, only one Hyallela sp. was captured during the course of my study.

PERSISTENCE OF ENDEMIC AQUATIC INVERTEBRATES IN SCHOOL

SPRINGS REFUGE

As of August 2010, all three endemic invertebrates were present in School Springs refuge. During the course of this study, I collected 307 P. pisteri, 39 S. c. calida, and 37

A. relictus in my dip-net samples. An additional 21 A. relictus were observed during

casual visual observation of habitats in School Springs refuge, and eight were caught in

59

Texas Tech University, Darrick S. Weissenfluh, December 2010 traps during fish sampling (see Chapter 2). Only A. relictus collected in dip-net samples

and only adult S. c. calida are included in my analyses.

Presence and abundance of the three endemic aquatic invertebrates captured in dip-

net samples varied through the study period. P. pisteri was the most abundant endemic

invertebrate and was captured in 76% of surveys. I collected the smallest number of

individuals in April 2010 and greatest number of individuals in April 2009. On 29 April

2009, 60 (20%) P. pisteri individuals were collected. The second most abundant endemic

invertebrate S. c. calida, was collected in 53% of surveys. The smallest and greatest

number of individuals were collected in January 2010 and April 2010, respectively. On 1

April 2010, 9 (24%) of S. c. calida were collected. Additionally, 83 Microcylloepus

similis individuals were collected. A. relictus were collected in 53% of surveys and were

the least abundant endemic invertebrate. Of the surveys when they were collected, the

fewest were captured in May-June 2009 and the greatest number was collected in April

2010. During a single survey, 1 April 2010, 16 (24%) A. relictus individuals were

sampled with early instars (< 5 mm) comprising 69% of those captured. Invertebrates

from 13 native and one non-native invertebrate genera and an additional five genera of unidentified taxa also were collected (Table 3.1).

Based on my collections, there is evidence that at least one endemic invertebrate, A.

relictus, successfully reproduced during the course of my study. There are four juvenile

and one adult A. relictus instars; all five instars were captured during sampling, although lengths of 3 mm and 4 mm were under-represented in my sampling (Figure 3.5). I also

60

Texas Tech University, Darrick S. Weissenfluh, December 2010 collected 115 Elmidae larvae; however, I could not distinguish between larvae of

Stenelmis and Microcylloepus in the field, so it is unclear whether Stenelmis reproduced.

Also, no egg masses of P. pisteri were collected during sampling, so I do not know if P.

pisteri are successfully reproducing in School Springs refuge.

DISTRIBUTION AND DISPERSAL OF ENDEMIC AQUATIC

INVERTEBRATES IN SCHOOL SPRINGS REFUGE

Endemic aquatic invertebrates initially were translocated into the upper 20 meters of

School Springs refuge, which includes reaches 1, 2, and the upper part of 5. During the

course of my study, P. pisteri was collected only in reaches 1, 2, and 5, and thus, had the

most limited distribution of the three endemic invertebrates studied herein (Figure 3.6).

In contrast, both S. c. calida and A. relictus dispersed varying distances downstream from

the translocation site. S. c. calida were collected in dip-net samples from reach 1 through

reach 30 (Figure 3.7) and A. relictus were collected from reach 2 through reach 32

(Figure 3.8). Elmidae larvae (possibly including both species) comprised of 108 (94%)

individuals were captured in reaches 1, 2, and 5 and as far downstream as reach 21. Both

S. c. calida and A. relictus dispersed beyond the translocation site by April 2009;

however, neither of these species was collected outside of the translocation site between

October and March.

Because they are thermophilic, I expected the majority of P. pisteri, S. c. calida, and

A. relictus individuals to stay within translocation site. P. pisteri were present in 37 of 56

collections in the translocation site, whereas none were collected in 164 samples

61

Texas Tech University, Darrick S. Weissenfluh, December 2010 downstream. Therefore, there is no evidence for P. pisteri dispersal outside of the

translocation site (Table 3.2; χ2 = 130.27, df = 1, P < 0.01). Of 56 collections in the

translocation site, S. c. calida were present in 16 and A. relictus were present in 12.

Although both S. c. calida and A. relictus dispersed from the translocation site, it was not

significant, as S. c. calida were collected in only 7 of 174 samples downstream (Table

3.3; χ2 = 26.34, df = 1, P < 0.01) and A. relictus were collected in only 9 of 174 samples

downstream (Table 3.4; χ2 = 12.29, df = 1, P < 0.01). These results indicate all three

endemic invertebrates occurred more frequently than expected in the upper 20 meters.

For P. pisteri the majority of collections were made in pebble and mud substrates and

pool and riffle habitats. There was no association with substrates (Table 3.5; χ2 = 0.16, df

= 3, P > 0.05); however, there was strong evidence for association with pool and riffle

habitats, but the pool in this case was the spring source (Table 3.6; χ2 = 14.65, df = 3, P <

0.01). S. c. calida showed no association with any substrate type (Table 3.7; χ2 = 1.02, df = 3, P > 0.05), but there was strong evidence of an association with pool and riffle

habitats (Table 3.8; χ2 = 8.06, df = 3, P < 0.05). A. relictus were primarily collected in

pebble, cobble, and mud substrates and riffle habitat. The presence of A. relictus was not

associated with specific substrate (Table 3.9; χ2 = 5.84, df = 3, P > 0.05), but A. relictus

presence was associated with riffle habitat (Table 3.10; χ2 = 14.82, df = 3, P < 0.01).

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Texas Tech University, Darrick S. Weissenfluh, December 2010

ENDEMIC AQUATIC INVERTEBRATE DISTRIBUTION AND THEIR

ASSOCIATION WITH CHEMICAL AND PHYSICAL STREAM PROPERTIES

IN SCHOOL SPRINGS REFUGE

There was significant heterogeneity in physical and chemical variables among

reaches (P < 0.01) and between habitats (P < 0.01) in School Springs refuge. Spearman

rank correlations also were calculated to determine if physical and chemical stream properties were correlated with the distance from spring source. All chemical variables

were significantly correlated with distance from spring source, except velocity. Because

School Springs refuge issues from a single point source with a stable discharge water

temperature, it was expected that these variables were associated with the distance from

spring source.

Across all samples, four chemical variables (conductivity, DO, water temperature,

and pH) were significantly correlated with P. pisteri and S. c. calida abundance, whereas

two chemical variables, DO and water temperature, were significantly correlated with A.

relictus abundance. Slope was not correlated with P. pisteri or S. c. calida abundance,

but was significantly correlated with A. relictus abundance. None of the other physical or

chemical variables measured were correlated with P. pisteri, S. c. calida, or A. relictus

abundance (Table 3.11) and visual analyses of scatterplots of the same variables failed to

reveal any additional correlations (Figures 3.9, 3.10, 3.11).

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Texas Tech University, Darrick S. Weissenfluh, December 2010

DISCUSSION

EVALUATING THE TRANSLOCATION SUCCESS OF ENDEMIC AQUATIC

INVERTEBRATES IN SCHOOL SPRINGS REFUGE

Translocations have played a crucial role in the conservation of imperiled fish

(Minckley 1995) and many other taxa, including amphibians, birds, and mammals

(Griffith et al. 1989; Reinert 1991; Hodder and Bullock 1997; Bullock 1998; Reynolds et

al. 2008). Although translocations of invertebrates have either been conducted less

frequently or are not well-documented in the literature, general guidelines for evaluating

the success of translocations already exist. The criteria for evaluating translocations

generally include selecting an appropriate translocation site, conducting the translocation,

and post-translocation monitoring (Williams et al. 1988). In the following discussion, I

compare the translocation of endemic aquatic invertebrates into School Springs refuge

with the guidelines Williams et al. (1988) discuss in detail.

School Springs refuge was built primarily to maintain a self-sustaining population of

C. n. pectoralis; however, the U.S. Fish and Wildlife Service also desired to establish

populations of P. pisteri, S. c. calida, and A. relictus. Spring-source water temperatures are similar throughout the WSC and both C. n. pectoralis and the endemic aquatic

invertebrates co-occur elsewhere in the WSC, so it is possible that all of the WSC springs

historically supported populations of these species. However, School Springs refuge is

the only spring in the WSC, besides Mexican Spring, which dried in 1973, where neither

P. pisteri, S. c. calida, nor A. relictus were known to have occurred historically. The long

64

Texas Tech University, Darrick S. Weissenfluh, December 2010 history of human disturbance in the WSC may explain their absence in those springs.

Nevertheless, those species are restricted to only a couple of low-flow springs and are unable to disperse to new springs, so they are at great risk of extinction. School Springs refuge was chosen as a translocation site for P. pisteri, S. c. calida, and A. relictus because it is possible they historically occupied the site. However, there were also other considerations.

Another reason for choosing School Springs refuge as a translocation site for P. pisteri, S. c. calida, and A. relictus was that all of the WSC springs have altered

hydrology and the U.S. Fish and Wildlife Service intends to restore those habitats.

Successful establishment of self-sustaining populations of these species at School Springs

refuge means there is less risk of losing those populations at sites selected for restoration.

Whether to establish a single large preserve or several small preserves was debated in the

1970s and 1980s (Diamond 1975; Simberloff and Abele 1976). From this debate, the

concept of spreading of risk emerged as primary argument for several small preserves.

Establishing populations of P. pisteri, S. c. calida, and A. relictus at School Springs

refuge expanded their distribution, to reduce their risk of extinction.

In addition to School Springs, P. pisteri and A. relictus are restricted to just two other

springs, Marsh Spring and North Scruggs (Hershler and Sada 1987; Polhemus and

Polhemus 1994), although the latter species also was present in North and South Indian

springs. Cochran (1949) first described S. c. calida from Devils Hole; however,

Schmude (1999) identified four new populations of S. c. calida located at Point-of-Rocks

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Springs, and Marsh and North Scruggs springs in the WSC, although he was not sure

whether these populations occurred naturally or were translocated.

Two locations, Marsh and North Scruggs springs, were selected by the U.S. Fish and

Wildlife Service as source sites for P. pisteri, S. c. calida, and A. relictus translocation

into School Springs refuge. Those locations were chosen because annual surveys

conducted by the U.S. Fish and Wildlife Service indicated the endemic aquatic

invertebrates were most abundant in those springs. Furthermore, two springs, as opposed

to one, were chosen as source sites to reduce the likelihood of depleting populations at

either site and to increase the genetic variability in the translocated population.

Marsh and North Scruggs springs also were considered ideal source sites because

neither spring was known to contain M. tuberculatus. However, M. tuberculatus was

later captured in Marsh Spring, so it is unclear whether the population in School Springs

refuge came from Marsh Spring, or whether eradication attempts in School Springs

refuge were unsuccessful. These observations emphasize the importance of evaluating

source sites extensively for non-native species before translocations occur.

For invertebrates, there are no guidelines as to how many individuals should be

translocated from a source population, except that a large number generally maintains

genetic variability (Frankman 1996). However, in the case of small spring systems such

as the WSC, it is important to not remove too many individuals. From North Scruggs,

666 P. pisteri, 981 riffle beetles (S. c. calida and M. similis), and 10 A. relictus were

translocated into School Springs refuge, whereas 982 P. pisteri, 364 riffle beetles, and 14

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Texas Tech University, Darrick S. Weissenfluh, December 2010

A. relictus were translocated from Marsh Spring into School Springs refuge. The total

number of individuals of each species translocated is representative of their abundance in

each spring, meaning that P. pisteri and riffle beetles were more abundant then A.

relictus.

Despite the abundance of publications on translocations, monitoring should be improved and success clearly defined (Minckley 1995; Fischer and Lindemayer 2000).

Frequency and intensity of monitoring will always depend on the availability of resources

(time, money, etc.), but should focus on (1) survival, (2) establishment, (3) population growth, and (4) research opportunities, although it is important to consider the efficiency of sampling a species and its life expectancy, as well (Minckley 1995).

I conducted systematic monitoring focused on P. pisteri, S. c. calida, and A. relictus on a monthly basis from February 2009 to March 2010, which is more frequent than the quarterly monitoring proposed by Williams et al. (1989). A. relictus were first captured

during the third survey of my research (April 2009), seven months after their

translocation into School Springs refuge. If I had monitored on a quarterly basis, it may

have taken me longer to detect A. relictus, given their low abundance in School Springs refuge. Also, I monitored School Springs refuge more frequently than quarterly to determine if P. pisteri, S. c. calida, and A. relictus presence was correlated with physical and chemical properties, based on a priori hypotheses.

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Texas Tech University, Darrick S. Weissenfluh, December 2010

EVALUATING THE SUCCESS AND DESIGN OF SCHOOL SPRINGS REFUGE

The placement of habitats in School Springs refuge mostly was contingent on the

topography. However, reaches 1 and 2 were specifically constructed to be pool and riffle

habitat, respectively. Reach 1 was the spring-source, which supplies water to the

remainder of the channel and is the warmest water in the system and was, therefore,

considered important habitat for P. pisteri, S. c. calida, and A. relictus, which are

thermophilic. Riffle habitat, primarily consisting of pebble and sand substrate, was

created in reach 2 because this section of the stream had the greatest slope. Both

substrate and slope have been associated with the presence of P. pisteri, S. c. calida, and

A. relictus elsewhere in the WSC (Parker et al. 2000). Run habitat was created in reach

5, the reach directly below reach 2, because the slope was not great enough to create riffle

habitat. Therefore, reach 1, 2, and 5 were intentionally constructed and placed in their

respective locations in the upper 20 m of School Springs refuge, as those locations were

expected to contain the most suitable habitat for P. pisteri, S. c. calida, and A. relictus.

The spatial distribution of Pyrgulopsis sp. typically occurs near springs and decline in

abundance downstream (Hershler and Sada 1987, 2002; Hershler and Liu 2008). In

School Springs refuge, the same was true for P. pisteri. Throughout the course of my

study, they did not disperse beyond the upper 20 m where they were translocated.

In contrast, both A. relictus and S. c. calida dispersed beyond the translocation site during the summer months. Of the three endemic invertebrates studied, A. relictus is the

most mobile and dispersed to reach 32, the wash, which was the furthest reach from the

68

Texas Tech University, Darrick S. Weissenfluh, December 2010 translocation site. I did not expect to observe A. relictus in the wash, which was

considered to be marginal habitat because of its mud substrate and wide temperature

fluctuations. Based on the limited number of A. relictus collected during the course of

my study, thorough analysis of its dispersal was precluded. However, A. relictus may

disperse to the lower reaches of School Springs refuge to escape high temperatures, to

find prey, or because seasonal abundance exceeds carrying capacity of the upper reaches.

Although only a single A. relictus was observed, it was 90 m downstream from the

nearest introduction site (reach 5). No literature describes the dispersal of naucorids after

translocation, but I suspect the distance of dispersal may be determined by the amount of

suitable habitat or prey abundance, as these factors have been shown to influence the

dispersal of other aquatic invertebrates. I also did not expect S. c. calida to disperse as

far as reach 30, approximately 60 m from the nearest introduction site. The dispersal of

A. relictus and S. c. calida downstream is biologically significant, especially if they

successfully reproduce beyond the translocation site or, in contrast, if their dispersal

downstream leads to isolation so that individuals are unable to reproduce. Therefore,

future monitoring should determine whether downstream dispersal of A. relictus and S. c.

calida is genetically beneficial.

I observed two A. relictus individuals on shallow boulder shelves in reach 19, the

largest pool; however, I observed 21 individuals in pools at night. In May 2010, I also

observed six S. c. calida during night surveys of pool habitats. Based on these

observations, both A. relictus and S. c. calida may disperse at night to avoid predation by

C. n. pectoralis. Further study is necessary to assess whether S. c. calida and A. relictus 69

Texas Tech University, Darrick S. Weissenfluh, December 2010 are active at night in other WSC springs and whether night surveys for these species are practical.

A diversity of habitats, including pools, runs, and riffles, were constructed throughout

School Springs refuge to meet the habitat needs of C. n. pectoralis, P. pisteri, S. c. calida,

and A. relictus. Based on previous studies (Hershler and Sada 1987; Parker et al. 2000),

P. pisteri, S. c. calida, and A. relictus primarily occupy spring source and riffle habitats.

My results are consistent with those observations. Both P. pisteri and S. c. calida were

strongly associated with pool (spring source) and riffle habitats. Although A. relictus was

not captured in the spring source during the course of my study, their presence was

strongly associated with riffle habitat. Therefore, these habitat types should be included

in restoration projects involving these species, so long as water temperature also is

sufficiently high.

Prior to renovation, the predominant substrate in School Springs refuge was mud.

Although no mud was added intentionally after renovation, mud has blown into the

channel and forms the most abundant substrate in School Springs refuge. Pebble and

cobble substrate was added to the channel to diversify the substrate; however, neither P.

pisteri, S. c. calida, nor A. relictus were strongly associated with any particular substrate.

Individuals of all three species were captured in sand, pebble, cobble, and mud substrates;

however, P. pisteri and A. relictus were present in mud substrate more than any other

substrate and S. c. calida were present in mud substrate more than any other substrate

except pebble. However, the substrate at almost every sample site included some mud

70

Texas Tech University, Darrick S. Weissenfluh, December 2010 substrate, which makes it unclear whether it is important for P. pisteri, S. c. calida, and A.

relictus.

To my knowledge, no study has associated P. pisteri, S. c. calida, or A. relictus

presence with physical and chemical stream properties. The presence of P. pisteri and S.

c. calida was uncorrelated with physical variables; however, slope was correlated with A. relictus presence. Both P. pisteri and S. c. calida presence were correlated with all four

chemical variables, whereas the presence of A. relictus only was correlated with DO.

In conclusion, available habitat and substrate in School Springs refuge is able to sustain the translocated populations of P. pisteri, S. c. calida, and A. relictus, at least in the short term. Importantly, at least one species, A. relictus has successfully reproduced

in School Springs refuge based on presence of early instars. In spite of this, it is unclear

whether P. pisteri or S. c. calida is reproducing, as neither eggs nor larvae of either

species has been positively identified. However, both P. pisteri and S. c. calida are short-

lived species, which suggests they must have successfully reproduced in School Springs

refuge.

MANAGEMENT RECOMMENDATIONS

The translocation of P. pisteri, S. c. calida, and A. relictus appears successful;

however, additional translocations of these species and other native species may be

warranted. Throughout the duration of my study, only 307 adult P. pisteri, 39 adult S. c.

calida and 37 adult A. relictus were collected. Although these species have persisted for

more than two years since they were translocated, their abundances are low. However, 71

Texas Tech University, Darrick S. Weissenfluh, December 2010 before additional individuals are translocated, I recommend the U.S. Fish and Wildlife

Service complete comprehensive surveys of Marsh, North and South Scruggs springs, and School Springs refuge for P. pisteri, S. c. calida, and A. relictus during March-June

to determine their relative abundances and establish goals relating to minimum

population size of each species in each spring.

It also may be desirable to translocate additional Hyallela sp. into School Springs

refuge, which is typically one of the most abundant invertebrates in other WSC springs

(Parker et al. 2000). Only one individual was captured during my surveys, despite the

translocation of more than 1,000 individuals. Therefore, it appears the Hyallela sp.

translocation was unsuccessful. Because Parker et al. (2000) suggested Hyallela sp. is a

primary food source of A. relictus, I suspect they may be an important element of the

invertebrate community missing from School Springs refuge.

If additional endemic aquatic invertebrate translocations to School Springs refuge occur, I suggest the U.S. Fish and Wildlife Service continue monitoring on a monthly basis. It is clear from my results that it may take several months to detect translocated species, such as A. relictus. Alternatively, if no further translocations occur, it may be

sufficient to continue monitoring P. pisteri, S. c. calida, and A. relictus on a quarterly

basis, to verify persistence of these species.

There are other objectives that may increase or decrease the frequency of monitoring,

such as the genetic variability of P. pisteri, S. c. calida, and A. relictus, as well as the

presence of non-native species. The number of invertebrates necessary to maintain

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Texas Tech University, Darrick S. Weissenfluh, December 2010 genetic variability is not well-understood. It may be desirable to evaluate the genetic

variability of each P. pisteri, S. c. calida, and A. relictus population to ensure their

genetic variation does not bottleneck. At least one non-native aquatic invertebrate, M.

tuberculatus, currently is present in School Springs refuge and other non-native aquatic

invertebrate invasions are possible from downstream springs. It is clear from my study

that M. tuberculatus continued to disperse downstream in School Springs refuge, so it

may be desirable to continue monitoring its dispersal and determine whether M.

tuberculatus negatively influence the presence of P. pisteri, S. c. calida, and A. relictus.

If so, monitoring more frequently, may be warranted.

Ultimately, the continuation of monitoring, and the frequency with which it is

conducted, should depend on the objectives set forth. Therefore, before any additional

monitoring occurs, I suggest the U.S. Fish and Wildlife Service clearly define success

and outline its objectives with regard to these species in School Springs refuge to ensure

the results of any monitoring activities will be able to answer the questions which arise.

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LITERATURE CITED

Analytical software. 2008. STATISTIX 9, user’s manual version 9. Analytical Software, Tallahassee, Florida.

Balian, E. V., H. Segers, K. Martens, and C. Leveque. 2008. The freshwater animal diversity assessment: an overview of results. Hydrobiologia 595:627-637.

Beyer, H. 2003. Hawth’s Analysis Tools. URL: http://www.SpatialEcology.com.

Bullock, J. M. 1998. Community translocation in Britain: setting objectives and measuring consequences. Biological Conservation 84:199-214.

Crews, S. C., and L. E. Stevens. 2009. Spiders of Ash Meadows National Wildlife Refuge, Nevada. The Southwestern Naturalist 54:331-340.

Darwall, W. R. T., K. G. Smith, D. Allen, M. B. Seddon, G. M. Reid, V. Clausnitzer, and V. J. Kalkman. 2009. Freshwater biodiversity: a hidden resource under threat. Pages 43-54 in J.-C.Vié, C. Hilton-Taylor, and S. N. Stuart, editors. Wildlife in a changing world – an analysis of the 2008 IUCN red list of threatened species. Gland, Switzerland: IUCN.

Deacon, J. E., and M. S. Deacon. 1979. Research on endangered fishes in the National Parks with special emphasis on the Devils Hole pupfish. Pages 9-19 in R. M. Linn, editor. Proceedings of the first conference on scientific research in the National Parks. U. S. National Park Service Transactions and Proceedings Series 5. Washington D.C.

Deacon, J. E., A. E. Williams, C. D. Williams, and J. E. Williams. 2007. Fueling population growth in Las Vegas, NV: how large scale groundwater withdrawal could burn regional biodiversity. Bioscience 57:688-698.

Diamond, J. M. 1975. Assembly of species communities. Pages 342-444 in M. L. Cody and J. M. Diamond, editors. Ecology and Evolution of Communities. Belknap Press, Cambridge, MA. ESRI, Inc. 2006. ArcGIS 9.2. ESRI, Inc., Redlands, CA.

Fischer, J., and D. B. Lindenmayer. 2000. An assessment of the published results of animal relocations. Biological Conservation 96:1-11.

Frankman, R. 1996. Relationship of genetic variation to population size in wildlife. Conservation Biology 10:1500-1508.

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Fraser, J., and C. Martinez. 2002. Restoring a desert oasis. Endangered Species Bulletin 27:18-19.

Griffith, B., J. M. Scott, J. W. Carpenter, and C. Reed. 1989. Translocation as a species conservation tool: status and strategy. Science 245:477-480.

Hershler, R., and D. W. Sada. 1987. Springsnails (Gastropoda: Hydrobiidae) of Ash Meadows, Amargosa basin, California-Nevada. Proceedings of the Biological Society of Washington 100: 776-843.

Hershler, R., and D. W. Sada. 2002. Biogeography of great basin aquatic snails of the genus Pyrgulopsis. Smithsonian Contributions to the Earth Sciences 33: 255-276.

Hershler, R., and H-P. Liu. 2008. Ancient vicariance and recent dispersal of springsnails (Hydrobiidae: Pyrgulopsis) in the Death Valley system, California-Nevada. The Geological Society of America 439:91-101.

Hodder, K. H., and J. M. Bullock. Translocations of native species in the UK: implications for biodiversity. Journal of Applied Ecology 34:547-565.

Kodric-Brown, A. and J. H. Brown. 2007. Native fishes, exotic mammals, and the conservation of desert springs. Frontiers in Ecology and the Environment 5:549- 553. Kreamer, D. K., and A. E. Springer. 2008. The hydrology of desert springs in North America. Pages 35-48 in L. E. Stevens and V. J. Meretsky, editors. Aridland springs in North America: ecology and conservation. University of Arizona Press, Tucson.

Mehlhop, P. and C. C. Vaughn. 1994. Threats to and sustainability of ecosystems for freshwater mollusks. Pages 68-77 in W. Covington and L. F. Dehand, editors. Sustainable ecological systems: implementing an ecological approach to land management. General Technical Report RM-247, U.S. Forest Service, Rocky Mountain Range and Forest Experiment Station, Fort Collins, CO.

Miller, R. R. 1948. The cyprinodont fishes of the Death Valley system of eastern California and southwestern Nevada. Miscellaneous Publications of the Museum of Zoology, University of Michigan 68:1-155.

Minckley, W. L. 1995. Translocation as a tool for conserving imperiled fishes: experiences in western United States. Biological Conservation 72:297-309.

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Parker, M. S., G. G. Scoppettone, and M. B. Neilson. 2000. Ecological investigation of two naucorid species (Ambrysus amargosus and A. relictus) endemic to thermal springs of the Ash Meadows National Wildlife Refuge. Prepared for U.S. Fish and Wildlife Service, unpublished report. FWS Document Control Number 14320-8-6002:1-53.

Polhemus, J. T, and D. A. Polhemus. 1994. A new species of Ambrysus STÅL (sic.) from Ash Meadows, Nevada (Heteroptera: Naucoridae). Journal of New York Entomological Society 102:261-265.

Reinert, H. K. 1991. Translocation as a conservation strategy for amphibians and reptiles: some comments, concerns, and observations. Herpetologica 47:357-363.

Ricciardi, A., and J. B. Rasmussen. 1999. Extinction rates of North American freshwater fauna. Conservation Biology 13:1220-1222.

Reynolds, M. H., N. E. Seavy, M. S. Vekasy, J. L. Klavitter, and L. P. Laniawe. 2008. Translocation and early post-release demography of endangered Laysan teal. Animal Conservation 11:160-168.

Sada, D. W., and G. L. Vinyard. 2002. Anthropogenic changes in biogeography of great basin aquatic biota. Smithsonian Contributions to the Earth Sciences 33:277-293.

Schmude, K. L. 1999. Riffle beetles in the genus Stenelmis (Coleoptera: Elmidae) from warm springs in southern Nevada: new species, new status, and a key. Entomological News 110:1-12.

Scoppettone, G. G, P. H. Rissler, C. Gourley, and C. Martinez. 2005. Habitat restoration as a means of controlling non-native fish in a Mojave Desert oasis. Restoration Ecology 113:247-256.

Simberloff, D. S., and L. G. Abele. 1976. Experimental zoogeography of islands: effects of island size. Ecology 57:629-648. Soltz, D. L. 1974. Variation in life history and social organization of some populations of Nevada pupfish, Cyprinodon nevadensis. Ph.D. thesis, University of California, Los Angeles.

Unmack, P.J., and W. L. Minckley. 2008. The demise of desert springs. Pages 12-34 in L. E. Stevens and V. J. Meretsky, editors. Aridland springs in North America: ecology and conservation. University of Arizona Press, Tucson.

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USFWS (U.S. Fish and Wildlife Service). 2009. Desert National Wildlife Refuge Complex: Final Comprehensive Conservation Plan and Environmental Impact Statement. URL: http://www.fws.gov/desertcomplex/ccp.htm. Williams, J. E., D. W. Sada, C. D. Williams, and Other Members of the Western Division Endangered Species Committee. 1988. Introductions of threatened and endangered fishes. Fisheries 13:5-11.

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Table 3.1. The total number of each invertebrate family or genus (italized), arranged by abundance, sampled in School Springs refuge, Ash Meadows National Wildlife Refuge, Nevada.

Family or Genus Abundance Tryonia 716 Pyrgulopsis* 307 Melanoides** 256 Elmidae larvae*** 108 Chironomidae 84 Microcylloepus 83 Baetis 47 Dugesia 42 Stenelmis* 39 Ambrysus* 37 Argia 38 Chrysomelidae 26 Ceratopogonidae 20 Gomphidae 17 Ostracoda 13 Culicidae larvae 12 Unknown 5 Hyallela 1 Helicopsyche 1 *Endemic to Ash Meadows National Wildlife Refuge **Non-Native ***Includes two species of riffle beetle larvae (genera Stenelmis and Microcylloepus)

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Table 3.2. Contingency table (2 x 2) showing presence-absence frequencies of Pyrgulopsis pisteri in dip net samples from the upper 20 meters of School Springs refuge versus the remainder of the spring. χ2 = 130.27, df = 1, P < 0.01.

Presence Absence Total Upper 20 m 37 19 56 Remainder of Spring 0 164 174 37 183 220

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Table 3.3. Contingency table (2 x 2) showing presence-absence frequencies of Stenelmis calida calida in dip net samples from the upper 20 meters of School Springs refuge versus the remainder of the spring. χ2 = 26.34, df = 1, P < 0.01.

Presence Absence Total Upper 20 m 16 40 56 Remainder of Spring 7 157 174 23 197 220

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Table 3.4. Contingency table (2 x 2) showing presence-absence frequencies of Ambrysus relictus in dip net samples from the upper 20 meters of School Springs refuge versus the remainder of the spring. χ2 = 12.29, df = 1, P < 0.01.

Presence Absence Total Upper 20 m 12 44 56 Remainder of Spring 9 155 174 21 199 220

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Table 3.5. Contingency table (4 x 2) showing presence-absence frequencies of Pyrgulopsis pisteri in dip net samples from School Springs refuge collected in each substrate type. χ2 = 0.16, df = 3, P > 0.05.

Presence Absence Total Sand 4 23 27 Pebble 12 57 69 Cobble 5 22 27 Mud 16 81 97 37 183 220

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Table 3.6. Contingency table (4 x 2) showing presence-absence frequencies of Pyrgulopsis pisteri in dip net samples from School Springs refuge collected in each habitat type. χ2 = 14.65, df = 3, P < 0.01.

Presence Absence Total Pool 12 45 57 Riffle 21 60 81 Run 4 49 53 Wash 0 29 29 37 183 220

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Table 3.7. Contingency table (4 x 2) showing presence-absence frequencies of Stenelmis calida calida in dip net samples from School Springs refuge collected in each substrate type. χ2 = 1.02, df = 3, P > 0.05.

Presence Absence Total Sand 3 24 27 Pebble 9 60 69 Cobble 3 24 27 Mud 8 89 97 23 197 220

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Table 3.8. Contingency table (4 x 2) showing presence-absence frequencies of Stenelmis calida calida in dip net samples from School Springs refuge collected in each habitat type. χ2 = 8.06, df = 3, P < 0.05.

Presence Absence Total Pool 10 47 57 Riffle 10 71 81 Run 3 50 53 Wash 0 29 29 23 197 220

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Table 3.9. Contingency table (4 x 2) showing presence-absence frequencies of Ambrysus relictus in dip net samples from School Springs refuge collected in each substrate type. χ2 = 5.84, df = 3, P > 0.05.

Presence Absence Total Sand 2 25 27 Pebble 6 63 69 Cobble 6 21 27 Mud 7 90 97 21 199 220

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Table 3.10. Contingency table (4 x 2) showing presence-absence frequencies of Ambrysus relictus in dip net samples from School Springs refuge collected in each habitat type. χ2 = 14.82, df = 3, P < 0.01.

Presence Absence Total Pool 0 57 57 Riffle 15 66 81 Run 5 48 53 Wash 1 28 29 21 199 220

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Table 3.11. Spearman rank correlations, corrected for ties, detailing how Pyrgulopsis pisteri, Stenelmis calida calida, and Ambrysus relictus abundance is correlated with chemical and physical variables in School Springs refuge. Significant P - values are denoted by * P < 0.05 and ** P < 0.01. Pyrgulopsis Stenelmis calida Ambrysus pisteri calida relictus (r) (r) (r) Conductivity (μS) 0.3830 ** 0.2610 ** 0.1040 DO (mg/l) -0.4610 ** -0.2980 ** -0.1450 * pH -0.5070 ** -0.1870 ** -0.0142 Water Temperature (°C) 0.4520 ** 0.2860 ** 0.1530 * Velocity (cm/s) 0.0954 0.0791 0.1980 Water Depth (cm) -0.0088 0.0946 -0.1040 Algae Density (%) -0.0168 0.0524 0.0200 Vegetation Density (%) -0.0398 0.1300 -0.1360 Slope 0.0773 -0.0416 0.2190 **

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Figure 3.1. Warm Springs Complex and School Springs refuge study area within Ash Meadows National Wildlife Refuge, Nevada.

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Figure 3.2. School Springs refuge, Ash Meadows, Nevada, 1969-2010. A. 1969-The spring terminates in a dug-out pool. B. 1983-Four concrete pools were constructed to improve habitat for C. n. pectoralis at School Springs. C. 2008-School Springs was rehabilitated, which included diversifying the habitats for Pyrgulopsis pisteri, Stenelmis calida calida, Ambrysus relictus, and C. n. pectoralis. D. 2010-Algae were abundant in the largest pool, reach 19. Photograph credit: U.S. Fish and Wildlife Service.

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Figure 3.3. School Springs refuge habitat as-built depicting reach segments, as well as pool, run, and riffle habitat types. Note: The wash is not included on this map because its length varies seasonally.

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Figure 3.4. Examples of School Springs refuge habitats: pool (A), riffle (B), run (C), and wash (D).

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Figure 3.5. Ambrysus relictus length-frequency histogram of all individuals captured during dip net sampling in School Springs refuge from February 2009 to April 2010.

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Figure 3.6. Locations in School Springs refuge where Pyrgulopsis pisteri was collected.

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Figure 3.7. Locations in School Springs refuge where Stenelmis calida calida was collected. 95

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Figure 3.8. Locations in School Springs refuge where Ambrysus relictus was collected.

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Figure 3.9. Scatterplot analyses depicting Pyrgulopsis pisteri abundance with chemical and physical stream properties in School Springs refuge.

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Figure 3.10. Scatterplot analyses depicting Stenelmis calida calida abundance with chemical and physical stream properties in School Springs refuge.

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Figure 3.11. Scatterplot analyses depicting A. relictus abundance with chemical and physical stream properties in School Springs refuge.

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APPENDIX A

SCHOOL SPRINGS REFUGE CHANNEL CHARACTERISTICS FOR EACH REACH Table 4.1. Habitat type, slope, surface area, reach length, mean reach width, and mean reach depth for each reach in School Springs refuge.

Reach Reach Surface Length Mean Reach Mean Reach ID Habitat Slope Area (%) (m) Width (m) Depth (m) 1 1Pool 0.001 0.01 4.270 0.15 0.05 2 Riffle 0.090 0.06 10.36 0.30 0.03 3 2Pool 0.000 0.00 1.520 0.61 1.25 4 3Dry 0.060 0.00 7.620 0.30 0.00 5 Run 0.020 0.06 7.620 0.46 0.05 6 Riffle 0.100 0.01 1.220 0.46 0.02 7 Pool 0.000 0.01 0.610 0.46 0.09 8 Run 0.120 0.01 1.830 0.30 0.05 9 Pool 0.000 0.01 0.910 0.76 0.15 10 Riffle 0.050 0.03 3.660 0.46 0.05 11 Pool 0.000 0.08 2.740 1.52 0.24 12 Riffle 0.090 0.02 2.440 0.46 0.02 13 Pool 0.020 0.11 5.180 1.22 0.05 14 Riffle 0.130 0.01 1.520 0.30 0.03 15 Pool 0.000 0.02 1.220 0.91 0.12 16 Riffle 0.080 0.01 0.910 0.46 0.02 17 Pool 0.000 0.02 1.220 0.91 0.12 18 Riffle 0.090 0.01 0.910 0.61 0.01 19 Pool 0.000 0.05 1.220 2.44 0.61 20 Riffle 0.140 0.01 1.830 0.30 0.02 21 Run 0.060 0.02 3.050 0.30 0.03 22 Run 0.030 0.05 3.960 0.76 0.01 23 Riffle 0.130 0.01 1.520 0.46 0.01 24 Pool 0.000 0.02 1.220 0.91 0.61 25 Riffle 0.110 0.05 4.270 0.61 0.01 26 Pool 0.000 0.03 1.520 1.07 0.18 27 Run 0.080 0.02 3.350 0.30 0.02 28 Pool 0.005 0.05 1.830 1.37 0.15 29 Pool 0.005 0.09 3.050 1.68 0.30 30 Riffle 0.120 0.03 6.100 0.30 0.02 31 Run 0.050 0.09 21.64 0.23 0.03 32 Wash 0.040 0.00 102.1 0.30 0.02 1Spring-source north 2Spring-source south 3Not sampled

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APPENDIX B

SUMMARY STATISTICS FOR PHYSICAL AND CHEMICAL STREAM CHARACTERISTICS IN EACH REACH FROM SCHOOL SPRINGS REFUGE. ALL SAMPLES FROM 6 MARCH 2009 TO 3 MARCH 2010 WERE COMBINED Table 4.2. Summary statistics for conductivity, DO, algae density, vegetation density, salinity, TDS, velocity, water depth, water temperature, and pH in each reach of School Springs refuge.

Reach 1 (Pool) N Mean SD Minimum Maximum Conductivity (mS) 133 787.67 24.51 732.00 823.00 DO (mg/l) 133 2.75 0.52 1.73 4.40 Algae Density (%) 133 41.02 19.84 0.00 90.00 Vegetation Density (%) 133 39.66 17.40 10.00 85.00 Salinity (ppt) 133 0.30 0.00 0.30 0.30 TDS (ppm) 133 353.40 10.90 325.00 366.00 Velocity (cm/s) 118 4.46 5.93 0.00 29.00 Water Depth (cm) 133 2.69 0.74 1.25 4.00 Water Temperature (°C) 133 33.87 1.90 24.90 35.60 pH 133 7.43 0.18 7.11 7.72

Reach 2 (Riffle) N Mean SD Minimum Maximum Conductivity (mS) 48 782.02 20.33 730.00 807.00 DO (mg/l) 48 4.98 0.71 3.15 5.92 Algae Density (%) 48 10.40 15.38 0.00 45.00 Vegetation Density (%) 48 1.25 1.39 0.00 5.00 Salinity (ppt) 48 0.30 0.00 0.30 0.30 TDS (ppm) 48 346.06 15.22 325.00 366.00 Velocity (cm/s) 44 6.47 5.78 1.00 20.12 Water Depth (cm) 48 1.43 0.22 1.00 1.75 Water Temperature (°C) 48 33.42 1.03 31.20 35.20 pH 48 7.60 0.16 7.26 8.03

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Reach 5 (Run) N Mean SD Minimum Maximum Conductivity (mS) 257 780.79 24.75 695.00 820.00 DO (mg/l) 257 4.92 0.98 2.71 7.78 Algae Density (%) 257 17.55 15.87 0.00 90.00 Vegetation Density (%) 257 4.28 3.64 0.00 25.00 Salinity (ppt) 257 0.30 0.00 0.30 0.30 TDS (ppm) 257 347.23 10.53 322.00 371.00 Velocity (cm/s) 257 6.22 4.18 0.00 18.00 Water Depth (cm) 257 2.21 1.04 1.20 6.00 Water Temperature (°C) 257 33.06 1.66 30.40 35.40 pH 257 7.45 0.16 7.45 8.27

Reach 6 (Riffle) N Mean SD Minimum Maximum Conductivity (mS) 15 784.40 11.04 764.00 794.00 DO (mg/l) 15 6.96 0.66 4.71 7.28 Algae Density (%) 15 31.00 19.20 0.00 50.00 Vegetation Density (%) 15 1.27 1.98 0.00 5.00 Salinity (ppt) 15 0.30 0.00 0.30 0.30 TDS (ppm) 15 352.07 4.23 343.00 357.00 Velocity (cm/s) 15 8.47 5.59 2.00 18.00 Water Depth (cm) 15 3.89 4.13 1.20 11.80 Water Temperature (°C) 15 33.75 0.88 31.30 34.40 pH 15 7.87 0.11 7.73 8.14

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Reach 7 (Pool) N Mean SD Minimum Maximum Conductivity (mS) 6 686.33 103.28 553.00 753.00 DO (mg/l) 6 4.64 0.30 4.25 4.84 Algae Density (%) 6 1.67 2.58 0.00 5.00 Vegetation Density (%) 6 0.33 0.52 0.00 1.00 Salinity (ppt) 6 0.27 0.05 0.20 0.30 TDS (ppm) 6 348.67 6.71 340.00 353.00 Velocity (cm/s) 6 2.00 1.55 1.00 4.00 Water Depth (cm) 6 3.17 0.52 2.50 3.50 Water Temperature (°C) 6 31.43 0.88 30.30 32.00 pH 6 7.94 0.11 7.79 8.01

Reach 8 (Run) N Mean SD Minimum Maximum Conductivity (mS) 8 730.00 17.20 690.00 745.00 DO (mg/l) 8 5.45 0.21 5.32 5.78 Algae Density (%) 8 10.63 6.23 0.00 15.00 Vegetation Density (%) 8 0.00 0.00 0.00 0.00 Salinity (ppt) 8 0.30 0.00 0.30 0.30 TDS (ppm) 8 354.00 2.78 348.00 357.00 Velocity (cm/s) 8 7.71 1.92 3.00 8.53 Water Depth (cm) 8 1.66 0.27 1.25 2.00 Water Temperature (°C) 8 30.00 0.71 29.50 31.00 pH 8 7.80 0.11 7.69 7.97

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Reach 9 (Pool) N Mean SD Minimum Maximum Conductivity (mS) 44 768.68 24.65 727.00 797.00 DO (mg/l) 44 6.03 1.28 5.20 8.80 Algae Density (%) 44 8.30 11.96 0.00 35.00 Vegetation Density (%) 44 8.30 10.11 0.00 35.00 Salinity (ppt) 44 0.30 0.00 0.30 0.30 TDS (ppm) 44 349.95 10.30 342.00 366.00 Velocity (cm/s) 44 2.86 2.59 1.00 8.00 Water Depth (cm) 44 3.07 0.56 2.50 4.25 Water Temperature (°C) 44 32.42 1.20 30.50 34.60 pH 44 8.01 0.02 7.94 8.03

Reach 10 (Riffle) N Mean SD Minimum Maximum Conductivity (mS) 92 735.66 57.31 585.00 797.00 DO (mg/l) 92 6.07 1.33 3.62 8.32 Algae Density (%) 92 9.45 9.43 0.00 35.00 Vegetation Density (%) 92 0.28 0.65 0.00 5.00 Salinity (ppt) 92 0.30 0.00 0.30 0.30 TDS (ppm) 92 348.65 11.16 318.00 363.00 Velocity (cm/s) 78 12.86 6.37 4.00 28.65 Water Depth (cm) 92 2.05 1.35 0.50 8.63 Water Temperature (°C) 92 31.66 1.85 28.70 35.00 pH 92 8.02 0.16 7.71 8.42

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Reach 11 (Pool) N Mean SD Minimum Maximum Conductivity (mS) 403 767.20 29.48 711.00 808.00 DO (mg/l) 403 6.86 1.28 4.22 8.79 Algae Density (%) 403 22.04 20.82 0.00 60.00 Vegetation Density (%) 403 0.28 0.55 0.00 5.00 Salinity (ppt) 403 0.30 0.00 0.30 0.30 TDS (ppm) 403 343.57 10.42 323.00 358.00 Velocity (cm/s) 403 1.38 1.08 0.00 4.00 Water Depth (cm) 403 5.48 3.11 0.88 9.50 Water Temperature (°C) 403 32.61 2.58 27.70 35.70 pH 403 8.10 0.21 7.80 8.55

Reach 12 (Riffle) N Mean SD Minimum Maximum Conductivity (mS) 53 751.72 30.86 709.00 796.00 DO (mg/l) 53 6.80 0.85 4.60 8.20 Algae Density (%) 53 15.49 20.77 0.00 80.00 Vegetation Density (%) 53 2.57 1.87 0.00 5.00 Salinity (ppt) 53 0.30 0.00 0.30 0.30 TDS (ppm) 53 348.11 13.80 317.00 364.00 Velocity (cm/s) 38 12.29 8.37 4.00 33.00 Water Depth (cm) 53 2.02 0.74 0.50 4.50 Water Temperature (°C) 53 31.89 2.38 28.00 35.30 pH 53 8.10 0.16 7.73 8.49

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Reach 13 (Pool) N Mean SD Minimum Maximum Conductivity (mS) 160 752.52 25.61 701.00 795.00 DO (mg/l) 160 6.93 1.03 3.56 7.95 Algae Density (%) 160 47.09 34.86 0.00 95.00 Vegetation Density (%) 160 55.81 24.87 20.00 100.00 Salinity (ppt) 160 0.30 0.00 0.30 0.30 TDS (ppm) 160 341.63 19.19 299.00 360.00 Velocity (cm/s) 137 3.92 3.79 0.00 17.00 Water Depth (cm) 160 2.96 0.90 1.00 5.00 Water Temperature (°C) 160 31.97 1.93 28.00 35.20 pH 160 8.23 0.18 7.80 8.51

Reach 14 (Riffle) N Mean SD Minimum Maximum Conductivity (mS) 22 738.82 47.44 687.00 800.00 DO (mg/l) 22 6.64 1.23 3.73 8.19 Algae Density (%) 22 4.82 6.91 0.00 20.00 Vegetation Density (%) 22 44.59 44.96 0.00 95.00 Salinity (ppt) 22 0.30 0.00 0.30 0.30 TDS (ppm) 22 336.45 12.06 310.00 356.00 Velocity (cm/s) 22 9.13 5.25 4.27 19.00 Water Depth (cm) 22 1.55 0.13 1.25 1.80 Water Temperature (°C) 22 31.46 3.36 27.80 35.50 pH 22 8.05 0.10 7.87 8.19

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Reach 15 (Pool) N Mean SD Minimum Maximum Conductivity (mS) 104 730.02 31.18 696.00 779.00 DO (mg/l) 104 6.11 0.93 4.46 7.12 Algae Density (%) 104 31.35 37.81 0.00 90.00 Vegetation Density (%) 104 1.62 1.91 0.00 5.00 Salinity (ppt) 104 0.30 0.00 0.30 0.30 TDS (ppm) 104 338.33 10.93 321.00 350.00 Velocity (cm/s) 104 3.25 3.55 1.00 10.00 Water Depth (cm) 104 8.87 4.30 6.00 17.00 Water Temperature (°C) 104 29.98 2.71 27.20 34.30 pH 104 8.07 0.12 7.84 8.15

Reach 16 (Riffle) N Mean SD Minimum Maximum Conductivity (mS) 15 716.87 31.45 667.00 792.00 DO (mg/l) 15 6.42 0.79 5.56 8.13 Algae Density (%) 15 5.07 6.77 0.00 20.00 Vegetation Density (%) 15 0.40 0.51 0.00 1.00 Salinity (ppt) 15 0.30 0.00 0.30 0.30 TDS (ppm) 15 336.13 6.05 321.00 347.00 Velocity (cm/s) 15 17.47 9.58 3.00 27.00 Water Depth (cm) 15 2.47 1.52 1.00 4.50 Water Temperature (°C) 15 29.45 2.34 27.20 35.40 pH 15 7.99 0.14 7.82 8.17

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Reach 17 (Pool) N Mean SD Minimum Maximum Conductivity (mS) 26 731.54 23.20 707.00 777.00 DO (mg/l) 26 4.70 1.72 3.68 8.20 Algae Density (%) 26 57.69 34.56 0.00 80.00 Vegetation Density (%) 26 0.00 0.00 0.00 0.00 Salinity (ppt) 26 0.30 0.00 0.30 0.30 TDS (ppm) 26 329.92 11.01 316.00 356.00 Velocity (cm/s) 26 7.65 3.61 1.00 10.00 Water Depth (cm) 26 3.80 0.57 2.50 4.25 Water Temperature (°C) 26 29.90 2.24 27.00 34.20 pH 26 8.18 0.04 8.11 8.25

Reach 18 (Riffle) N Mean SD Minimum Maximum Conductivity (mS) 29 759.76 11.54 715.00 767.00 DO (mg/l) 29 7.49 0.35 6.96 8.60 Algae Density (%) 29 26.90 8.06 0.00 30.00 Vegetation Density (%) 29 0.03 0.19 0.00 1.00 Salinity (ppt) 29 0.30 0.00 0.30 0.30 TDS (ppm) 29 342.31 3.21 338.00 357.00 Velocity (cm/s) 28 10.79 2.57 5.00 17.00 Water Depth (cm) 29 0.76 0.10 0.50 1.00 Water Temperature (°C) 29 9.64 12.47 3.40 34.70 pH 29 8.13 0.10 7.95 8.42

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Reach 19 (Pool) N Mean SD Minimum Maximum Conductivity (mS) 2293 720.31 39.07 623.00 786.00 DO (mg/l) 2293 7.42 1.86 3.26 12.95 Algae Density (%) 2293 57.89 37.60 0.00 100.00 Vegetation Density (%) 2293 2.67 7.65 0.00 60.00 Salinity (ppt) 2293 0.30 0.02 0.30 0.40 TDS (ppm) 2293 339.08 10.65 310.00 358.00 Velocity (cm/s) 2197 1.29 2.95 0.00 30.00 Water Depth (cm) 2293 13.09 8.10 1.50 47.00 Water Temperature (°C) 2293 30.37 15.32 23.20 299.00 pH 2293 8.24 0.17 7.91 8.72

Reach 20 (Riffle) N Mean SD Minimum Maximum Conductivity (mS) 4 747.75 43.29 683.00 774.00 DO (mg/l) 4 8.52 0.90 7.31 9.50 Algae Density (%) 4 3.75 7.50 0.00 15.00 Vegetation Density (%) 4 7.50 2.89 5.00 10.00 Salinity (ppt) 4 0.30 0.00 0.30 0.30 TDS (ppm) 4 319.50 9.95 311.00 330.00 Velocity (cm/s) 4 28.75 9.60 22.00 43.00 Water Depth (cm) 4 0.69 0.13 0.50 0.75 Water Temperature (°C) 4 32.70 3.99 26.80 35.60 pH 4 8.20 0.09 8.06 8.31

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Reach 21 (Run) N Mean SD Minimum Maximum Conductivity (mS) 63 715.49 48.28 599.00 773.00 DO (mg/l) 63 7.16 1.52 3.54 9.18 Algae Density (%) 63 11.59 15.24 0.00 50.00 Vegetation Density (%) 63 14.54 16.91 1.00 60.00 Salinity (ppt) 63 0.30 0.00 0.30 0.30 TDS (ppm) 63 332.27 10.26 310.00 346.00 Velocity (cm/s) 63 10.36 4.80 4.00 24.00 Water Depth (cm) 63 1.98 1.47 1.25 5.75 Water Temperature (°C) 63 29.61 3.74 24.60 34.30 pH 63 8.21 0.13 7.92 8.34

Reach 22 (Run) N Mean SD Minimum Maximum Conductivity (mS) 186 711.10 66.50 498.00 783.00 DO (mg/l) 186 7.07 1.15 3.70 9.66 Algae Density (%) 186 2.92 9.81 0.00 90.00 Vegetation Density (%) 186 1.55 3.24 0.00 15.00 Salinity (ppt) 186 0.30 0.02 0.20 0.30 TDS (ppm) 186 330.58 14.19 304.00 351.00 Velocity (cm/s) 177 7.28 7.34 1.00 60.00 Water Depth (cm) 186 1.52 0.61 0.75 3.50 Water Temperature (°C) 186 31.05 3.79 23.60 35.80 pH 186 8.34 0.17 8.06 8.72

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Reach 23 (Riffle) N Mean SD Minimum Maximum Conductivity (mS) 2 607.00 0.00 607.00 607.00 DO (mg/l) 2 7.79 0.00 7.79 7.79 Algae Density (%) 2 0.00 0.00 0.00 0.00 Vegetation Density (%) 2 0.00 0.00 0.00 0.00 Salinity (ppt) 2 0.30 0.00 0.30 0.30 TDS (ppm) 2 343.00 0.00 343.00 343.00 Velocity (cm/s) 2 43.89 0.00 43.89 43.89 Water Depth (cm) 2 1.25 0.00 1.25 1.25 Water Temperature (°C) 2 23.50 0.00 23.50 23.50 pH 2 8.33 0.17 8.05 8.72

Reach 24 (Pool) N Mean SD Minimum Maximum Conductivity (mS) 55 722.27 28.90 665.00 742.00 DO (mg/l) 55 7.15 0.55 5.66 7.46 Algae Density (%) 55 22.18 40.31 0.00 100.00 Vegetation Density (%) 55 2.05 3.51 0.00 10.00 Salinity (ppt) 55 0.30 0.00 0.30 0.30 TDS (ppm) 55 314.60 20.17 301.00 349.00 Velocity (cm/s) 44 3.27 1.02 2.00 6.00 Water Depth (cm) 55 3.79 1.02 2.00 4.50 Water Temperature (°C) 55 31.51 3.49 25.40 34.00 pH 55 8.43 0.18 8.15 8.75

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Reach 25 (Riffle) N Mean SD Minimum Maximum Conductivity (mS) 47 704.11 57.81 560.00 773.00 DO (mg/l) 47 7.56 1.31 4.13 9.84 Algae Density (%) 47 3.85 3.47 0.00 10.00 Vegetation Density (%) 47 5.51 7.14 0.00 45.00 Salinity (ppt) 47 0.30 0.00 0.30 0.30 TDS (ppm) 47 330.19 11.38 310.00 349.00 Velocity (cm/s) 47 17.74 6.78 7.00 33.00 Water Depth (cm) 47 1.44 0.47 0.50 2.25 Water Temperature (°C) 47 29.79 4.60 23.10 35.70 pH 47 8.34 0.18 8.06 8.77

Reach 26 (Pool) N Mean SD Minimum Maximum Conductivity (mS) 67 697.19 42.32 622.00 746.00 DO (mg/l) 67 7.79 1.41 4.74 9.15 Algae Density (%) 67 38.88 35.59 0.00 90.00 Vegetation Density (%) 67 7.48 5.71 1.00 15.00 Salinity (ppt) 67 0.31 0.03 0.30 0.40 TDS (ppm) 67 340.85 10.46 328.00 365.00 Velocity (cm/s) 67 5.15 4.30 0.91 16.00 Water Depth (cm) 67 3.40 1.47 0.75 6.75 Water Temperature (°C) 67 28.18 4.12 22.80 33.00 pH 67 8.29 0.08 8.18 8.53

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Reach 27 (Run) N Mean SD Minimum Maximum Conductivity (mS) 8 688.75 55.80 602.00 748.00 DO (mg/l) 8 8.14 0.38 7.66 8.56 Algae Density (%) 8 1.25 2.31 0.00 5.00 Vegetation Density (%) 8 6.63 17.53 0.00 50.00 Salinity (ppt) 8 0.30 0.00 0.30 0.30 TDS (ppm) 8 335.25 4.33 331.00 344.00 Velocity (cm/s) 8 17.77 2.33 16.00 23.00 Water Depth (cm) 8 1.09 0.30 0.50 1.25 Water Temperature (°C) 8 29.43 4.60 22.70 34.40 pH 8 8.36 0.14 8.20 8.75

Reach 28 (Pool) N Mean SD Minimum Maximum Conductivity (mS) 64 573.84 90.27 506.00 732.00 DO (mg/l) 64 7.26 0.52 5.70 8.30 Algae Density (%) 64 29.22 14.89 0.00 50.00 Vegetation Density (%) 64 0.36 1.09 0.00 5.00 Salinity (ppt) 64 0.30 0.00 0.30 0.30 TDS (ppm) 64 334.36 5.36 324.00 347.00 Velocity (cm/s) 51 3.04 3.68 1.00 20.00 Water Depth (cm) 64 7.75 2.91 1.50 12.00 Water Temperature (°C) 64 22.53 4.60 19.30 32.60 pH 64 8.36 0.18 8.05 8.70

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Reach 29 (Pool) N Mean SD Minimum Maximum Conductivity (mS) 501 689.47 58.58 547.00 769.00 DO (mg/l) 501 7.56 1.03 5.16 10.36 Algae Density (%) 501 13.24 22.18 0.00 95.00 Vegetation Density (%) 501 1.04 2.92 0.00 11.00 Salinity (ppt) 501 0.29 0.02 0.20 0.30 TDS (ppm) 501 333.58 7.59 315.00 347.00 Velocity (cm/s) 501 1.01 0.86 0.00 4.00 Water Depth (cm) 501 10.59 5.31 0.89 27.00 Water Temperature (°C) 501 28.83 4.88 22.00 35.80 pH 501 8.35 0.13 8.03 8.57

Reach 30 (Riffle) N Mean SD Minimum Maximum Conductivity (mS) 18 684.33 77.22 420.00 738.00 DO (mg/l) 18 7.81 0.97 4.46 8.67 Algae Density (%) 18 7.28 9.54 0.00 30.00 Vegetation Density (%) 18 44.44 28.23 0.00 85.00 Salinity (ppt) 18 0.30 0.00 0.30 0.30 TDS (ppm) 18 327.11 15.96 296.00 345.00 Velocity (cm/s) 16 12.26 7.55 4.27 35.00 Water Depth (cm) 18 1.83 0.48 1.08 2.75 Water Temperature (°C) 18 29.41 4.88 19.30 33.90 pH 18 8.36 0.15 8.08 8.86

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Texas Tech University, Darrick S. Weissenfluh, December 2010

Reach 31 (Run) N Mean SD Minimum Maximum Conductivity (mS) 112 678.29 50.67 564.00 751.00 DO (mg/l) 112 7.45 0.97 4.48 8.71 Algae Density (%) 112 9.82 13.96 0.00 40.00 Vegetation Density (%) 112 51.25 28.97 5.00 100.00 Salinity (ppt) 112 0.30 0.00 0.30 0.30 TDS (ppm) 112 323.76 18.22 248.00 351.00 Velocity (cm/s) 101 11.19 4.73 1.00 23.77 Water Depth (cm) 112 2.21 0.91 1.00 7.00 Water Temperature (°C) 112 28.13 5.34 8.30 35.50 pH 112 8.39 0.18 7.98 8.92

Reach 32 (Wash) N Mean SD Minimum Maximum Conductivity (mS) 156 675.54 78.44 471.00 957.00 DO (mg/l) 156 6.47 1.47 3.79 9.10 Algae Density (%) 156 6.29 13.02 0.00 60.00 Vegetation Density (%) 156 71.77 22.51 1.00 100.00 Salinity (ppt) 156 0.30 0.04 0.20 0.50 TDS (ppm) 156 327.71 35.81 3.12 458.00 Velocity (cm/s) 150 3.01 4.09 0.00 21.00 Water Depth (cm) 156 1.55 0.74 0.50 6.00 Water Temperature (°C) 156 27.32 5.15 15.20 34.80 pH 156 8.31 0.29 7.19 9.13

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