Consequences of Low-head Dams on Crayfish Distribution and Gene Flow in Alabama Streams by Emily Elizabeth Hartfield A thesis submitted to the Graduate Faculty of Auburn University in partial fulfillment of the requirements for the Degree of Master of Science Auburn, Alabama December 13, 2010 Keywords: crayfish, low-head dams, gene flow, stream ecology, population isolation Copyright 2010 by Emily Elizabeth Hartfield Approved by Jack W. Feminella, Chair, Professor of Biology Guenter A. Schuster, Professor of Biology Scott R. Santos, Associate Professor of Biology Mike M. Gangloff, Assistant Professor of Biology ii Abstract Dams are numerous in many eastern US streams, and instream physicochemical and biotic impacts of dams can alter flow and sediment regimes and channel geomorphology as well as reducing longitudinal movement by fishes and other mobile organisms. In addition, dams can fragment populations, thus decreasing genetic diversity while increasing extinction vulnerability. I quantified freshwater crayfish abundance and their habitats at reaches upstream (1000-5000 m), immediately downstream (mill reaches), and >500 m downstream of 22 low- head milldams within 9 Alabama drainages in 2006?2008. Eleven dams were intact, 5 were partially breached, and 6 were considered relict with more natural flow regimes. On streams with intact dams, crayfish abundance was lower at mill reaches than at reaches upstream of impoundments or further downstream, whereas on streams with breached dams abundance was higher at upstream reaches than at mill or downstream reaches. In contrast, longitudinal patterns in crayfish abundance were similar among sites on streams with relict dams. Predatory fish abundance was higher at mill reaches on streams with intact dams than at sites upstream or further downstream, suggesting that predatory fish aggregations at dam reaches were responsible for low crayfish abundance. Genetic diversity and population connectivity of 2 crayfish species (Cambarus striatus and C. coosae) also was quantified from upstream, mill, and downstream iii reaches of 2 focal, intact dam sites by sequencing a fragment of the mitochondrial cytochrome oxidase I (COI) gene. Cambarus striatus in Sandy Creek showed evidence of upstream population isolation with movement limited to downstream migration across the dam, whereas C. coosae in Hatchett Creek showed no evidence of population structure. Our results suggest that small low-head dams and their reservoirs can alter abundance and impede longitudinal migration of some freshwater crayfishes. iv Acknowledgments I would like to thank my committee, Jack W. Feminella, Guenter A. Schuster, Scott R. Santos, and Mike M. Gangloff, for encouragement and advice. I would also like to thank Susan Balenger, Steven Bryant, Steve Butler, Leslie de Souza, Paul Freeman, Lee Hamm, Shanna Hanes, Nathan Kirk, Ely Kosnicki, Nathan Lujan, Quint Lupton, Tyler Mosely, Nick Ozburn, Guenter Schuster, Erin Singer, Geoff Sorell, Hilary Strickland, Anna Thomas, Keith Ray, Nobuo Ueda, and David Werneke for field and laboratory assistance. Also, thanks to all of the private landowners for allowing us access to the their property and providing information about the history of their land. This project was funded in part by a State Wildlife Grant from the Alabama Department of Conservation and Natural Resources, and a research grant-in-aid from the Auburn University Graduate School. v Table of Contents Abstract ........................................................................................................................................ ii Acknowledgments ......................................................................................................... iv List of Tables .............................................................................................................................. vi List of Figures ................................................................................................................vii I. Chapter 1. Introduction and Literature Review ................................................. 1 References .................................................................................................................. 11 I. Chapter 2. Consequences of Low-head Dams on Crayfish Distribution and Gen Flow in Alabama Streams ..............................................................................22 Introduction ................................................................................................22 Materials and Methods ...............................................................................25 Results ........................................................................................................36 Discussion ..................................................................................................40 Results ........................................................................................................49 Appendix A ............................................................................................................60 vi List of Tables Chapter 2. Consequences of Low-head Dams on Crayfish Distribution and Gene Flow in Alabama Streams Table 1. Study streams, drainages, mill dam names, their conditions (see text), geographic locations, and key dates of construction and/or breach where known ........................................................................................................................... 28 Table 2. Summary of physicochemical and crayfish data from the study streams ...................................................................................................................................... 37 Table 3. Genetic diversity and differentiation among subpopulations of the crayfish Cambarus striatus and C. coosae ..................................................................... 41 vii List of Figures Chapter 2. Consequences of Low-head Dams on Crayfish Distribution and Gene Flow in Alabama Streams Figure 1. Map of Study Sites across Alabama ............................................................. 26 Figure 2. Examples of dams showing different conditions and layout of study reaches for streams with intact, breached, and relict dams ................................. 29 Figure 3. Graphs showing catch per trap effort of crayfish on streams with different dam conditions across all study reaches, catch per trap effort of crayfish at each study reach on streams with dams in different conditions and abundance of predatory fishes at each study reach on streams with dams in different condtions ................................................................................................................................... 38 Figure 4. Haplotype networks showing nesting levels used to infer historical processes for Cambarus striatus and C. coosae .......................................................... 42 1 Chapter One: Introduction and Literature Review Dams are numerous and widespread throughout the state of Alabama and elsewhere in the US, with >10,000 dams occurring in Alabama alone (AL Office of Water Resources http://www.adeca.alabama.gov/) and up to 44% of the mainstem Alabama, Coosa, and Tallapoosa rivers being impounded (Irwin et al. 2007). These structures were built for flood control, hydroelectricity, water storage, recreation, and irrigation. However, most dams within the state are not large hydroelectric structures. Of the >2200 Alabama dams included in the US Army Corps of Engineers National Inventory of Dams (NID) 2009 data set, 69% were <7.5 m in height and 29% were between 7.5 and 15.25 m tall. It is important to note that the NID includes only dams of > 2 m height and 50 acre-ft (61700 m3)of storage or > 7.5 m height and 15 acre-ft (18510 m3) of water storage (http://crunch.tec.army.mil/). The effects of large dams on aquatic organisms and their habitats in large rivers have been well documented, whereas little research has been done to assess effects of small, surface-release, or low-head dams on low-order streams (Watters 1996, Dean et al. 2002, Lessard and Hayes 2003). Low-head dams are those with a hydraulic height of <15m and are typically overflow or spill-way structures 2 (Poff and Hart 2002). According to census records, >65,000 low-head dams existed in the Eastern US by 1840, most of which were built for water-powered milling (Walter and Merritts 2008). Physicochemical Impacts Physicochemical impacts of dams can be dramatic. Perhaps the most obvious effects of dams on streams are changes in the hydrologic regime, channel geomorphology, water temperature, and chemistry, both within the impounded footprint and downstream of dams. Dams and dam alteration may alter the magnitude of minimum flow events typically increases and number of maximum flow events decreases (Poff et al. 1997); in addition, the rate of change in flow (flashiness) increases, although the duration and magnitude of these events often decrease (Poff et al. 1997). Timing of seasonal high and low flow events is altered, resulting in more predictable and less variable flow regimes (Magilligan and Nislow 2005, Graf 2006). In turn, lower magnitude of high-flow events downstream reduces nutrient uptake by flood plains through the deposition of silt during floods, thereby reducing flood plain?stream nutrient exchange (Welcomme 1975, Baxter 1977, Kingsford 2000, Junk and Wantzen 2004). Decreased current velocity in the impounded section increases sediment deposition upstream of the dam, which usually causes tailwaters to become sediment-starved and downstream sections to exhibit increased scour, streambed lowering, and bed coarsening (Baxter 1977, Chien 1985, Graf 2005). Moreover, reduced magnitude of high-flow events and decreased deposition often cause 3 tailwaters to become less geomorphically complex, with fewer bars and islands and reduced shallow-water habitat (Poff et al. 1997, Graf 2006). At the watershed level, such dramatic changes may transform slow-flowing marshy streams into fast-flowing gravel-bottomed systems (Walter and Merritts 2008). In addition to streambed alterations, dams also alter natural thermal regimes (Baxter 1977). Reduced current velocity and increased solar inputs within reservoirs can increase surface water temperatures and plankton growth rates (Baxter 1977). Reduction in current velocity in impounded sections also may cause depth stratification, resulting in a colder but oxygen-poor hypolimnetic zone relative to surface waters (Hart et al. 2002). Increased surface water temperature in impounded waters can result in higher densities of primary producers (e.g., algae and cyanobacteria), which, by increased respiration rates, may result in anoxia or hypoxia (Carmago et al. 2005). In addition, most low- head dams are overflow dams and often result in a significant increase in temperature of dam tailwaters (Lessard and Hayes 2003). The resulting low dissolved oxygen levels can, in turn, cause fish kills and release of macronutrients normally bound to bottom sediments (Correll 1998). Such releases may cause reservoirs to act as nutrient sources, thereby creating downstream eutrophication (Wright 1967). Dams may also affect the stream nitrogen cycle. Nitrification has been shown to increase in impounded streams, especially in surface waters, where decreased current velocity and increased temperature may accelerate transformation of NH4+ to NO3 (Polak 2004, Straus et al. 2004). In contrast, in 4 deeper anoxic sections ammonium denitrification and increases in N concentrations also can occur (Allen 1995). In addition, the physical force of water flowing over dams may cause tailwaters to become supersaturated with oxygen and atmospheric N2 in downstream sections (Morris et al. 1968). Following dam construction, increases in N2 concentrations in downstream reaches of >20% have been documented to cause gas bubble disease and mortality in fish (Beiningen and Ebel 1968, Rucker 1972, Baxter 1977). Biological Impacts Altered flow regimes from impoundments have been shown to affect stream animal assemblages (Fraser 1972, Cushman 1985, Irvine 1985, Travinchek et al. 1995, Gerhke et al. 2002, McLaughlin et al. 2006) and even riparian vegetation (Janson et al. 2000). Coarsening of the streambed by erosion of sediment-poor tailwaters reduces habitat availability for benthic species by decreasing habitat heterogeneity, which, in turn, may reduce species diversity and richness (Armitage and Blackburn 1990, Hauer et al. 1989, Poff et al. 1997). Alterations in temperature regimes from impoundments also may alter organism distribution and behavior. Increased temperatures downstream of overflow dams can eliminate thermal cues vital to some invertebrate life cycles (Lehmkuhl 1974, Ward and Stanford 1982, Irvine 1985). In addition, increased water temperature affects metabolic rates for fish and invertebrates, which, in turn, increase demands for food to maintain growth and survival (Gibbons 1976, Wotton 1995, Perry et al. 1987, Vinson 2001, Lessard and Hayes 2003). Within reservoirs the deep, 5 cold, anoxic water often is a fish- and mollusc-free zone (Headrick and Carline 1993, Dean et al. 2002). The few studies that were designed to examine effects of small dams have reported similar alterations but smaller in magnitude than those resulting from large dams (Graf 2006). In particular, Cumming (2004) found that low-head dams increase summer maximum temperatures and decrease fish richness at reaches upstream of impoundments. Taylor et al. (2001) documented shifts in fish composition, from a cyprinid- to a centrarchid-dominated assemblage after impoundment of a stream by a low-head earthen dam. Habitat fragmentation and population isolation Physical barriers presented by dam structures (including the dam itself, the impounded zone, and the affected tailwaters) impede longitudinal movements of stream organisms (Baxter 1977, Watters 1996, Dean et al. 2002). Genetic drift can result after such separations as rare alleles become common or fixed in a population while other alleles become less frequent or disappear (Lande 1976). Natural selection may expedite the divergence between reproductively isolated populations occurring in different habitats (Felsenstein 1976), but isolation has a greater influence on genetic divergence between populations than selection (Dillon 1984, Finlay et al. 2006). The degree of divergence between separated populations can be quantified by analyzing the accumulation of fixed mutations in the genome of each population (Nei 1977). 6 Impeded migration of fishes and other mobile organisms by small dams has been observed, preventing individuals from reaching feeding and/or spawning habitat. Among fish, decreased longitudinal connectivity across streams exacerbates population isolation (Neraas and Spruell 2001, Olden et al. 2001). One-way (downstream) migration of fish, commonly observed in impounded systems, reduces genetic diversity and population size in upstream reaches (Jager et al. 2001, Morita and Yamamoto 2002, Yamamoto et al. 2004). Dams have similar effects on freshwater mussels by restricting migration and distribution of their host fish through impounded sections (Watters 1996, Kelner and Sietman 2000). Similar consequences on other invertebrates also have been observed. Watanabe and Omura (2007) demonstrated greater genetic differentiation among sub-populations of caddisflies separated by large reservoirs than among sub-populations on unimpounded reference streams. It is unknown how low-head dams affect habitat conditions or longitudinal movements of stream crayfishes, although it is likely that migration of some species is impeded (Miya and Hamano 1988). Habitat fragmentation is of great conservation concern because ecological theory predicts that isolated populations can decrease in size and genetic diversity, making them more vulnerable to extinction (MacArthur and Wilson 1967, Lande 1988, Lande 1999). Without sufficient immigration from neighboring populations, natural stochastic events or anthropogenic impacts that reduce population size can lead to loss of genetic diversity and inbreeding depression (Charlesworth and Charlesworth 1987, Crnokrak and Roff 1999), thus reducing 7 population fitness and ultimately causing local extinction (Hansson and Westerberg 2002, Reed and Frankham 2003, Watanabe and Omura 2007). Further, by reducing the likelihood of interaction among populations, fragmentation also reduces the chance of successful recolonization after local extinctions, thus threatening metapopulation persistence (Saunders et al. 1991, Young et al. 1996, Fagan 2002). Crayfish diversity in Alabama Alabama is a freshwater biodiversity ?hotspot,? as it supports 60% of North America?s native mussel species, 43% of native freshwater snails, 38% of native fishes, and 24% of native crayfishes, many of which are endemic to the southeastern US(Lydeard and Mayden 1995, Crandall et al. 2000, Schuster and Taylor 2004). In addition, Alabama?s streams are considered its most imperiled ecosystems due, in part, to flow modification from impoundments (Dudgeon et al. 2005). Freshwater crayfishes are highly diverse within Alabama, with at least 83 species in 6 genera and 25 subgenera occurring in the state (Schuster and Taylor 2004). Diverse life history strategies are represented, including cave and spring dwellers, and primary, secondary, and tertiary burrowers. Primary burrowers dig elaborate burrows in flood plains and moist low-lands where they spend most of their lives. Secondary burrowers dig more simplistic borrows, usually opening to a permanent body of water such as streams, lakes, and sloughs, and may forage in the open water. In contrast, tertiary burrowers live primarily in permanent 8 flowing water and dig burrows only during drought or when females tend eggs (Taylor and Schuster 2004). In addition, some species can facultatively use multiple burrowing strategies (Schuster and Taylor 2004, Finlay et al. 2006). Crayfish are polyphagous, consuming primarily macroalgae, but also feed on detritus and animal prey (Creed 1994, Momot 1995). Predators of crayfish include fish, wading birds, mammals, and larger crayfish (Stein and Magnuson 1976, Englund and Krupa 2000). Male crayfish show cyclical reproduction, molting into the form I (reproductively active) stage at the onset of breeding, and back to form II (reproductively inactive) stage after reproduction (Crandall and Fitzpatrick 1996). Form I and form II males are distinguished by the sclerotized condition of the gonopods, which are use to transfer sperm to females. Females of some species also may exhibit form alteration (Wetzel 2002), and typically carry fertilized eggs and newly hatched young on the venter of the abdomen until their second or third juvenile molt. Breeding for most species occurs in autumn, winter, and early spring (Taylor and Schuster 2004); however, life history and reproductive strategies of many of freshwater crayfish species, including those in Alabama, are unknown. Most of Alabama?s stream crayfish occur in the genera Cambarus, Procambarus, and Orconectes. The genera are readily distinguished by the gonopods of form I males, and most form II males. The form I male gonopods are important in identification of most crayfish species (Hobbs 1981, Crandall and Fitzpatrick 1996, Taylor and Schuster 2004); however, distinguishing species within a genus is considerably more difficult, if only form II or female specimens 9 are available. It is likely that there are undescribed crayfish species in Alabama, and several museum specimens at the Auburn University Museum and the University of Alabama Museum are listed as undescribed (Schuster and Taylor 2004, E. Hartfield, personal observations). Unlike many aquatic animals, crayfish may migrate both through water and, over land. Long-distance dispersal occurs primarily during floods (Lodge et al. 2000). Overland migration is limited by tolerance to desiccation, which varies among species (Larson and Magoulick 2008) and is usually limited to rain events, when relative humidity is high. For this reason, migration mostly occurs within a single drainage, although inter-drainage movements do occur (Fetzner and Crandall 2003). In addition, introduction of some species may occur via bait- bucket transfer by humans (Lodge et al. 2000). Population genetics studies have revealed extensive gene flow and large effective population sizes in several species of subterranean- and surface-dwelling crayfish; however, molecular data also have suggested recent declines in genetic variability and effective population size in some widespread non-burrowing taxa (Buhay and Crandall 2005, Finlay et al. 2006). Such declines are of special concern because species with restricted ranges are highly vulnerable to genetic isolation resulting from habitat fragmentation (Lande 1999). Objectives Currently, there is little information available on historic distribution and population sizes, or ecology of Alabama?s crayfishes, especially endemic species 10 (Butler et al. 2003). Only 5 species of crayfish in Alabama have been listed as High Conservation Concern (Schuster and Taylor 2004). In the southeastern US, including Alabama, the high prevalence of small dams along streams and rivers has the potential to affect many imperiled aquatic species primarily through habitat fragmentation and population isolation (Travnicheck et al. 1995, Jager et al. 2001, Dean et al. 2002, Lessard and Hayes 2003, Irwin et al. 2007). Thus, the general paucity of knowledge of crayfish population size, life history, and ecology in Alabama, coupled with the high prevalence of impoundments, requires elucidation of the effects of impoundments on crayfish species assemblages, and increased ecological data on crayfish population dynamics within the state and elsewhere in the southeastern US. The objectives of my research were to examine if presence and condition of low-head dams affect crayfish assemblages, and whether small low-head dams impede gene flow among crayfish populations. These objectives are important to crayfish conservation because 1) Alabama is a species-rich region with many endemic and undescribed taxa, 2) range, lifehistory, and ecology of many Alabama crayfish are unknown, and 3)patchy and fragmented habitats increase reproductive isolation and may increase likelihood of extinction/extirpation (Taylor et al. 2007). Finally, low-head dams are widespread globally and increasing in abundance (Wu et al. 2003), so an increased understanding of how these structures contribute to population fragmentation may enhance management of imperiled aquatic species across regional and global scales. 11 References Allan, J.D. 1995. Stream ecology: structure and function of running waters. Chapman and Hall, New York. Armitage, P.D., and J.H. Blackburn. 1990. Environmental stability and communities of Chironomidae (Diptera) in a regulated river. Regulated Rivers: Research and Management 5:319-328 Baxter, R.M. 1977. Environmental effects of dams and impoundments. Annual Review of Ecology and Systematics 8:255-283. Beiningen, K.T. and W.J. Ebel. 1968. Effect of John Day Dam on dissolved nitrogen concentrations and salmon in the Columbia River, 1968. Transactions of the American Fisheries Society 99:664-671. Buhay, J.E. and K. A. Crandall. 2005. Subterranean phylogeography of freshwater crayfishes shows extensive gene flow and surprisingly large population sizes. Molecular Ecology 14:4259-4273. Butler, R.S., RJ. DeStefano, and G.A. Schuster. 2003. Crayfish: an overlooked fauna. Endangered Species Bulletin 28:10-12. Carmago, J.A., A. Alonso, M. de la Puente. 2005. Eutrophication downstream from small reservoirs in mountain rivers of Central Spain. Water Research 39:3376-3384. Charlesworth, D. and B. Charlesworth. 1987. Inbreeding depression and its evolutionary consequences. Annual Review of Ecology and Systematics 18:237-268. Chien, N. 1985. Changes in river regime after the construction of upstream 12 reservoirs. Earth Surface Processes and Landforms 10:143-159. Correll, D.L. 1998. The role of phosphorus in eutrophication of receiving waters: a review. Journal of Environmental Quality 27:261-266. Crandall, K.A. and J.F. Fitzpatrick Jr. 1996. Crayfish molecular systematics: using a combination of procedures to estimate phylogeny. Systematic Biology 45:1-26. Crandall, K.A, D.J. Harris, J.W. Fetzner Jr. 2000. The monophyletic origin of freshwater crayfish estimated from nuclear and mitochondrial DNA sequences. Proclamations of the Royal Society of London 267:1679-1686. Creed, R.P. 1994. Direct and indirect effects of crayfish grazing in a stream community. Ecology 75:2091-2103. Crnokrak, P., and D.A. Roff. 1999. Inbreeding depression in the wild. Heredity 83:260-270. Cumming, G.S. 2004. The impact of low-head dams on fish species richness in Wisconsin, USA. Ecological Applications 14:1495-1506. Cushman, R.M. 1985. Review of ecological effects of rapidly varying flows downstream of hydroelectric facilities. North American Journal of Fisheries Management 5:330-339. Dean, J., D. Edds, D. Gillette, J. Howard, S. Sherraden, and J. Tiemann. 2002. Effects of lowhead dams on freshwater mussels in the Neosho River, Kansas. Transactions of the Kansas Academy of Science 105:323-240. Dillon, R. T., Jr. 1984. Geographic distance, environmental difference, and divergence between isolated populations. Systematic Zoology 33:69-82. 13 Dudgeon, D., A.H. Arthington, M.O.Gessner, Z.I. Kawabata, D.J. Knowler, C. Leveque, R.J. Naiman, A.H. Prieur-Richard, D. Soto, M.L.J. Stiassny, and C.A. Sullivan. 2006. Freshwater biodiversity: importance, threats, status and conservation challenges. Biological Review 81:163-182. Englund, G. and J.J. Krupa. 2000. Habitat use by crayfish in stream pools: influence of predators, depth, and body size. Freshwater Biology 43:75- 83. Fagan, W.F. 2002. Connectivity, fragmentation, and extinction risk in dendritic metapopulations. Ecology 83:3243-3249. Felsenstein, J. 1976. The theoretical population genetics of variable selection and migration. Annual Review of Genetics 10:253-280. Fetzner, J. W. and K. A. Crandall. 2003. Linear habitats and the nested clad analyses: an empirical evaluation of geographic versus river distances using an Ozark crayfish (Decapoda: Cambaridae). Evolution 57:2101- 2118. Finlay, J. B., J. E. Buhay, and K. A. Crandall. 2006. Surface to subsurface freshwater connections: phylogeographic and habitat analyses of Cambarus tenebrosus, a facultative cave-dwelling crayfish. Animal Conservation 9:375-387. Fraser, J.C. 1972. Regulated discharge and the stream environment. Pages 263- 286 in R. Olgesby, C.A. Carlson, and J. McCann, editors. River Ecology and Management. Academic Press, New York. Gehrke, P.C., D.M. Gilligan, and M. Barwick. 2002. Changes in fish 14 communities of the Shoalhaven River 20 years after construction of Tallowa Dam, Australia. River Research and Applications 18:265-286. Gibbons, J.W. 1976. Thermal alteration and the enhancement of species populations. In G.W. Esch, R.W. McFarlane, editors. Thermal Ecology II, ERDA Symposium Series. Graf, W.L. 2005. Geomorphology and American dams: the scientific, social, and economic context. Geomorphology 71:3-26. Graf, W.L. 2006. Downstream hydrologic and geomorphic effects of large dams on American rivers. Geomorphology 79:336-360. Hansson, B., and L. Westerberg. 2002. On the correlation between heterzygosity and fitness in natural populations. Molecular Ecology 11:2467-2474. Hart, D.D., T.E. Johnson, K.L. Bushaw-Newton, R.J. Horwitz, A.T. Bednarek, D.F. Charles, D.A. Kreeger, and K.J. Velinsky. 2002. Dam removal: challenges and opportunities for ecological research and river restoration. BioScience 52:669-682. Hauer, F.R., J.A. Stanford, J.V. Ward. 1989. Serial discontinuities in a Rocky Mountain river. II. Distribution and Abundance of Trichoptera. Regulated Rivers: Research and Management 3:177-182. Headrick, M.R. and R.F. Carline. 1993. Restricted summer habitat and growth of northern pike in two southern Ohio impoundments. Transactions of the American Fisheries Society 122:228-236. Hobbs, H.H., Jr. 1981. The crayfishes of Georgia. Smithsonian Contributions to Zoology 318:1-549. 15 Irvine, J.R. 1985. Effects of successive flow perturbations on stream invertebrates. Canadian Journal of Fisheries and Aquatic Sciences 42:1922-1927. Irwin, E., G. Turner, K. Mickett, and T. Piper. 2007. ACT Aquatic Gap and water quality monitory: Final Report. AL Department of Conservation of Natural Resources. Jager, H.I., J.A. Chandler, K.B. Lepla, and W.V. Winkle. 2001. A theoretical study of river fragmentation by dams and its effects on white sturgeon populations. Environmental Biology of Fishes 60:347-361. Jansson, R., C. Nilsson, and B. Renofalt. 2000. Fragmentation of riparian floras in rivers with multiple dams. Ecology 81:899-903. Junk, W.J. and K.M. Wantzen. 2004. The flood pulse concept: new aspects, approaches, and applications ? an update. Pages 117-140 in R.L. Welcomme and T. Petr, editors. Proceedings of the Second International Symposium on the Management of Large Rivers for Fisheries Volume II, FAO, RAP Publication. Kelner, D.E. and B.E. Sietman. 2000. Relic populations of the Ebony Shell, Fusconaia ebena (Bivalvia: Uniondiae), in the Upper Mississippi River Drainage. Journal of Freshwater Ecology 15:371-378. Kingsford, R.T. 2000. Ecological impacts of dams, water diversions and river management on floodplain wetlands in Australia. Austral Ecology 25:109-127. Lande, R. 1976. Natural selection and random genetic drift in phenotypic 16 evolution. Evolution 10:314-334. Lande, R. 1988. Genetics and demography in biological conservation. Science 241:1455-1460. Lande, R. 1999. Extinction risks from anthropogenic, ecological, and genetic factors. Pages 1-22 in L.F. Landweber and A.P. Dobson, editors. Genetics and the extinction of species. Princeton University Press, Princeton, New Jersey. Larson, E.R. and D.D. Magoulick. 2008. Comparative life history of native (Orconectes eupunctus) and introduced (Orconectes neglectus) crayfishes in the Spring River Drainage of Arkansas and Missouri. American Midland Naturalist 160:323-341. Lehmkuhl, D.M. 1974. Thermal regime alterations and vital environmental physiological signals in aquatic systems. Pages 216-222 in J.W. Gibbons, and R.R. Sharitz, editors. Thermal ecology, AEC Symposium Series. Lessard, J.L. and D.B. Hayes. 2003. Effects of elevated water temperature on fish and macroinvertebrate communities below small dams. River Research and Applications 19:721-732. Lodge, D. M., C. A. Taylor, D. M. Holdich, and J. Skurdal. 2000. Nonindigenous crayfishes threaten North American freshwater biodiversity. Fisheries 25(8):7-20. Lydeard, C. and R.L. Mayden. 1995. A diverse and endangered aquatic ecosystem of the Southeast United States. Conservation Biology 9:800- 805. 17 MacArthur, O.H. and E.O. Wilson. 1967. The theory of island biogeography. Princeton University Press, Princeton, New Jersey. Magilligan, F.J., and K.H. Nislow. 2005. Changes in hydrologic regime by dams. Geomorphology 71:61-78. McLaughlin, R.L., L. Porto, D.L.G. Noakes, J.R. Baylis, L.M. Carl, H.R. Dodd, J.D.Goldstein, D.B. Hayes, and R.G. Randall. 2006. Effects of low-head barriers on stream fishes: taxonomic affiliations and morphological correlates of sensitive species. Canadian Journal of Fisheries and Aquatic Sciences 63:766-779. Miya, Y. and T. Hamano. 1988. The influence of a dam having no fish-way on the distribution of decapod crustaceans in the Yukinoura River, Nagasaki, Japan. Nippon Suisan Gakkaishi 54:429-435. Momot, W.T. 1995. Redefining the role of crayfish in aquatic ecosystems. Reviews in Fisheries Science 3:33-63. Morita, K. and S. Yamamoto. 2002. Effects of habitat fragmentation by damming on the persistence of stream-dwelling charr populations. Conservation Biology 16:1318-1323. Morris, L.A., R.N. Langemeier, T.R. Russell, and A. Witt, Jr. 1968. Effects of main stem impoundments and channelization upon limnology of the Missouri River, Nebraska. Transactions of the American Fisheries Society 97:380-388. Nei, M. 1977. F-statistics and analysis of gene diversity in subdivided populations. Annals of Human Genetics 41:225-233. 18 Neraas, L.P. and P. Spruell. 2001. Fragmentation of riverine systems: the genetic effects of dams on bull trout (Salvelinus confluentus) in the Clark Fork River system. Molecular Ecology 10:1153-1164. Olden, J.D., D.A. Jackson, and P.R. Peres-Neto. 2001. Spatial isolation and fish communities in drainage lakes. Oecologia 127:572-585. Perry, S.A., W.B. Perry, J.A. Stanford. 1987. Effects of thermal regime on size, growth rates and emergence of two species of stoneflies (Plecoptera: Taeniopterygidae, Pteronarcyidae) in the Flathead River, Montana. American Midland Naturalist 117:83-93. Poff, N.L., J.D. Allan, M.B. Bain, J.R. Karr, K.L. Prestegaard, B.D. Richter, R.E. Sparks, and J.C. Stromberg. 1997. The natural flow regime: a paradigm for river conservation and restoration. BioScience 47:769-784. Poff, N.L., and D.D. Hart. 2002. How dams vary and why it matters for the emerging science of dam removal. BioScience 52:659-668. Polak, J. 2004. Nitrification in the surface water of the Wloclawek Dam Reservoir: the process contribution to biochemical oxygen demand (N- BOD). Polish Journal of Environmental Studies 13:415-424. Reed, D.H., and R. Frankham. 2003. Correlation between fitness and genetic diversity. Conservation Biology 17:230-237. Rucker, R.R. 1972. Gas bubble disease of salmonids: a critical review. U.S. Bureau of Sport Fisheries and Wildlife Technical Paper Number 58. Saunders, D.A., R.J. Hobbs, and C.R. Margules. 1991. Biological consequences of ecosystem fragmentation: a review. Conservation Biology 5:18-32. 19 Schuster, G.A., and C.A. Taylor. 2004. Report of the crayfishes of Alabama: literature and museum database review, species list with abbreviated annotations and proposed conservation statuses. Illinois Natural History Survey, Center for Biodiversity Technical Report. Stein, R. A. and J.J. Magnuson. 1976. Behavioral response of crayfish to a fish predator. Ecology 57:751-761. Strauss, E.A., W.B. Richardson, L.A. Bartsch, J.C. Cavanaugh, D.A. Bruesewitz, H. Imker, J.A. Heinz, and D.M. Soballe. 2004. Nitrification in the Upper Mississippi River: patterns, controls, and contribution to the NO3- budget. Journal of the North American Benthological Society 23:1-14. Taylor, C.A., J.H. Knouft, and T.M. Hilland. 2001. Consequences of stream impoundment on fish communities in a small North American drainage. Regulated Rivers: Research and Management 17:687-698. Taylor, C.A. and G.A. Schuster. 2004. The crayfishes of Kentucky. Illinois Natural History Survey. Special Publication No. 28:1-219. Taylor, C.A., G.A. Schuster, J.E Cooper, R.J. DiStefano, A.G. Eversole, R. Hamr, H.H. Hobbs III, H.W. Robinson, C.E. Skelton, R.G. Thoma. 2007. A reassessment of the conservation status of crayfishes of the United State and Canada after 10+ years of increased awareness. Fisheries 32(8):372- 389. Travnicheck, V.H., M.B. Bain, M.J. Maceina. 1995. Recovery of a warmwater 20 fish assemblage after the initiation of minimum-flow release downstream from a hydroelectric dam. Transactions of the American Fisheries Society 124:836-844. U.S. Department of State, Compendium of the Enumeration of the Inhabitants and Statistics of the United States as Obtained at the Department of State, from the Returns of the Sixth Census (Thomas Allen, Washington, DC, 1841). Vinson, M.R. 2001. Long-term dynamics of an invertebrate assemblage downstream from a large dam. Ecological Applications 11:711-730. Walter, R.C., and D.J. Merritts. 2008. Natural streams and the legacy of water powered mills. Science 319:299-304. Ward, J.V. and J.A. Stanford. 1982. Thermal responses in the evolutionary ecology of aquatic insects. Annual Review of Entomology 27:97-117. Watanabe, K. and T. Omura. 2007. Relationship between reservoir size and genetic differentiation of the stream caddisfly Stenopsyche marmorata. Biological Conservation 136:203-211. Watters, G.T. 1996. Small dams as barriers to freshwater mussels (Molusca: Pelecypoda: Unionidae) and their hosts. Biological Conservation 75:79- 85. Welcomme, R.L. 1975. The fisheries ecology of African floodplains. CIFA Technical Paper Number 3. Rome: FAO. Wetzel, J.E. 2002. Form alteration of adult female crayfishes of the genus 21 Orconectes (Decapoda: Cambaridae). American Midland Naturalist 147:326-337. Wotton, R.S. 1995. Temperature and lake-outlet communities. Journal of Thermal Biology 20:121-125. Wright, J.C. 1967. Effect of impoundments of productivity, water chemistry, and heat budgets of rivers. Reservoir fisheries resources. American Fisheries Society, Washington, D.C. Wu, J., J. Huang, X. Han, Z. Xie, and X. Gao. 2003. Three-Gorges Dam- experiment in habitat fragmentation? Science 300:1239-1240. Yamamoto, S. K. Morita, I. Koizumi, and K. Maekawa. 2004. Genetic differentiation of white-spotted charr (Salvelinus leucomaenis) populations after habitat fragmentation: spatial-temporal changes in gene frequencies. Conservation Genetics 5:529-538. Young, A., T. Boyle, and T. Brown. 1996. The population genetic consequences of habitat fragmentation for plants. Trends in Ecology and Evolution 11:413-418. 22 Chapter Two: Consequences of Low-head Dams on Crayfish Distributions and Gene Flow in Alabama Streams Introduction Fragmented populations are more vulnerable to local extinction than contiguous populations (MacArthur and Wilson 1967, Lande 1988, Frankham 1997, Lande 1999). Without sufficient immigration from neighboring populations, natural stochastic events and/or anthropogenic impacts that reduce population size can lead to loss of genetic diversity and inbreeding depression (Charlesworth and Charlesworth 1987, Crnokrak and Roff 1999), thus reducing population fitness and ultimately causing local extinction (Hansson and Westerberg 2002, Reed and Frankham 2003, Watanabe and Omura 2007). Following extirpation, fragmentation reduces the likelihood of successful recolonization of patches, which may reduce the size of and exchange within metapopulations and further threatening persistence (Saunders et al. 1991, Young et al. 1996, Fagan 2002). Streams are special cases when considering habitat fragmentation because of their linear structure. Many lotic organisms are restricted to the wetted channel, so migration can only occur bidirectionally (i.e., upstream or downstream; Fagan 2002, Hughes et al. 2009). Moreover, stream habitats often are naturally patchy (Townsend 1989). Terrestrial dispersal may alleviate instream constraints for some aquatic species, but even these organisms are 23 restricted to or near stream corridors (Finn et al. 2007). Thus, a single barrier bisecting a stream corridor may have dire consequences by restricting or eliminating connectivity for populations on either side. Dams are extreme examples of anthropogenic barriers that can fragment stream populations, and effects of large dams on lotic biota and their habitats in large rivers can be profound. Environmental consequences of these structures include severe alteration of assemblage structure of both animals (Fraser 1972, Cushman 1985, Irvine 1985, Travinchek et al. 1995, Gerhke et al. 2002, McLaughlin et al. 2006) and vegetation (Janson et al. 2000). In addition, increased temperatures in tailwaters of overflow dams can increase energetic demands and physiological stresses for downstream fauna (Gibbons 1976, Wotton 1995, Perry et al. 1987, Vinson 2001, Lessard and Hayes 2003). Interruptions in longitudinal dispersal by large dams have been well- documented for several species. Dams may halt upstream migration of fishes and other mobile animals, preventing individuals from reaching feeding and/or spawning habitat, which, in turn, drive population declines (Raymond 1979, Larinier 2001, Neraas and Spruell 2001, Olden et al. 2001). The one-way (downstream) migration of fish, commonly observed in impounded systems, reduces genetic diversity and population size, particularly in upstream reaches (Jager et al. 2001, Morita and Yamamoto 2002, Yamamoto et al. 2004). Dams can have similar impacts on freshwater mussels by restricting migration and distribution of their host fish through impounded sections and thus breaking mussel life cycles (Watters 1996, Kelner and Sietman 2000). 24 Longitudinal movement by crayfishes is common within streams (Fetzner and Crandall 2003, Lodge et al. 2000; however, unlike many lotic species, crayfishes are not restricted to the wetted channel. Overland dispersal is possible for some species (Viosca 1953, Penn 1956, Cappelli and Magnuson 1983), particularly those that are desiccation-tolerant (Larson et al. 2009). Given that crayfish dispersal ability varies among species, it is difficult to predict the degree to which instream barriers, such as dams, affect population connectivity. Furthermore, the systematics, life history and ecology are poorly known for most crayfish species in North America, and a number of species remain undescribed (Taylor et al. 2007). Understanding impacts of dams crayfishes is critical for their conservation. Consequences of dams on aquatic organisms and their habitats in large rivers have been well-documented, whereas comparatively little work has been done to assess effects of small, surface-release, and/or low-head dams on smaller streams (Watters 1996, Dean et al. 2002, Lessard and Hayes 2003). Low-head dams are those with a hydraulic height of <8 m that typically have over-dam flow or lateral spillways (IFC Consulting Report 2005). Such structures are pervasive across the Eastern US; according to census records, >65,000 low-head dams existed in the region by 1840, most of which were built for water-powered milling (Walter and Merritts 2008). The few studies designed to examine effects of small dams have reported similar types of alterations as with large dams, although effects are considerably smaller in magnitude (Graf 2006). Low-head dams are increasing in abundance globally (Wu et al. 2003), so it is essential to understand 25 the degree to which these structures contribute to population fragmentation and reproductive isolation of aquatic species. This study was designed to examine the degree to which variation in environmental conditions associated with low-head dams affected the abundance, distribution, and gene flow of crayfish populations across a range of dam conditions within impounded or historically impounded streams in Alabama. Materials and Methods Study sites I studied 22 low-head dams in 9 of the major drainages within Alabama (Fig. 1). Study sites spanned the full range of physiographic provinces with the state, including Highland Rim, Cumberland Plateau, Alabama Ridge and Valley, and Eastern Gulf Coastal Plain (Hackney et al. 1993). Study streams ranged from 3rd to 5th-order, and sites drained predominantly forested watersheds, although a mosaic of agricultural, suburban and urban land uses also was present. Because of the high variability across provinces and drainages, habitat types and conditions and associated animal assemblages greatly varied among sites. Dams are abundant throughout Alabama, and many were discovered during previous surveys and from discussions with local residents (Gangloff and Feminella 2007). Sites were selected based on the 3 dam condition categories (below), landowner approval, and accessibility of the dam and upstream and further downstream reaches. Also, sites were selected due to presence of endangered or threatened aquatic taxa (i.e., mussels, snails, fish, or crayfish) 26 Figure 1. Map of study sites across Alabama. Open circles represent intact dams, open squares represent breached dams, and black hexagons represent relict dams. 27 nearby. Of the selected sites, 11 dams were intact, 5 were breached, and 6 were considered relict. Intact dams were those where the impoundment structure had a functional spillway, over-dam flow, and a reservoir with slow current velocity relative to nearby free-flowing reaches (Fig. 2A). Breached dams were those where the impoundment structure was either partially broken or the spillway was open and the impoundment zone was absent or <50m long at base flow (Fig. 2B). Last, relict dams were those where impoundment structures were almost entirely eroded (e.g., often showing only pilings on the bank to indicate original dam location, Fig. 2C), usually from advanced age and the action of hydrologic forces during flood events, thus allowing the return of more typical free-flowing conditions. Study design I established three 150-m study reaches at each study dam: 1) upstream of impounded reaches (= upstream, 1500-5000 m), 2) immediately downstream of the dam (= mill; 0-150 m), and 3) 500 to 3000 m downstream of the dam (= downstream; usually <1000 m) (Fig. 2D). I estimated longitudinal length of pre- breached impoundment for breached and relict dams using information from local landowners or long-term residents. Thus, I established this reach an equivalent distance upstream of dams in all conditions (Fig. 2D). 28 Table 1. Study streams, drainages, milldam names and their condition (see text), geographic location, and key dates of construction and/or breach, where known. Stream Drainage Mill Cond- ition Lat. (N) Long. (W) Dates Big Flat Alabama Rikard's Mill Intact 31.782 -87.223 Built 1868 Cahaba Alabama Grant's Mill Relict 33.509 -86.644 Little Cahaba Cahaba Unnamed Mill Breached 33.451 -86.694 Lost Black Warrior Boshell Mill Intact 33.855 -87.414 Built 1885 Brushy Black Warrior Unnamed Mill Intact 34.292 -87.273 Built 1966 Blue Springs Black Warrior Chamblee Mill Relict 34.060 -86.662 Halawakee Chatahoo-chee Bean's Mill Intact 32.697 -85.267 Built 1834 Osanippa Chatahoo-chee Ferguson Mill Relict 32.778 -85.193 Little Uchee Chatahoo- chee Meadow Mill Intact 32.528 -85.253 Pea Choctaw-hatchee Shellgrove Mill Relict 31.521 -85.869 Built 1890?s Big Canoe Coosa Goodwin Mill Breached 33.819 -86.384 Breached 1990?s Yellow Leaf Coosa Shannon Mill Intact 32.935 -86.611 Hatchett Coosa Old AL Power Mill Intact 33.068 -86.096 Built 1920?s Chocta- faula Tallapoosa Vaughn's Mill Breached 32.513 -85.578 Built 1940 Breached 1990?s Loblockee Tallapoosa Macon's Mill Intact 33.653 -85.584 Sandy Tallapoosa Jones? Mill Intact 32.751 -85.560 Built 1830?s Little Hilabee Tallapoosa Carr?s Mill Relict 33.204 -85.943 Breached 1940?s Paint Rock Tennessee Butler?s Mill Relict 34.579 -86.301 Built 1820?s Turkey Tennessee Masterson Mill Intact 34.538 -87.283 Built 1870?s Butta- hatchee Tombigbee Unnamed Mill Breached 34.126 -87.837 Built 1920?s Breached 1980?s New Tombigbee Kelly's Mill Intact 33.930 -87.680 Pearce's Mill Tombigbee Pearce's Mill Breached 34.122 -87.836 Built 1920?s Breached 1980?s 29 Figure 2. Examples of dams showing intact (a), breached (b), and relict (c) conditions, and layout of study reaches (d) for streams with intact (left) and breached or relict (right). Square symbols represent upstream sites, stars represent mill sites, and circles represent downstream sites. 30 I measured a full range of physicochemical variables (mean current velocity, stream depth, substrate size, and channel width) thought to vary among sites concurrently with biotic sampling over Spring-Summer and Fall 2007 and 2008. I established cross-stream transects every 10 m in each reach and measured wetted width, and current velocity (measured at 0.6x the channel depth, Gore 2006), depth (n=5/transect), substrate size (pebble counts method [Wolman 1954], n=20/transect at random intervals), and proportion of unmeasured substrate in the channel (wood, bedrock, organic matter, sand/silt). Streamwater chemistry data included conductivity (Sharp C66 meter), pH (Sharp pH52), and dissolved oxygen (YSI 55), measured from grab samples from each reach and in the impoundment zone or its equivalent during summer base flow (July 2008). Conditions at all reaches on a stream were measured within 1 or 2 d of each other to ensure similar physical conditions among reaches. I measured water temperature continuously (3-h intervals) using iButton data loggers deployed at downstream, mill, and upstream reaches; I deployed a 4th logger in the impoundment zone of intact dams, or in a reach of equivalent distance upstream from the dam for breached and relict sites. Crayfish and fish sampling I quantified crayfishes in each reach using a combination of trapping, seining, and electrofishing, thus minimizing sampling bias of any single method (Rabeni et al.1997, Ratcliffe and DeVries 2004). I set 8 crayfish traps, baited with perforated cans of cat food, over night once in each reach during May- 31 September. All reaches on a stream were typically sampled the same night or within a day of each other to ensure similar conditions upstream to downstream. I used kick-seining (mesh size=0.25mm) and a backpack electrofishing unit (Smith-Root LR-24 electrofishing unit) to quantify fishes, potential biotic controls on crayfishes, in the full range of habitat unit types available in study reaches (i.e., riffles, runs, pools, backwaters). Crayfish specimens were kept alive on ice and later preserved in 95% ethanol in the laboratory where they were identified to species using keys in Hobbs (1981, 1989) and Taylor and Schuster (2004). I anesthetized fishes in tricane methanesulfonate (MS-222), fixed them in a 10% formalin solution, and then transferred them to 70% ethanol for permanent storage. In the laboratory, I identified fishes to species and classified them according to feeding guilds (Berkman and Rabeni 1987, Boschung and Mayden 2004, B. Helms unpublished data). All specimens were deposited in the Auburn University Learning Center. Statistical analysis I used a General Linear Model (PROC GLM, SAS 9.1) with reach (upstream, mill, downstream) as a fixed factor and site (intact, breached, relict dam condition) as a random factor with an interaction term to test for differences in substrate size and percentages of bedrock, wood, organic matter, and fine sediments (i.e., silt and sand). For each reach I calculated reach- and stream-specific catch per unit effort (CPUE) using mean trap success (mean number crayfishes per trap) as a measure 32 of abundance. Traps are an efficient way to sample crayfishes in vegetated or highly structured environments (Feminella and Resh 1989). Moreover, our streams were spread throughout a range of physiographic provinces with a large array of habitat conditions (turbidity, substrate particle size, macrophyte and woody debris abundance, etc.) so traps were the most replicable method of sampling crayfish abundance (Hubert 1996). In contrast, to guard against potential trap bias against trap-shy species, for crayfish richness I used electrofishing and kick-seining to obtain more accurate estimates (Jordan et al. 2000). Crayfish abundance and richness data were non-normal and could not be normalized using transformations; thus, I used non-parametric Kruskal-Wallace tests (Zar 1999) to test the null hypothesis that crayfish abundance and richness did not differ among streams of contrasting dam conditions (i.e., intact vs. breached vs. relict sites). I used Friedman?s test (Zar 1999) to test the null hypothesis that crayfish abundance and richness did not differ among reaches for streams of similar dam conditions (i.e., upstream vs. mill vs. downstream reaches of relict, breached, and intact dams). Multiple comparisons were made using the Friedman?s Test on each pair of treatments (?=0.05). In addition, I used a General Linear Model (PROC GLM, SAS 9.1) with reach (upstream, mill, downstream) as a fixed factor, site as a random factor, and mean water depth as a covariate was used for each dam condition (intact, breached, relict) to test the null hypothesis that predatory fish abundance did not differ among reaches. Predatory fishes were defined as those individuals in the families Centrarchidae, Cottidae, Esocidae, and Ictaluridae, all of which are 33 known to consume crayfishes (Berkman and Rabeni 1987, Boschung and Mayden 2004). I did not use a minimum size threshold for counting predatory fish and instead assumed that fish large enough to be caught in seines were potential predators of juvenile and/or adult crayfishes. Crayfish population genetics For 2 sites with intact dams, Sandy Creek (Tallapoosa Drainage) and Hatchett Creek (Coosa Drainage), I quantified gene flow among crayfish subpopulations of impounded streams. Genetic analyses focused on Cambarus striatus (the ambiguous crayfish) from Sandy Creek, and Cambarus coosae (the Coosa crayfish) from Hatchett Creek. I chose these species because they were locally common (i.e., had no conservation concerns) and were easily collected and identified in their respective streams. In addition, the strongly disparate distributions of the 2 species promoted a contrast in their degree of endemism. Cambarus striatus is widespread throughout the Southeastern US; in contrast, C. coosae is restricted to the Coosa and Cahaba River Drainages (Hobbs 1981). Tissue samples used for genetic analysis were taken from the abdominal muscle of crayfishes from each reach and stored in 95% ethanol before animals were preserved (Fetzner and Crandall 2003). I extracted genomic DNA using a 2X CTAB extraction protocol as detailed in Coffroth et al. (1992) and amplified a portion of the mitochondrial cytochrome oxidase subunit I (COI) gene using the polymerase chain reaction (PCR) for all individuals using the primers HCO2198 and LCO1490 (Folmer et 34 al. 1994). The COI gene has been used widely in decapod phylogenetic and phylogeography studies because of its relatively high rate of substitution (Tontelj et al. 2005). Reactions were conducted in 25 ?L using the following reagent concentrations: 2.5 ?L 10 x buffer (1.5 ?M), 0.5 ?L dNTPs (10 ?M), 0.1 ?L Taq polymerase, 0.5 ?L magnesium chloride (25 ?M), 1 ?L of each primer (10 ?M), and 1 ?L DNA template (~10-50 ng). Thermocycling was performed with a PTC- 100TM thermocycler (MJ Reactions) using the following program: an initial denaturing step of 96?C for 3 min, followed by 40 cycles of 94?C for 1 min, annealing at 50?C for 1 min, and 72?C for 1 min, with a final elongation of 72?C for 5 min. I purified amplified products with MontageTM PCR Filter Units (Millipore) according to the supplier?s recommendations and sequenced using an ABI 3100 Genetic Analyzer (Applied Biosystems) in both directions. I then edited sequences by comparing each read to its compliment strand using Sequencher v4.6 (Gene Codes Corporation) and aligned manually with Se-Al v2.0a11 (available at http://evolve.zoo.ox.ac.uk/). Genetic data analysis I assessed levels of genetic polymorphism and structure among subpopulations (e.g., upstream, mill, and downstream) of each species with DnaSP v4.06 (Rozas et al. 2003). Specifically, I calculated nucleotide (?) and haplotype (Hd) diversity estimates (Nei 1987) within each subpopulation and overall (i.e., across all 3 subpopulations). In these cases, ? represents the mean 35 number of nucleotide differences per locus between any 2 sequences whereas Hd reflects the haplotype richness within a subpopulation. I estimated genetic differentiation of subpopulations using the nearest neighbor statistic, Snn (Hudson et al. 1992), which quantifies the frequency of the most similar sequence of a given haplotype being recovered from the same locality. I also quantified population structure and gene flow using Fst and Nm (Hudson et al. 1992), where Fst measures the proportion of genetic variation found among subpopulations within a larger population, Nm is the effective number of migrants exchanged between subpopulations per generation, N = the number of individuals in each subpopulation, and m is the fraction of migrants in each subpopulation per generation. To separate population history from population structure, I constructed networks for the COI haplotypes within each species using TCS v1.21 (Clement et al. 2000) and used in a nested clade analysis (NCA, Templeton et al. 1987). I tested the null hypothesis of no geographic association among haplotypes by calculating clade distance (Dc) and nested clade distance (Dn) by 5000 permutations in the GeoDis v2.5 software package (Posada et al. 2000). Dc measures the geographical range of a haplotypes at each nested level whereas Dn measures the evolutionary distance between haplotypes in different levels of the nesting. I assessed the output of GeoDis in the context of the most recent (i.e., April 2009) NCA inference key. This approach can help explain what evolutionary events (e.g., restricted gene flow) led to current levels of genetic 36 diversity and geographic distribution of haplotypes within a species (Templeton 2005). Results Instream habitat, and crayfish and predatory fish assemblage variation Instream Habitat.?There was high variation in substrate characteristics across reaches, with mean particle width ranging from 0.6 to 272.2 mm, mean bedrock cover from 0 to 65%, and mean sand and silt cover from 1 to 78% (Table 2). Only substrate composition differed significantly among reaches, and only for intact dam sites. There, percentage of fine particles (as sand and silt) in the stream bed was significantly lower at mill than at upstream reaches (22.1 vs. 35.6%, respectively, p=0.015); % sand and silt at downstream reaches (28.8%) was not significantly different from upstream or mill reaches (p>0.05). Crayfish.?A total of 20 crayfish taxa was collected from the 22 study sites (Appendix A). Richness and total catch did not differ among streams with different dam conditions (H=0.244, p=0.885 and H=0.033, p=0.984, respectively). However, when expressed as CPUE there were significant differences among dam conditions (H=7.923, p=0.019). Mean CPUE for streams with breached dams was significantly lower than that of relict and intact dams (Fig. 3A). 37 Table 2. Summary of physicochemical and crayfish data from the study streams. Up= site 1500-5000 m upstream of the dam, Mill= site immediately downstream of the dam, Down= site 500-3000 m downstream of the dam. Cond. = specific conductance, DO = dissolved oxygen, Temp = water temperature, CPUE=crayfish catch per trap and crayfish species found at each reach type. Mean + SD. Dam condition Study reach Current velocity (m/s) Stream Depth (m) Wetted width (m) Mean substrate (mm) % Bedrock % Woody debris % Organic matter % Sand / silt Cond. (uS/cm) pH DO (mg/L) Temp (C) CPUE Intact Overall 0.05 +0.04 0.26 +0.12 11.97 +5.03 116.91 +59.59 18.50 +16.56 9.41 +9.24 8.44 +7.42 28.85 +13.63 111.92 +134.69 8.33 +0.58 6.18 +1.20 22.93 +5.29 3.00 +0.15 Up 0.04 +0.03 0.26 +0.15 10.46 +5.26 113.10 +63.99 16.18 +15.87 11.57 +10.6 8 8.29 +5.37 35.64 +14.94 114.63 +134.49 8.52 +0.69 5.88 +0.99 21.29 +5.67 4.34 +0.48 Mill 0.07 +0.05 0.26 +0.11 13.09 +4.75 133.47 +53.42 18.02 +13.41 6.22 +6.26 8.68 +8.66 22.09 +11.45 116.88 +149.93 8.12 +0.53 5.94 +1.47 24.24 +4.90 1.58 +0.31 Down 0.05 +0.05 0.26 +0.10 12.35 +5.22 103.78 +63.60 21.31 +20.91 10.44 +10.2 0 8.35 +8.58 28.81 +11.92 105.11 +137.28 8.38 +0.56 6.71 +1.08 23.28 +5.54 3.08 +.049 Breached Overall 0.16 +0.07 0.28 +0.11 13.08 +6.12 138.88 +77.83 20.50 +17.19 5.98 +4.35 7.91 +7.11 29.21 +19.04 313.71 +420.00 8.47 +0.50 7.62 +2.00 21.09 +4.00 0.99 +0.06 Up 0.16 +0.07 0.30 +0.30 14.04 +8.18 124.36 +101.94 24.14 +17.43 8.03 +3.89 7.79 +6.32 35.70 +29.03 326.71 +446.05 8.40 +0.61 6.28 +2.32 20.27 +4.61 1.73 +0.21 Mill 0.15 +0.09 0.30 +0.11 13.96 +3.89 143.22 +55.38 18.17 +14.53 5.94 +4.61 7.36 +8.87 27.37 +9.73 240.14 +225.43 8.40 +0.57 7.55 +1.54 21.85 +3.31 0.56 +0.11 Down 0.18 +0.05 0.25 +0.06 11.23 +6.06 149.08 +79.66 19.20 +21.14 3.96 +4.13 8.59 +7.01 24.57 +13.88 374.29 +572.21 8.60 +0.32 8.76 +1.64 21.17 +4.56 0.69 +0.15 Relict Overall 0.11 +0.06 0.32 +0.17 19.28 +0.73 109.98 +127.81 18.96 +21.08 7.25 +5.05 7.11 +5.43 31.41 +16.11 146.89 +81.43 8.23 +0.15 5.99 +1.91 26.76 +1.52 2.77 +0.17 Up 0.08 +0.05 0.30 +0.15 19.85 +12.0 4 65.61 +75.06 11.18 +13.29 9.52 +2.98 10.26 +5.55 33.58 +22.26 147.00 +91.33 8.20 +0.10 3.82 +0.10 26.60 +1.65 1.85 +1.36 Mill 0.14 +0.09 0.28 +0.12 19.83 +9.66 163.57 +165.24 23.58 +21.61 4.60 +4.36 6.42 +4.99 24.86 +8.69 144.67 +92.68 8.17 +0.12 7.41 +0.12 26.00 +2.83 3.50 +3.54 Down 0.11 +0.05 0.38 +0.25 18.17 +5.15 100.74 +134.44 22.12 +28.22 7.64 +6.78 4.66 +5.19 35.80 +15.93 149.00 +97.86 8.33 +0.21 6.75 +0.21 27.43 +0.12 2.95 +2.71 38 Figure 3. Catch per trap effort of crayfish on streams with intact, breached, and relict dams across all study reaches (a), catch per trap effort of crayfish at each study reach on streams with dams in different conditions (b), and abundance of predatory fishes (as total catch) (c), at each study reach on streams for dams in different conditions (see text). White columns represent upstream reaches, black columns are dam reaches, and gray columns are downstream reaches. Mean + SE. 39 For streams with intact dams, CPUE was lower at mill reaches than at upstream or downstream reaches (Xr2=21.88, p<0.0001). For streams with breached dams, CPUE was significantly higher upstream than at mill or downstream sites (Xr2=10.83, p=0.005). For streams with relict dams, CPUE did not differ among sites (Xr2=1.90, p=0.39; Fig. 3B). Predatory fish.?Individuals from 5 fish families known to consume crayfishes were collected during the study (B. Helms unpublished data). For streams with intact dams, there were significantly higher abundances of predatory fishes at mill reaches (i.e., immediately below dams) than at upstream or downstream reaches (Fig. 3C). The numerically dominant taxa of predators were centrarchids. Crayfish population genetics A total of 562 base pairs of COI were obtained from each of 22 Cambarus striatus at Sandy Creek. From these, 19 (3.38%) polymorphic sites were recovered from 12 haplotypes found at all 3 reaches (upstream, mill, and downstream). ? was higher upstream than overall or at mill or downstream reaches, whereas Hd was higher overall than at any single reach except the mill reach, where only 2 individuals were collected, and each represented a different haplotype. Snn was significant and approached 1 (Snn=0.705, p=0.009), implying genetic differentiation between upstream and downstream subpopulations. Furthermore, whereas Fst was not high (0.378), Nm was <1 (0.410), suggesting limited gene flow among subpopulations (Table 3). Of the 12 haplotypes, 8 40 (67%) were found at the downstream reach, with 2 being shared by downstream and mill reaches, and 4 haplotypes found at the upstream reach. The upstream reach shared no haplotypes with downstream and mill reaches (Fig. 4A), again suggesting limited to no longitudinal gene flow among subpopulations. The NCA of C. striatus found one 2-step clade to be significant (2-1: Dn=1.420, p<0.0001; I-T: Dc=-1.467, p=0.013), with a conclusion of contiguous range expansion based on the inference key. All other clades yielded inconclusive outcomes in the NCA. A total of 598 base pairs of COI were obtained from each of 50 Cambarus coosae at Hatchett Creek. Overall, 16 haplotypes were collected across the 3 study reaches, with a total of 15 (2.51%) polymorphic sites. ? was higher upstream than overall or at mill or downstream reaches, with Hd being highest at the mill reach. Fst was low, whereas Nm was high, but Snn was not significant (Table 3), suggesting little to no genetic differentiation among subpopulations. This pattern was visually apparent in the TCS network: of the 16 identified haplotypes, the most commonly recovered (i.e., haplotype 1) was found at the upstream, mill, and downstream sites at a similar frequency. The remaining haplotypes were singletons and/or unique to a reach (Fig. 4B). The NCA of C. coosae revealed no clades with significant divergence. Discussion A wide range of biotic and abiotic factors influence crayfish abundance and distribution in streams. Abiotic factors involve appropriate substrate sizes and/or types (i.e., the availability of refugia), water depth, hydrologic 41 Table 3. Genetic diversity and differentiation among subpopulations of the crayfish Cambarus striatus and C. coosae. ? = nucleotide diversity, Hd = haplotype diversity, Fst = genetic variation among subpopulations within the metapopulation, and Nm = effective number of migrants exchanged between subpopulations per generation, Snn = estimate of genetic differentiation of the crayfish subpopulations (see text). Species Population No. of indiv. No. of haplo- types ? Hd Fst Nm Snn C. striatus Overall 22 12 0.0107 0.887 0.3778 0.41 0.70519 p=0.009 Upstream 6 4 0.0108 0.867 Mill 2 2 0.0018 1.0000 Downstream 16 8 0.0084 0.824 C. coosae Overall 50 16 0.0016 0.593 0.0094 26.81 0.35432 Upstream 15 7 0.0019 0.625 Mill 13 5 0.0013 0.667 Downstream 22 8 0.0015 0.545 42 Figure 4. Halpotype networks showing nesting levels used to infer historical processes for Cambarus striatus at Sandy Creek (a) and C. coosae at Hatchett Creek (b). Numbered circles each represent a unique sampled haplotype whereas small open circles represent unsampled (i.e. missing) haplotypes. The size of a circle is proportional to the frequency at which that haplotype was recovered. Shading corresponds to sites where individuals were collected (white = upstream, black = mill, and gray = downstream). Note that despite variable lengths, each branch implies a single mutational difference between haplotypes. For example, haplotype 1 in 4a differs from haplotype 5 by one mutation, whereas haplotype 5 differs from haplotype 7 by 4 mutations. 43 permanence, and water chemistry (Bovbjerg 1970, Jordan et al. 2000, Flinders 2003, Ratcliffe and DeVries 2004, Larson et al. 2009). Biotic factors can include inter- and intraspecific competition and predation by fish, wading birds, and mammals, each of which may affect crayfish behavior, distribution, and overall assemblage structure (Bovbjerg 1970, Stein and Magnuson 1976, Garvey et al. 1994, Englund and Krupa 2000). On streams with intact mill dams crayfish I found crayfish abundance to be lower at mill reaches than at upstream or downstream reaches. In addition, abundance of predatory fish was higher at mill reaches (vs. upstream or downstream) of these same streams. Dams may serve as sources of fish aggregations downstream of the impoundment (Agostinho et al. 2007), which may, in turn, act as ?predator gauntlets? (sensu Hein and Crowl 2010, see also Creed 2006), decreasing crayfish abundance in these reaches through direct consumption by fish, behavioral avoidance, or a combination of these factors. Irrespective of the source, presence of predators immediately downstream of impoundments may impose additional limits on longitudinal dispersal by crayfishes, thus exacerbating the influence of the physical barrier on crayfish movements. Interactions between abiotic factors and biotic factors also may be driving reductions in crayfish abundance at intact mill sites. Our data did show evidence of reduced amounts of fine sediments immediately downstream of intact dams (unpublished data). Decreased current velocity in the impounded reaches can greatly increase sediment deposition upstream of the dam, often causing tailwaters to become sediment-starved and downstream sections to exhibit 44 increased scour, streambed lowering, and bed coarsening (Baxter 1977, Chien 1985, Graf 2005). Decreased deposition causes tailwaters to become less geomorphically complex, with fewer bars and islands and reduced shallow-water habitat (Poff et al. 1997, Graf 2006). Moreover, coarsening of the stream bed reduces habitat availability for benthic species by decreasing habitat heterogeneity (Hauer et al. 1989, Armitage and Blackburn 1990). At the watershed scale, these geomorphic changes may transform slow-flowing marshy streams into fast- flowing gravel-bottomed systems (Walter and Merritts 2008). Our habitat data revealed few significant differences among sites on streams with intact dams, but the influence of abiotic factors on aquatic populations may be transient and/or difficult to detect empirically because of high spatial and temporal variation; thus, it is possible that our sampling regime may have not adequately characterized the potentially critical habitat conditions affecting crayfishes. It is important to note that there was high variability in structural conditions of breached dams, as well as habitat across these breached sites. In this context, dams at most (60%) of our breached sites were breached intentionally by landowners or managers for safety or conservation concerns, whereas all other breaches ostensibly occurred by natural erosion during storm events. Some breaches (both intentional and natural) were the result of removal of a small portion of the dam, whereas the spillways were opened on others allowing free- flowing conditions. Duration of time following intentional breaches was easily determined from the landowners, whereas time since natural breaches often was less clear, as these structures often were frequently unmaintained and/or occurred 45 in remote areas. According to landowner accounts, most dams were breached between the late 1980s and the early 1990s (E. Hartfield, personal observation), indicating a 20- to 30-y post-breach period for biotic and abiotic conditions to change within our study sites. Surprisingly, breached dams appeared to have a greater negative impact on crayfish abundance than intact or relict dams, as many breached mill and downstream reaches supported fewer crayfishes than upstream reaches. Large deposits of fine sediments are characteristic of the reaches immediately upstream of breached (and intact) dams, often extending for >1000 m upstream of impoundments. At the dam, the breach often constricted the stream to a narrow exit point where water velocity is greatly increased during major storm events, resulting in a ?pressure hose effect.? It is possible that episodic pulses of fine sediments from behind the breached dams may reduce the availability of crevices, interstitial space, and other refugia, reducing crayfish habitat quality, heterogeneity, and stability for considerable distances downstream of the breach. This effect could explain why our sampling regime did not show higher fine sediment loads in mill or downstream reaches compared with upstream reaches. The lack of significant differences in abundance among sites on streams with relict dams suggests that after total dam removal, crayfish abundance homogenizes longitudinally and may return to pre-impoundment levels. Assessment of recovery after dam removal was beyond the scope of this paper, as exact estimates of time since dam failure at relict sites often could not be obtained. According to landowner accounts, most relict dams failed between the 46 1940s and 1960s; if these estimates are correct, then recovery of crayfish assemblages may take years to decades, depending on how much of the structure was removed. The time for captured sediment in formerly impounded reaches to move through the system may play a key role in recovery time, making time since breach or removal of dams an important factor when studying recovery time of these systems. Unfortunately, such historical data are not typically available. Genetic diversity measures (Hd and ?) for C. striatus in Sandy Creek were not higher downstream than upstream, as would often be expected if upstream subpopulations were isolated for long periods and experienced subsequent reductions in diversity. However, values estimated for Snn, Fst and Nm all suggest genetic structure exists for C. striatus in Sandy Creek, with differentiation and limited gene flow among subpopulations. Abundance of C. striatus at the mill reach was low, but the 2 haplotypes found also occurred downstream, suggesting unrestricted gene flow between these reaches. The upstream reach shared no haplotypes with downstream or mill reaches, but 2 haplotypes found downstream (haplotypes 10 and 12 in Fig. 4A) were more similar to upstream haplotypes than to other downstream haplotypes. Presence of upstream haplotypes in downstream reaches implies the occurrence of migration/dispersal over the dam, but only in the downstream direction. Such downstream unidirectional movement across a dam has been documented in fish populations (Neraas and Spruell 2001). In contrast, for C. coosae in Hatchett Creek, Snn, Fst and Nm values suggested no population structuring. This result also was reflected in the haplotype network, where one haplotype was numerically dominant across all 3 47 subpopulations. Overall genetic diversity (as Hd and ?) for the endemic C. coosae was much lower than the widespread C. striatus, which is consistent with the idea that restricted-range endemic species are less genetically diverse than widespread species (Frankham 1997). Differences in apparent gene flow between the 2 species also may be the result of the relatively young age of the Hatchett Creek dam (~80 y) compared to the age of Jones Mill dam on Sandy Creek (at least 160 y). This time difference would be equal to many generations of crayfishes, and may explain why the COI gene fragment used in this study revealed no population structuring. It is possible, that use of more sensitive genetic markers, such as microsatellites (Avise 2004), might detect finer-scale (i.e., shorter-term) genetic structuring within this system. Aging and degraded low-head dams are a hazard to recreational activities and may also threaten survival of sensitive aquatic taxa. Our data suggest that streams with intact dams negatively affect crayfishes by creating predaceous fish aggregations downstream of dams that may reduce crayfish abundance. In addition, low-head dams and their resulting reservoirs have the potential to limit longitudinal movements by crayfishes, and serve as an additional source of fragmention of these populations. As such, dams make excellent targets for restoration projects. However, partial removal (i.e., breaching) of these structures may actually increase threats to downstream biota because of decreased habitat stability due to potentially rapid and catastrophic delivery of multiple decades of sediment buildup directly behind dams. When designing instream restorations in such regulated systems that support at-risk freshwater biota, extensive case-by- 48 case evaluations are needed to weigh the costs and benefits of dam removal (Stanley and Doyle 2003). 49 References Agostinho, C.S., C.R. Pereira, J.J. Oliveira, I.S. Frieta, E.E. Marque. 2007. Movements through a fish ladder: temporal patterns and motivations to move upstream. Neotropical Ichthyology 5:161-167. Armitage, P.D., and J.H. Blackburn. 1990. Environmental stability and communities of Chironomidae (Diptera) in a regulated river. Regulated Rivers: Research and Management 5:319-328. Avise, J.C. 2004. Molecular markers, natural history, and evolution. 2nd edition. Sinauer Associates, Inc, Sunderland, Massachusetts. Baxter, R.M. 1977. Environmental effects of dams and impoundments. Annual Review of Ecology and Systematics 8:255-283. Berkman, H. E., and C. F. Rabeni. 1987. Effect of siltation on stream fish communities. Experimental Biology of Fishes 18:285-294. Boschung, H. T., and R. L. Mayden. 2004. Fishes of Alabama. Smithsonian Books, Washington D.C. Bovjerg, R.V. 1970. Ecological isolation and competitive exclusion in two crayfish (Orconectes virilis and Orconectes immunis). Ecology 51:225- 236. Capelli, G.M., and J.J. Magnuson. 1983. Morphoedaphic and biogeographic analysis of crayfish distribution in northern Wisconsin. Journal of Crustacean Biology 3:548-564. Charlesworth, D. and B. Charlesworth. 1987. Inbreeding depression and its evolutionary consequences. Annual Review of Ecology and Systematics 18:237-268. 50 Chien, N. 1985. Changes in river regime after the construction of upstream reservoirs. Earth Surface Processes and Landforms 10:143-159. Clement, M., D. Posada, K. Crandall. 2000. TCS: a computer program to estimate gene genealogies. Molecular Ecology 9:1657-1659. Coffroth, M.A., H.R. Lasker, M.E. Diamond, J.A. Bruenn, and E. Bermingham. 1992. DNA fingerprinting of a gorgonian coral: a method for detecting clonal structure in a vegetative species. Marine Biology 114:317-325. Creed, R.P. 2006. Predator transitions in stream communities: a model and evidence from field studies. Journal of the North American Benthological Society 25:533-544. Crnokrak, P., and D.A. Roth. 1999. Inbreeding depression in the wild. Heredity 83:260-270. Cushman, R.M. 1985. Review of ecological effects of rapidly varying flows downstream of hydroelectric facilities. North American Journal of Fisheries Management 5:330-339. Dean, J., D. Edds, D. Gillette, J. Howard, S. Sherraden, and J. Tiemann. 2002. Effects of lowhead dams on freshwater mussels in the Neosho River, Kansas. Transactions of the Kansas Academy of Science 105:323-240. Englund, G., and J.J. Krupa. 2000. Habitat use by crayfish in stream pools: influence of predators, depth, and body size. Freshwater Biology 43:75- 83. Fagan, W.F. 2002. Connectivity, fragmentation, and extinction risk in dendritic metapopulations. Ecology 83:3243-3249. 51 Feminella, J.W., and V.H. Resh. 1989. Submersed macrophytes and grazing crayfish: an experimental study of herbivory in a California freshwater marsh. Holarctic Ecology 12:1-8. Fetzner, J. W., and K. A. Crandall. 2003. Linear habitats and the nested clade analyses: an empirical evaluation of geographic versus river distances using an Ozark crayfish (Decapoda: Cambaridae). Evolution 57:2101- 2118. Finn, D.S., M.S. Blouin, and D.A. Lytle. 2007. Population genetic structure reveals terrestrial affinities for a headwater stream insect. Freshwater Biology 52:1881-1897. Flinders, C.A., and D.D. Magoulick. 2003. Effects of stream permanence on crayfish community structure. American Midland Naturalist 149:134-147. Folmer, O., M. Black, W. Hoeh, R. Lutz, and R. Vrijenhoek. 1994. DNA primers for amplification of mitochondrial cytochrome c oxidase subunit I from diverse metazoan invertebrates. Molecular Marine Biology and Biotechnology 3:294-299. Frankham, R. 1997. Do island populations have less genetic variation than mainland populations? Heredity 78:311-327. Fraser, J.C. 1972. Regulated discharge and the stream environment. Pages 263- 286 in R. Olgesby, C.A. Carlson, and J. McCann (editors). River ecology and management. Academic Press, New York. 52 Gangloff, M.M. and J.W. Feminella. 2007. Stream channel geomorphology influences mussel abundance in southern Appalachian streams. Freshwater Biology 52:64-74. Garvey, J.E., R.A. Stein, and H.M. Thomas. 1994. Assessing how fish predation and interspecific prey competition influence a crayfish assemblage. Ecology 75:532-547. Gehrke, P.C., D.M. Gilligan, and M. Barwick. 2002. Changes in fish communities of the Shoalhaven River 20 years after construction of Tallowa Dam, Australia. River Research and Applications 18:265-286. Gibbons, J.W. 1976. Thermal alteration and the enhancement of species populations. Pages 27-31 in G.W. Esch and R.W. McFarlane, editors. Thermal Ecology II, ERDA Symposium Series. Gore, J.A. 2006. Discharge measurements and streamflow analysis. Pages 51- 101 in F.R. Hauer and G.A. Lamberti (editors). Methods in stream ecology. 2nd edition. Academic Press, Amsterdam. Graf, W.L. 2005. Geomorphology and American dams: the scientific, social, and economic context. Geomorphology 71:3-26. Graf, W.L. 2006. Downstream hydrologic and geomorphic effects of large dams on American rivers. Geomorphology 79:336-360. Hackney, C.T., S.M Adams, and W.H. Martin (editors). 1993. Biodiversity of the Southeastern United States. John Wiley & Sons, New York. Hansson, B., and L. Westerberg. 2002. On the correlation between heterzygosity and fitness in natural populations. Molecular Ecology 11:2467-2474. 53 Hauer, F.R., J.A. Stanford, J.V. Ward. 1989. Serial discontinuities in a Rocky Mountain river. II. Distribution and Abundance of Trichoptera. Regulated Rivers: Research and Management 3:177-182. Hein, C.L., and T.A. Crowl. 2010. Running the predator gauntlet: do freshwater shrimp (Atya lanipes) migrate above waterfalls to avoid fish predation? Journal of the North American Benthological Society 29:431-443. Hobbs, H.H., Jr. 1981. The crayfishes of Georgia. Smithsonian Contributions to Zoology 318:1-549. Hobbs, H.H., Jr. 1989. An illustrated checklist of the American crayfishes (Decapoda: Astacidae, Cambaridae, and Parastacidae). Smithsonian Contributions to Zoology 480:1-236. Hubert W.A. 1996. Passive capture techniques. Pages 157-192 in L.A. Nielson and D.L. Johnson (editor). Fisheries techniques. 2nd edition. American Fisheries Society, Bethesda, Maryland. Hudson, R. R., D.D. Boos, and N.L. Kaplan. 1992. A statistical test for detecting geographic subdivision. Molecular Biology and Evolution 9:138-151. Hughes, J.M., D.J. Schmidt, and D.S. Finn. 2009. Genes in streams: using DNA to understand the movement of freshwater fauna and their riverine habitat. BioScience 59:573-583. IFC Consulting. 2005. A summary of existing research on low-head dam removal Projects. Technical Report, American Association of State Highway and Transportation Officials (AASHTO). 54 Irvine, J.R. 1985. Effects of successive flow perturbations on stream invertebrates. Canadian Journal of Fisheries and Aquatic Sciences 42:1922-1927. Jager, H.I., J.A. Chandler, K.B. Lepla, and W.V. Winkle. 2001. A theoretical study of river fragmentation by dams and its effects on white sturgeon populations. Environmental Biology of Fishes 60:347-361. Jansson, R., C. Nilsson, and B. Renofalt. 2000. Fragmentation of riparian floras in rivers with multiple dams. Ecology 81:899-903. Jordan, F., K.J. Babbitt, C.C. McIvor, and S.J. Miller. 2000. Contrasting patterns of habitat use by prawns and crayfish in headwater marsh of the St. John River, Florida. Journal of Crustacean Biology 20:769-776. Kelner, D.E., and B.E. Sietman. 2000. Relict populations of the Ebony Shell, Fusconaia ebena (Bivalvia: Uniondiae), in the Upper Mississippi River Drainage. Journal of Freshwater Ecology 15:371-378. Lande, R. 1988. Genetics and demography in biological conservation. Science 241:1455-1460. Lande, R. 1999. Extinction risks from anthropogenic, ecological, and genetic factors. Pages 1-22 in L.F. Landweber and A.P. Dobson (editors). Genetics and the extinction of species. Princeton University Press, Princeton, New Jersey. Larinier, M. 2001. Environmental issues, dams and fish migration. Pages 45-89 in G. Marmulla (editor). Dams, fish and fisheries: opportunities, challenges and conflict resolution. FAO Fish Technical Paper, Rome. 55 Larson, E.R., D.D. Magoulick, C. Turners, and K.H. Laycock. 2009. Disturbance and species displacement: different tolerances to stream drying and desiccation in a native and an invasive crayfish. Freshwater Biology 54:1899-1908. Lessard, J.L., and D.B. Hayes. 2003. Effects of elevated water temperature on fish and macroinvertebrate communities below small dams. River Research and Applications 19:721-732. Lodge, D. M., C. A. Taylor, D. M. Holdich, and J. Skurdal. 2000. Nonindigenous crayfishes threaten North American freshwater biodiversity. Fisheries 25(8):7-20. MacArthur, R.H., and E.O. Wilson. 1967. The theory of island biogeography. Princeton University Press, Princeton, New Jersey. McLaughlin, R.L., L. Porto, D.L.G. Noakes, J.R. Baylis, L.M. Carl, H.R. Dodd, J.D. Goldstein, D.B. Hayes, and R.G. Randall. 2006. Effects of low-head barriers on stream fishes: taxonomic affiliations and morphological correlates of sensitive species. Canadian Journal of Fisheries and Aquatic Sciences 63:766-779. Morita, K., and S. Yamamoto. 2002. Effects of habitat fragmentation by damming on the persistence of stream-dwelling charr populations. Conservation Biology 16:1318-1323. Nei, M. 1987. Molecular evolutionary genetics. Columbia University Press, New York. Neraas, L.P., and P. Spruell. 2001. Fragmentation of riverine systems: the 56 genetic effects of dams on bull trout (Salvelinus confluentus) in the Clark Fork River system. Molecular Ecology 10:1153-1164. Olden, J.D., D.A. Jackson, and P.R. Peres-Neto. 2001. Spatial isolation and fish communities in drainage lakes. Oecologia 127:572-585. Penn, G. H. 1956. The genus Procambarus in Louisiana. American Midland Naturalist 6:406-422. Perry, S.A., W.B. Perry, and J.A. Stanford. 1987. Effects of thermal regime on size, growth rates and emergence of two species of stoneflies (Plecoptera: Taeniopterygidae, Pteronarcyidae) in the Flathead River, Montana. American Midland Naturalist 117:83-93. Poff, N.L., J.D. Allan, M.B. Bain, J.R. Karr, K.L. Prestegaard, B.D. Richter, R.E. Sparks, and J.C. Stromberg. 1997. The natural flow regime. BioScience 47:769-784. Posada, D., K.A. Crandall, and A.R. Templeton. 2000. GeoDis: a program for the cladistic nested anaylsis of the geographical distribution of genetic haplotypes. Molecular Ecology 9:487-488. Rabeni, C.F., K.J. Collier, and S.M. Parkyn. 1997. Evaluating techniques for sampling stream crayfish (Paranephrops planiforms). New Zealand Journal of Marine and Freshwater Research 31:693-700. Ratfliffe, J.A., and D.R. DeVries. 2004. The crayfishes (Crustacea: Decapoda) of the Tallapoosa River Drainage, Alabama. Southeastern Naturalist 3:417- 430. Raymond, H.L. 1979. Effects of dams and impoundments on migrations of 57 juvenile Chinook salmon and steelhead from the Snake River, 1966 to 1975. Transactions of the American Fisheries Society 108:505-529. Reed, D.H., and R. Frankham. 2003. Correlation between fitness and genetic diversity. Conservation Biology 17:230-237. Rozas, J., J.C. Sanchez-DelBarrio, X. Messeguer, and R. Rozas. 2003. DnaSP, DNA polymorphism analysis by the coalescent and other methods. Bioinformatics 19:2496-2497. Saunders, D.A., R.J. Hobbs, and C.R. Margules. 1991. Biological consequences of ecosystem fragmentation: a review. Conservation Biology 5:18-32. Stanley, E.H. and M.W. Doyle. 1993. Trading off: the ecological effects of dam removal. Frontiers in Ecology and the Environment 1:15-22. Stein, R.A., and J.J. Magnuson. 1976. Behavioral response of crayfish to a fish predator. Ecology 57:751-761. Steinman, A.D., G.A. Lamberti, and P.R. Leavitt. 2006. Biomass and pigments of benthic algae. Pages 357-379 in F.R. Hauer and G.A. Lamberti (editors). Methods in stream ecology. 2nd edition. Academic Press, Amsterdam. Taylor, C.A. and G.A. Schuster. 2004. The crayfishes of Kentucky. Illinois Natural History Survey. Special Publication No. 28:1-219. Taylor, C.A., G.A. Schuster, J.E Cooper, R.J. DiStefano, A.G. Eversole, R. Hamr, H.H. Hobbs III, H.W. Robinson, C.E. Skelton, and R.G. Thoma. 2007. A reassessment of the conservation status of crayfishes of the United States and Canada after 10+ years of increased awareness. Fisheries 32(8):372- 389. 58 Templeton, A.R. 2005. Inference key for the nested haplotype tree analysis of geographical distances. http://darwin.uvigo.es/software/geodis.html Templeton, A.R., E. Boerwinkle, and C.F. Sing. 1987. A cladistic analysis of phenotypic associations with haplotypes from restriction endonuclease mapping. I. Basic theory and analysis of alcohol dehydrogenase activity in Drosophila. Genetics 117:343-351. Tontelj, P., Y. Machino, and B. Sket. 2005. Phylogenetic and phylogeographic relationships in the crayfish genus Austropotamobius inferred from mitochondrial COI gene sequences. Molecular Phylogenetics and Evolution 34:212-226. Townsend, C.R. 1989. The patch dynamics concept of stream community ecology. Journal of the North American Benthological Society 8:36-50. Travnicheck, V.H., M.B. Bain, and M.J. Maceina. 1995. Recovery of a warmwater fish assemblage after the initiation of minimum-flow release downstream from a hydroelectric dam. Transactions of the American Fisheries Society 124:836-844. Vinson, M.R. 2001. Long-term dynamics of an invertebrate assemblage downstream from a large dam. Ecological Applications 11:711-730. Viosca, H.J. 1953. All about crawfish: life history and habits. Louisiana Conservation 5:3-5. Walter, R.C., and D.J. Merritts. 2008. Natural streams and the legacy of water powered mills. Science 319:299-304. Watanabe, K. and T. Omura. 2007. Relationship between reservoir size and 59 genetic differentiation of the stream caddisfly Stenopsyche marmorata. Biological Conservation 136:203-211. Watters, G.T. 1996. Small dams as barriers to freshwater mussels (Mollusca: Pelecypoda: Unionidae) and their hosts. Biological Conservation 75:79- 85. Wolman, M.G. 1954. A method of sampling coarse river-bed material. Transactions of the American Geophysical Union 15:951-956. Wotton, R.S. 1995. Temperature and lake-outlet communities. Journal of Thermal Biology 20:121-125. Wu, J., J. Huang, X. Han, Z. Xie, and X. Gao. 2003. Three-Gorges Dam: experiment in habitat fragmentation? Science 300:1239-1240. Yamamoto, S. K. Morita, I. Koizumi, and K. Maekawa. 2004. Genetic differentiation of white-spotted charr (Salvelinus leucomaenis) populations after habitat fragmentation: spatial-temporal changes in gene frequencies. Conservation Genetics 5:529-538. Young, A., T. Boyle, and T. Brown. 1996. The population genetic consequences of habitat fragmentation for plants. Trends in Ecology and Evolution 11:413-418. Zar, J.H. 1999. Biostatistical analysis. 4th edition. Prentice-Hall, Upper Saddle River, New Jersey. 60 Appendix A. Crayfish species, streams where each species was collected, drainage, and study reaches (Up=upstream reach; Mill=mill reach; Down=downstream reach; X=present). Species Streams Drainage Up Mill Down Cambarus coosae Little Cahaba Creek Cahaba X X X Hatchett Creek Coosa X X X Big Canoe Creek Coosa X X X Yellow Leaf Creek Coosa X X X C. englishi Little Hillabee Tallapoosa X X C. girardianus Turkey Creek Tennessee X C. halli Choctafaula Creek Tallapoosa X X X Sandy Creek Tallapoosa X X X Loblokee Creek Tallapoosa X X X C. howardi Halawakee Creek Chatahoochee X X X Osanippa Creek Chatahoochee X X X C. latimanus Halawakee Creek Chatahoochee X X Hatchett Creek Coosa X X Big Canoe Creek Coosa X X X C. striatus Blue Springs Creek Black Warrior X X X Brushy Creek Black Warrior X X Little Cahaba River Cahaba X X Cahaba River Cahaba X Little Uchee Creek Chatahoochee X X Choctafaula Creek Tallapoosa X X Little Hillabee Creek Tallapoosa X X Sandy Creek Tallapoosa X X X Turkey Creek Tennessee X Pearce's Mill Creek Tombigbee X X New River Tombigbee X C. obstipus Brushy Creek Black Warrior X X Pearce?s Mill Creek Tombigbee X Falicambarus fodiens Choctafaula Creek Tallapoosa X Orconectes erichsonianus Blue Springs Creek Black Warrior X X X Cahaba River Cahaba X X Big Canoe Creek Coosa X X X Paint Rock River Tennessee X X O. forceps Paint Rock River Tennessee X X X 61 Appendix A. Crayfish species, streams where each species was collected, drainage, and study reaches (Up=upstream reach; Mill=mill reach; Down=downstream reach; X=present). Species Streams Drainage Up Mill Down O. holti Big Flat Creek Alabama X X O. perfectus New River Tombigbee X X X Lost Creek Black Warrior X X X O. putnami Turkey Creek Tennessee X X X O. spinosus Turkey Creek Tennessee X X O. validus Brushy Creek Black Warrior X X X Paint Rock River Tennessee X X Turkey Creek Tennessee X Buttahatchee River Tombigbee X X X Pearce?s Mill Creek Tombigbee X X X O. virilis Cahaba River Cahaba X X X Little Cahaba River Cahaba X X X Procambarus spiculifer Big Flat Creek Alabama X X X Halawakee Creek Chatahoochee X X X Little Uchee Creek Chatahoochee X X X Osanippa Creek Chatahoochee X X X Little Hillabee Tallapoosa X X X Loblockee Creek Tallapoosa X P. versutus Yellow Leaf Creek Coosa X X Choctafaula Creek Tallapoosa X X X Loblockee Creek Tallapoosa X X P. verrucosus Choctafaula Creek Tallapoosa X X Sandy Creek Tallapoosa X