fish assemblage structure, trophic ecology, and ...

FISH ASSEMBLAGE STRUCTURE, TROPHIC ECOLOGY, AND POTENTIAL EFFECTS OF INVADING ASIAN CARPS IN THREE MISSOURI RIVER TRIBUTARIES, SOUTH DAKOTA

BY CARI-ANN HAYER

A dissertation submitted in partial fulfillment of the requirements for the Doctor of Philosophy Major in Wildlife & Fisheries Sciences Specialization in Fisheries Sciences South Dakota State University 2014

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“In Japan, the Carp symbolizes strength, perseverance, the courage and ability to attain high goals while overcoming life’s difficulties – leading to consequent success.” – Unknown This dissertation is dedicated to my teacher, my inspiration, my friend, Rob Klumb

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ACKNOWLEDGEMENTS I would first like to acknowledge my advisors, Dr. Brian Graeb and Dr. Katie Bertrand, who entrusted me to conduct research on what I think was undoubtedly the greatest and most fascinating PhD project ever. They also pushed me throughout my PhD journey to become a better researcher, teacher, scientist and scholar. I would like to show gratitude toward my other committee members, Dr. Steve Chipps, Dr. Rob Klumb and Dr. Mike Wimberley, for their various inputs into this project and for their inquiries and insights, especially during my comprehensive exams, which forced me out of my comfort level and afforded me the opportunity to become an improved and more well-rounded scientist. Additionally, I learned that a person cannot know everything about everything and it is okay to state: “I don’t know”. I would like to express my gratitude to Jessica Howell for her unrelenting, enthusiastic and verbose assistance with fieldwork and for being a remarkable friend who always made me laugh. Way to go “Asian Carp Search and Seizure Crew”! I would like to acknowledge my “Comprehensive Exams Discussion Group” for many intense discussions, enlightening questions, numerous libations and priming me for my comprehensive exams. The core group consisted of: Jason Breeggemann, Dan Dembkowski, Dave Deslauriers, Mark Kaemingk and Tobias Rapp. I am grateful to all the technicians, both undergraduate and recently graduated, who worked with me over the years both in the field and in the laboratory. Without them, this project would never have come to fruition, the field work days would not have been as enjoyable and I would not have learned so much.

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I am thankful for having known Dr. Jenks during my tenure at South Dakota State University. Thanks for letting me use your desk space! I also want to thank Dr. Grovenburg and Dr. Jenks for introducing me to the Great Plains Natural Science Society and hiring me as an Outreach Coordinator and member of the workshop committees. This was an enlightening undertaking in which I learned to appreciate the inner workings of how a Journal and Society operates. I am particularly appreciative to the staff in the Natural Resource Department who kept things together and helped to keep my project on track. These hard-working people include Diane Drake, Carol Jacobsen, Terri Symens, Kate Tvedt and Dawn Van Ballegooyen. I would like to thank my friend and mentor, Dr. Steve Chipps for his advisement on various issues that were novel to me from reservoir productivity to mercury bioaccumulation. I am grateful to him for “keeping me around” long enough to find this PhD project. Without his encouragement and mentoring before this project, I probably would not have taken and/or considered this position. I would like to thank Dr. Charles “Chuck” Berry for being a helpful and inspiring mentor and supervisor that always pressed me to do my best work and who was always there when I needed something. Without Dr. Berry I never would have accepted the position to take on this PhD project. As such, I would not have been equipped for this project without his guidance.

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I would like to thank my fiancé, Jason, for his unwavering support, especially through comprehensive exams, for his imaginative mind to inspire and motivate mine, and for always being there when productivity and motivation waned. I would lastly and most notably like to recognize my family for their continuous support and strength throughout my unrelenting and seemingly everlasting schooling. Just knowing they were supporting me allowed me to persevere in my studies, stay “on course” and make them proud. Mom, Dad and Crissi-Lyn: thank you so much for everything you have done for me throughout my graduate journey which has taken many, many grueling years. There are no words to express my gratitude. I love you! “In every conceivable manner, the family is the link to our past, and bridge to our future.” – Alex Haley Finally, I would like to thank all the people that have touched my life in one way or another and in particular those I have lost too soon: Dr. Robert Klumb, Dr. David Willis, Alan Zimbrick, Steve Wolf and Anne Kraft. I think of you all often and wish you could be here when I walk across that stage one last time. Each chapter has been formatted for the appropriate journal where it has been submitted or is intended to be submitted for future publication. Chapter 2 is formatted for American Midland Naturalist and co-authors are Katie Bertrand, Michelle Bouchard, Brian Graeb and Jessica Howell. Chapter 3 is formatted for BioInvasions Records and co-authors include Katie Bertrand and Brian Graeb, Chapter 4 has been revised and resubmitted to Aquatic Invasions and co-authors include Katie Bertrand, Jason Breeggemann, Brian Graeb and Robert Klumb. Chapter 5 is formatted for Biological Invasions and co-authors

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include Katie Bertrand, Brian Graeb and Robert Klumb. Funding for this project was provided by the South Dakota Game Fish and Parks, Sport Fish Restoration Project Number F-15-R, and administered through South Dakota State University, Department of Natural Resource Management, formally known as the Department of Wildlife and Fisheries Sciences.

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CONTENTS LIST OF FIGURES ---------------------------------------------------------------------------------- x LIST OF TABLES --------------------------------------------------------------------------------- xiii ABSTRACT ------------------------------------------------------------------------------------------ 1 CHAPTER 1 INTRODUCTION ------------------------------------------------------------------ 3 CHAPTER 2 AN HISTORICAL WATERSHED LEVEL PERSPECTIVE ON THE PERSISTANCE AND OCCUPANCY OF FISHES IN PRAIRIE STREAMS ------- 25 Abstract ----------------------------------------------------------------------------------------------- 25 Introduction ------------------------------------------------------------------------------------------ 26 Study Area-------------------------------------------------------------------------------------------- 30 Methods ----------------------------------------------------------------------------------------------- 32 Results ------------------------------------------------------------------------------------------------ 35 Discussion -------------------------------------------------------------------------------------------- 39 Acknowledgements --------------------------------------------------------------------------------- 48 Literature Cited -------------------------------------------------------------------------------------- 49 CHAPTER 3 DISTRIBUTION OF ASIAN CARPS IN EASTERN SOUTH DAKOTA TRIBUTARIES TO THE MISSOURI RIVER ------------------------------- 85 Introduction ------------------------------------------------------------------------------------------ 85 Methods ----------------------------------------------------------------------------------------------- 86 Results ------------------------------------------------------------------------------------------------ 87 Discussion -------------------------------------------------------------------------------------------- 90 Acknowledgements --------------------------------------------------------------------------------- 91 References -------------------------------------------------------------------------------------------- 92

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CHAPTER 4 POPULATION DYNAMICS OF BIGHEAD AND SILVER CARP ON THE NORTHWESTERN FRONT OF THEIR NORTH AMERICAN INVASION ---------------------------------------------------------------------------------------- 104 Abstract --------------------------------------------------------------------------------------------- 104 Introduction ---------------------------------------------------------------------------------------- 105 Study Area------------------------------------------------------------------------------------------ 110 Methods --------------------------------------------------------------------------------------------- 111 Results ---------------------------------------------------------------------------------------------- 116 Discussion ------------------------------------------------------------------------------------------ 120 Acknowledgements ------------------------------------------------------------------------------- 127 References ------------------------------------------------------------------------------------------ 129 CHAPTER 5 ASSEMBLAGE AND TROPHODYNAMIC STRUCTURE OF THREE PRAIRIE RIVERS DURING THE INITIAL INVASION OF BIGHEAD HYPOPHTHALMICHTHYS NOBILIS AND SILVER CARP H. MOLITRIX------ 158 Abstract --------------------------------------------------------------------------------------------- 158 Introduction ---------------------------------------------------------------------------------------- 160 Methods --------------------------------------------------------------------------------------------- 168 Results ---------------------------------------------------------------------------------------------- 177 Discussion ------------------------------------------------------------------------------------------ 183 Acknowledgements ------------------------------------------------------------------------------- 189 References ------------------------------------------------------------------------------------------ 191 CHAPTER 6 CONCLUSIONS ----------------------------------------------------------------- 238 CURRICULUM VITAE ------------------------------------------------------------------------- 242

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LIST OF FIGURES Figure 1.1. Mean calculated length by age of male and female Bighead Carp collected in the lower Missouri River May to August 1998 and January to May 1999 ------------------ 22 Figure 2.A. 8-digit hydrologic unit (HUs) in the James, Vermillion, and Big Sioux Rivers in eastern South Dakota ---------------------------------------------------------------------------- 65 Figure 2.B. Mean James River basin discharge and standard error by time period ------- 66 Figure 2.C. Nonnative species richness across time periods in the James, Vermillion and Big Sioux rivers in eastern South Dakota-------------------------------------------------------- 67 Figure 2.D. Total species richness for entire, lower and upper river Hus across time periods ------------------------------------------------------------------------------------------------ 68 Figure 2.E. Common species occupancy between tributary and mainstem sites ----------- 69 Figure 2.F. Common species occupancy among hydrologic units for the James, Vermillion and Big Sioux rivers in eastern South Dakota ------------------------------------ 70 Figure 2.G. Nonnative species occupancy among hydrologic units for the James, Vermillion, and Big Sioux rivers in eastern South Dakota ------------------------------------ 71 Figure 3.1. Study area spanning North and South Dakota with the James, Vermillion, and Big Sioux rivers in the eastern part of the states ------------------------------------------------ 96 Figure 3.2. Northern most collections of Silver and Bighead Carp adults ------------------ 97 Figure 3.3. Northern most collections of Silver and Bighead Carp juveniles -------------- 98 Figure 3.4. Juvenile Silver Carp collected in North Dakota on the James River below Jamestown Dam ------------------------------------------------------------------------------------- 99 Figure 3.5. Northern most collections of Silver and Bighead Carp young of year ------ 100

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Figure 3.6. Two young of year Silver Carp collected in Shue Creek, a tributary to the James River ---------------------------------------------------------------------------------------- 101 Figure 4.1. Study area in eastern South Dakota showing ten standardized sites on the James, Vermillion, and Big Sioux rivers------------------------------------------------------- 143 Figure 4.2. Mean annual discharge between 1949 and 2012 with the 25th and 75th percentiles, the median, and the study period marked --------------------------------------- 144 Figure 4.3. Mean monthly discharge downloaded from the nearest US Geological Survey water gage to a sampling site for the James, Vermillion and Big Sioux rivers in eastern South Dakota between 2008 and 2012 --------------------------------------------------------- 145 Figure 4.4. Bighead and Silver Carp catches in South Dakota tributaries across sampling seasons, spring, summer, fall from 2009 - 2012 ---------------------------------------------- 146 Figure 4.5. Mean catch-per-unit effort for Silver Carp between 2009 - 2012 in the James, Vermillion, and Big Sioux rivers, South Dakota --------------------------------------------- 147 Figure 4.6. Length-frequency distribution for Silver Carp across seasons for 2011 and 2012 ------------------------------------------------------------------------------------------------- 148 Figure 4.7. Mean length-at-age and raw data for Silver Carp collected in South Dakota eastern tributaries 2009-2012 -------------------------------------------------------------------- 149 Figure 4.8. Predicted length-at-age for eastern South Dakota tributaries, middle Mississippi River, the Amur River, Russia, and Gobindsagar Reservoir, India --------- 150 Figure 4.9. Age frequency histograms for Silver Carp collected in South Dakota tributaries 2010-2012 ----------------------------------------------------------------------------- 151

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Figure 4.10. Condition of native planktivores, Bigmouth Buffalo, Ictiobus cyprinellus, Gizzard Shad, Dorosoma cepedianum, and Emerald Shiner, Notropis atherinoides versus total length for the James, Vermillion, and Big Sioux Rivers ------------------------------ 152 Figure 4.11. Length-weight regression of Silver Carp collected in South Dakota tributaries (James, Vermillion, Big Sioux rivers), the Missouri River below Gavins Point Dam, the Interior Highlands portion of the Missouri River, and the Illinois River------ 153 Figure 4.12. Mean catch-per-unit effort for native planktivores, Bigmouth Buffalo, Ictiobus cyprinellus, Gizzard Shad, Dorosoma cepedianum, and Emerald Shiner, Notropis atherinoides, across years (2009 - 2012) in South Dakota tributaries---------- 154 Figure 5.1. Study area with sampling sites in the James, Vermillion and Big Sioux rivers, South Dakota --------------------------------------------------------------------------------------- 211 Figure 5.2. Nonmetric multidimensional plot based on Bray-Curtis similarity by study site for all fish species collected, 2009 - 2012 in the James, Vermillion and Big Sioux rivers, South Dakota ------------------------------------------------------------------------------ 212 Figure 5.3. Stable isotope biplot with standard errors for all taxa collected -------------- 213 Figure 5.4. Stable isotope biplot with standard errors for individual native and non-native planktivores: Gizzard Shad, Bigmouth Buffalo, Emerald Shiner, Silver Carp and Bighead Carp ------------------------------------------------------------------------------------------------- 214 Figure 5.5. Total area polygons representing trophic niches in δ15N and δ13C space by year, 2009 – 2012, for the planktivore trophic guild ----------------------------------------- 215 Figure 5.6. Frequency of occurrence versus prey specific abundance of planktivore diets collected from the James, Vermillion and Big Sioux Rivers, 2011 - 2012 --------------- 218

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LIST OF TABLES Table 1.1. Comparison of feeding habits of Asian carps and native planktivores --------- 23 Table 2.1. Number of sampling occasions per Hydrologic Unit and location within the basin for eastern South Dakota -------------------------------------------------------------------- 74 Table 2.2. Persistence values and total mean value across decades with one standard error for the James, Vermillion and Big Sioux river basins between decades -------------------- 75 Table 2.3. Fish species collected, number of sites collected at and proportion of sites collected for the James, Vermillion and Big Sioux rivers since 1934 ----------------------- 76 Table 2.4. Status of missing species, declining species, and species additions to a basin 81 Table 4.1. Mean catch-per-unit effort by year (2009 – 2012) for standardized boat electrofishing on the James, Vermillion, and Big Sioux rivers ----------------------------- 155 Table 4.2. Length-weight regression for silver carp collected in the James, Vermillion, and Big Sioux Rivers in eastern South Dakota along with r2 and 95% upper and lower confidence intervals around the parameter estimates ---------------------------------------- 156 Table 5.1. Annual discharge over the study period, watershed area and distance to Missouri River for ten standardized sites on the James, Vermillion and Big Sioux rivers219 Table 5.2. Mean seasonal water quality variables with associated standard error and minimum and maximum values for the James, Vermillion and Big Sioux Rivers, South Dakota, 2009-2012 -------------------------------------------------------------------------------- 220

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Table 5.3. Morisita’s Index by sampling site for fishes and zooplankton/phytoplankton taxa collected in the James, Vermillion and Big Sioux rivers, South Dakota from 20092012 ------------------------------------------------------------------------------------------------- 222 Table 5.4. Community, guild and species diversity metrics for fish isotope samples collected in the James, Vermillion and Big Sioux rivers, South Dakota, 2009-2012 --- 224 Table 5.5. Dietary overlap (%) between planktivores using δ15N values from fish collected from the James, Vermillion and Big Sioux rivers, South Dakota, 2009-2012 225 Table 5.6. Frequency of occurrence for taxa collected from digestive tracts of planktivores. Fish were collected from the Big Sioux, James and Vermillion Rivers, South Dakota 2011-2012 ------------------------------------------------------------------------- 226 Table 5.7. Schoener’s index of diet overlap for planktivore species collected in the James, Vermillion and Big Sioux Rivers, South Dakota 2011-2012 ------------------------------- 231 Table 5.8. Morisita’s Index of diet similarity for planktivore species collected in the James, Vermillion and Big Sioux Rivers, South Dakota 2011-2012 ---------------------- 232

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APPENDICES Appendix 3.1. First collection records for adult, juvenile, and young-of-year silver and bighead carps for North and South Dakota ---------------------------------------------------- 102 Appendix 5.1. Number of fishes collected, number of fish or samples used for isotope analysis, trophic guild, trophic position and trophic position standard error ------------- 233

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ABSTRACT FISH ASSEMBLAGE STRUCTURE, TROPHIC ECOLOGY, AND POTENTIAL EFFECTS OF INVADING ASIAN CARPS IN THREE MISSOURI RIVER TRIBUTARIES, SOUTH DAKOTA CARI-ANN HAYER 2014 Invasive species are spreading in aquatic systems at an unprecedented rate and have the potential to disrupt the structure and function of these systems. Two invasive species, Bighead Carp and Silver Carp (Asian carps), are spreading in the Mississippi River Basin and more recently into the Missouri River basin and tributaries. Asian carps exhibit fast population growth, are highly fecund, are plankton consumers and are known to cause negative impacts (e.g., decreased condition, decreased growth, altered composition) on native planktivores and plankton. Knowledge of Asian carp population structure and trophodynamics in newly invaded populations is necessary to understand invasion dynamics in general and to predict the impacts Asian carps may be having on the newly invaded community. We found that historically, assemblage structure of eastern South Dakota tributaries to the Missouri River (i.e., James, Vermillion, Big Sioux rivers) was persistent and displayed signs of biotic resistance. Silver carp population abundance increased each year of sampling (e.g., 2009-2012) comprising 45% of catches in 2012; however Bighead Carp did not follow this same pattern as catches remained minimal during this period. Additionally, the population of Silver Carp displayed erratic recruitment in that 91 % of catches were dominated by the 2010 year class. Distribution into eastern South Dakota tributaries was stopped by three substantial dams preventing further natural spread. Isotope analysis and diet analysis revealed that Silver Carp and Gizzard Shad overlapped trophically and the isotopic trophic niche of Emerald Shiner, which is usually pelagic, potentially moved to more benthic habitats and food sources as a result of Silver Carp presence. There were

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no perceived effects of Silver Carp on Bigmouth Buffalo however. With continued population increases of Silver Carp, the potential for competition for plankton resources becomes greater and trophic impacts may become more pronounced, especially for gizzard shad and emerald shiner. Invasive species are usually studied well after an introduction and colonization, unlike this study which began on the cusp of the northwestern United States invasion. This research and continued monitoring of this invasion can help to provide insights into invasion ecology (e.g., competition, phonotypic plasticity, biotic resistance), complex invasive species (e.g., opportunistic and adaptable), and potential impacts (e.g., nutrient cycling, trophic cascades) on the native community and ecosystem.

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CHAPTER ONE: INTRODUCTION Riverscapes reflect the climate, geomorphology and land use of their drainage basins. Stream structure (e.g. width, depth, substrate, riparian zone, woody debris, topography, geology, soil, vegetation) and function (e.g., velocity, discharge) interact to form stream communities that are spatially and temporally variable (Wiens 2002; Haslam 2008; Winemiller et al. 2010). In addition, ecosystem biota are affected differentially by variations in riverscape ecology (e.g. in stream hydrology, invertebrate drift, lateral, longitudinal, and vertical exchanges, patch boundaries, or riverine connectivity with surrounding landscape; Weins 2002). River systems are highly dynamic and understanding how the riverscape influences aquatic community composition (Poff and Allan 1995), distribution and trophic structure will help to understand the current and future status of native aquatic communities (Quist et al. 2004). Lotic systems throughout North America are threatened by increased anthropogenic alterations such as channel modifications, hydrologic alterations, invasive species, land use changes and wetland drainage (Hoagstrom et al. 2006). As a result, the abundance and distribution of native fishes across North America have been declining (Wilcove and Bean 1994; Wilcove et al. 1998; Hoagstrom et al. 2006). Prairie streams are considered one of the most endangered regions in North America (Samson and Knopf 1994; Dodds et al. 2004) and are subject to frequent disturbances caused by climatic or physiochemical extremes (Matthews 1988). Fish in this region are tolerant of harsh natural conditions including extreme hydrological variability (Poff and Ward 1989). Mainstem prairie river environments are variable; however, tributaries can be even more unpredictable (Matthews 1988). Knowledge of the current and historic state of the prairie aquatic community is necessary as little is known about biotic adaptations to harsh,

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fluctuating environments of prairie streams (Matthews 1988). Matthews (1988) suggests that future studies should be devoted to basic ecological description and comparison among prairie streams in order to predict future impacts of anthropogenic alterations including non-native species. Invasive species are common in a majority of North American freshwater environments and are expanding their ranges through natural and human-mediated dispersal (Vitousek et al. 1996; Rahel 2000). The consequences of range expansion are poorly understood for most species and the ecosystems they invade, however they have been linked to the decline of many native fishes (Wilcove and Bean 1994; Wilcove et al. 1998). For example, 68 % of freshwater fishes in the lower 48 United States which have gone extinct since 1890 were negatively affected by nonnative fishes (Wilcove and Bean 1994). Information on the influence of invasive species on community structure and food webs across wide spatial and temporal scales is necessary to identify their impacts on aquatic communities and especially in harsh prairie ecosystems (Lazzaro 1987; Ibanez et al. 2014). Food webs are used to portray and understand predator-prey relationships within ecosystems and can be used to predict how ecosystems respond to disturbance (Vander Zanden et al. 2003) such as altered hydrologic regime or invasive species introduction (Vander Zanden and Rasmussen 1999). Food web analysis (e.g., stable isotopes) can trace and measure alterations in energy flow pathways resulting from anthropogenic ecosystem disturbances and direct ecosystem management efforts (Vander Zanden and Rasmussen 1999; Vander Zanden et al. 2003). Invasive species are major factors of

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global change as they alter not only community composition but also the interactions and processes that occur within ecosystems. Bighead Carp, Hypophthalmichthys nobilis and Silver Carp, H. molitrix (collectively referred to as Asian carps) are invasive species that are spreading rapidly throughout central North America, and most recently in the Missouri River basin. Asian carps were first reported in the lower Missouri River in 1989. Adult Bighead and Silver Carps were first documented in the Missouri River downstream of Gavin’s Point Dam in 1998 and 2003, respectively (Kolar et al. 2007) and have not been found upstream of this barrier (Klumb 2007). Asian carps began invading prairie tributaries of the Missouri River in eastern South Dakota in 2008, coincident with the beginning of this research project on the same systems.

Historic South Dakota tributary distribution A Bighead Carp was caught by an angler from the middle James River (a major tributary to the Missouri River) during summer 2007. The James River experienced extensive flooding throughout much of South Dakota during the summer 2007, enabling Asian carps (and other fishes) to migrate to upstream sites. This is the first account of an Asian carp species at this latitude in the James River, and it demonstrates the urgent need to examine the distribution, abundance and trophic ecology of Asian carps in the Great Plains. Additionally, Silver Carp (Hypophthalmichthys molitrix) have been observed at the confluence of the James River with the Missouri River (Klumb 2007), but the extent of their upstream distribution is unknown. The role that Missouri River tributaries such as

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the James, Vermillion and Big Sioux rivers serve as spawning and nursery habitats and energy sources is unknown.

Asian Carp habitats Adult Bighead Carp are considered a large riverine (e.g. Missouri, Mississippi Rivers) species that utilize low velocity (3m) except during spawning periods when they move to faster flowing habitats (Kolar et al. 2007). Silver Carp occur in numerous habitats including rivers and ponds, lakes, and backwaters connected to large rivers (Berg 1964; Kaul and Rishi 1993; Finley 1999) although they rarely spawn or recruit without access to riverine habitat (Kolar et al. 2007). Silver Carp occur in open waters of standing or slow-flowing waters (Rasmussen 2002) and are found in the middle to upper water column (Shetty et al. 1989). Both species are also reported to use tributaries that have low velocities except during periods of high rainfall when they will move between the river and tributary (Kolar et al. 2007). Data pertaining to juvenile Asian carps are lacking, however, they are reported to use similar habitats as the adults (Kolar et al. 2007). Migrations and movements are thought to be associated with reproductive and feeding behaviors (Kolar et al. 2007). Adults in their native range remain in the main river channel until water levels rise, at which time they migrate upstream to spawn and then move into floodplain lakes (Jennings 1988).

Asian carp life history Asian carps are highly fecund (Schrank and Guy 2002; Williamson and Garvey 2005) and fertility increases with age and body weight (Cudmore and Mandrak 2004;

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Kolar et al. 2007). In the lower Missouri River, fecundity for Bighead Carp ranged from 11,588 – 769, 964, with an average of 226,213 eggs (Schrank and Guy 2002). In the middle Mississippi fecundity of Silver Carp ranged from 57,283-328,538 eggs (Williamson and Garvey 2005). Sexual maturity of Asian carps varies significantly with environmental and climatic conditions. Maturity has ranged from 2-3 years in Southern China (Jennings 1988) to 6-8 years in temperate climates (Woynarovich and Horvath 1980) with males typically maturing 1 year earlier than females (Kuronuma 1968; Jennings 1988). The average size at maturity also is highly variable (Kolar et al. 2007). Spawning and early life history of Asian carps in North American rivers is not well documented; however in Asia, Bighead Carp spawn between April and June (Chang 1966; Verigin et al. 1978) and spawning is initiated by increasing water temperatures and rising water levels following heavy spring rains (Verigin 1979; Pfleiger 1997). Spawning associated with a rise in water levels (Schrank et al. 2001; Lohmeyer et al. 2007) may decrease egg mortality and increase the chances of larvae to enter energy rich floodwaters (Krykhtin and Gorbach 1981). Spawning habitats in their native range are characterized by high flowing (0.6-2.3 m/s) turbid water between 18 and 30°C (Verigin et al. 1978; Schrank et al. 2001) and are typically located where there is a mixing of water (i.e. confluences; Peters et al. 2006; Huet 1970). In the lower Missouri River, Asian carps have protracted spawning periods extending from early spring through fall (Galat et al. 2004a, b) with some spawning multiple times (Pfleiger 1997; Papoulias et al. 2006). Asian carps have rapid growth rates and are capable of reaching 18-23 kg in 4-5 years (Henderson 1978; Leventer 1987). In the Gavin’s Point reach of the Missouri

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River, Bighead Carp ranged from 322-1200 mm and 4 -19 kg (Wanner and Klumb 2009). They are also long lived reaching 8-10 years old in Lake Erie, Ontario (Johal et al. 2001; Morrison et al. 2004) and 7 years in the lower Missouri River (Schrank and Guy 2002). Information on age and growth of Asian carps is minimal because ageing this species is difficult (Schrank and Guy 2002; Kolar et al. 2007; Figure 1).

Asian Carps feeding habits Bighead Carp are filter feeders and are predominantly zooplanktivorous (Table 1). Larvae eat primarily protozoa and zooplankton, including rotifers, cladocerans and copepod nauplii. Larval, juvenile and adult Bighead Carp exhibit highly opportunistic feeding habits that are dependent on zooplankton biomass (Kolar et al. 2007). When zooplankton biomass is low, studies have shown that Bighead Carp switch to feed on phytoplankton (e.g. blue-green algae, diatoms, and green algae; Kolar et al. 2007). Silver Carp are also filter feeders and are predominantly phytoplanktivorous and consume smaller particles than Bighead Carp. Larvae feed on small zooplankton species and will switch to phytoplankton at around 45 mm (Marciak and Bogdan 1979). When phytoplankton biomass is low, studies have shown that Silver Carp also consume zooplankton (Spataru and Gophen 1985; Burke et al. 1986; Williamson and Garvey 2005).

Environmental effects Asian carps can alter water quality but the effects are inconsistent (Kolar et al. 2007). Silver Carp presence has increased, decreased, or not changed nutrient

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concentrations (Opuszynski 1980; Starling 1993; Matyas et al. 2003), decreased dissolved oxygen in water (Lambi et al. 1978; Vybornov 1989), and increased total nitrogen (Starling 1993). Silver Carps can increase turbidity leading to increased small algae production (Rowe 1972; Fedorenko and Fraser 1978; Lembi et al. 1978), and can effect sight feeding predators and reduce growth of aquatic macrophytes (Vogel and Beauchamp 1999; Radke and Kahl 2002; Kolar et al. 2007). Asian carps can have negative effects on the plankton community, benthic macroinvertebrates and native fishes by altering food webs through foraging, predation and competition (Kolar et al. 2007). Interspecific competition between Bighead and Silver Carps and native planktivores (e.g. Bigmouth Buffalo, Ictiobus cyprinellus, Gizzard Shad, Dorosoma cepedianum, Paddlefish, Polyodon spathula) has been documented (Schrank et al. 2003; Irons et al. 2007; Sampson et al. 2009). For example, Gizzard Shad and Bigmouth Buffalo body condition declined after Asian carp invasion in the Illinois River (Irons et al. 2007) and in another study, there was dietary overlap between Bigmouth Buffalo and Asian carps (Samson et al. 2009). Asian carps invasion in the middle Missouri River and tributaries could potentially alter the aquatic community. Asian carps have typical invasive characteristics in that they are long-lived, fast growing, tolerate a broad range of environmental conditions, have opportunistic feeding habits, have high reproductive capabilities and can reach high population densities (Kolar et al. 2007; Wanner and Klumb 2009). In the face of global change (e.g., invasive species), it is important to document the current and historic state of aquatic communities to understand invasive species effects on the aquatic ecosystem. This study sought to quantify the distribution, population dynamics and

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trophic ecology of Asian carps in three Missouri River tributaries: the Big Sioux, James and Vermillion rivers. I hypothesized that Asian carps will affect the native riverine ecosystem through direct competition (i.e., niche overlap with other planktivores) and may indirectly affect food web structure (e.g., piscivores) through cascading or middleout effects. The specific objectives are to: 1) characterize present and historic patterns of fish assemblage structure, 2) track the spatiotemporal distribution of Asian carps into prairie tributaries 3) quantify the population dynamics of Asian carps and other planktivores in Great Plains streams and 4) track the energy flow and feeding relations in prairie stream aquatic food webs. I conducted a field study that selected ten standardized sites across the James, Vermillion and Big Sioux Rivers in eastern South Dakota to account for variability associated with fish movement and distribution. Fish sampling was standardized at each of the ten sites and was conducted across three seasons (e.g., spring, May-June; summer, July-August; fall, September-October) to account for temporal variability in fish distribution. Asian carps are difficult to capture (Stancil 2003; Williamson and Garvey 2005; Conover et al. 2007; Klumb 2007; Wanner and Klumb 2009) even in areas with high observed carp concentrations. As a result, numerous active (e.g., boat electrofishing, backpack electrofishing) and passive (i.e., hoop nets, mini-fyke nets, trammel nets) gears were used to sample Asian carps. My goal was to maximize catch rates and adequately sample all size classes. A historical picture of the assemblage structure within the study area was examined in order to determine any effects Asian carps might have. A fish database

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dating back to 1934, was used to characterize historical patterns in fish assemblage structure among the James, Vermillion and Big Sioux river basins. Fish species richness, persistence and occupancy were calculated and compared across time periods and spatial units (e.g., hydrologic unit) within each basin. The most upstream location of Bighead Carp and Silver Carp adults, juveniles and young of year was recorded to document the northwestern United States distribution front. Population dynamics (e.g., growth, age) were calculated for Asian carps. Dynamic rate variables for Bighead and Silver Carp were compared among seasons and rivers. Finally, condition of native planktivores was analyzed to examine for any potential negative impacts resulting from Asian carps presence. Energy flow and feeding relationships in aquatic food webs was tracked using stable isotope analysis (SIA) and diet analysis. Stable isotope analysis is an important tool in the study of aquatic food webs as it helps identify pathways of nutrient transfers and can be used to estimate assimilation of food resources over time (Peterson and Fry 1987). Stable carbon isotopes (δ13C) were selected to assess energy sources and stable nitrogen isotopes (δ15N) were selected to measure the trophic level of each taxa in the aquatic community (e.g., fish, phytoplankton, plants, crayfish, invertebrates). Mass balance spectrometry was performed on all samples to determine δ13C and δ15N ratios using a continuous-flow isotope ratio mass spectrometer (Europa Scientific Ltd, United Kingdom) with a dual-ion collection beam for 13C/12C and 15N/14N isotope ratio measurements. Isotopic results for carbon and nitrogen were quantified as changes relative to isotopic standards. Community, guild, and species isotope metrics were calculated (e.g., nitrogen range, carbon range, total niche area) and examined for

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spatiotemporal patterns. Additionally, we quantified the frequency of occurrence, prey specific abundance, and the Manly-Chesson Index to evaluate diet composition and prey selection. Diet analysis was then related to isotope measures and diet overlap (e.g., Schoeners Index) among native and nonnative planktivores was also calculated.

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REFERENCES Berg, L. S. 1964. Freshwater fishes of the USSR and adjacent countries. 4th edition, Volume II. The Smithsonian Institution and the National Science Foundation, Washington D.C. Burke, J. S., D. R. Bayne, and H. Rea. 1986. Impact of silver and bighead carps on plankton communities of channel catfish ponds. Aquaculture 55:59-68. Chang, Y. F. 1966. Culture of freshwater fish in China. In E.O. Gandstad, editor. 1080. Chinese fish culture. Report 1. Technical report A-79. Aquatic plant control research program. U.S. Army Waterways Experiment Station, Washington, D.C. Conover, G., R. Simmonds and M. Whalen. 2007. Management and control plan for Bighead, Black, Grass and Silver Carps in the United States, Washington, DC. Cudmore, B. and N. E. Mandrak. 2004. Biological synopsis of grass carp (Ctenopharyngodon idella). Canadian Manuscript Report of Fisheries and Aquatic Sciences 2705(November):1-44, II, III-V. Dodds, W. K., K. Gido, M. R. Whiles, K. M. Fritz and W. J. Matthews. 2004. Life on the edge: The ecology of Great Plains prairie streams. BioScience 54:205-216. Finley, J. 1999. Biological summery of silver carp. Included in handout materials. Proceedings: Asian carp management and control workshop, April 19-20, St. Louis, Missouri. U.S. Fish and Wildlife Service, Columbia Fishes Resources Office, Columbia, Missouri. Galat, D. L., G. W. Whitledge and G. T. Gelwicks. 2004a. Influence of lateral connectivity on larval fish assemblage structure and habitat use in lower Missouri River floodplain water bodies. Final Report to Missouri Department of

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Conservation. U.S. Geological Survey, Missouri Cooperative Fish and Wildlife Research Unit, University of Missouri, Columbia. Galat, D. L., G. W. Whitledge, L. D. Patton, J. Hooker. 2004b. Larval fish use of lower Missouri River scour basins in relation to connectivity. Final Report to Missouri Department of Conservation. U.S. Geological Survey, Missouri Cooperative Fish and Wildlife Research Unit, University of Missouri, Columbia. Haslam, S. M. 2008. The riverscape and the river. Cambridge University Press, New York. Henderson, S. 1978. An evaluation of the filter feeding fishes, silver and bighead carp, for water quality improvement. Pages 121-136 in R.O. Smitherman, W. L. Shelton and J. H Grover, editors. Culture of exotic fishes symposium proceedings. American Fisheries Society, Fish Culture Section, Auburn, Alabama. Hoagstrom, C. W., C.-A. Hayer, J. G. Kral, S. S. Wall and C. R. Berry. 2006. Rare and declining fishes of South Dakota: a river drainage scale perspective. Proceedings of the South Dakota Academy of Science 85:171-211. Huet, M. 1970. Textbook of fish culture: breeding and cultivation of fish. Fishing News Ltd., Surrey and London. Ibanez I, Diez J. M. L. P. Miller, J. D. Olden, C. J. B. Sorte, D. M. Blumenthal, B. A. Bradley, C. M. D’Antonio, J. S. Dukes, R. I. Early, E. D. Grosholz, J. J. Lawler 2014. Integrated assessment of biological invasions. Ecological Applications 24:25-37. doi:10.1890/13-0776.1 Irons, K. S., G. G. Sass, M. A. McClelland and J. D. Stafford. 2007. Reduced condition factor of two native fish species coincident with invasion of non-native Asian

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carps in the Illinois River, USA - Is this evidence for competition and reduced fitness? Journal of Fish Biology 71:258-273. Jennings, D.P. 1988. Bighead Carp (Hypophthalmichthys nobilis): a biological synopsis. U.S. Fish and Wildlife Service, Washington, D.C. U.S. Fish and Wildlife Service Biological Report 88:1-47. Johal, M. S., H. R. Esmaeili and K. K. Tandon. 2001. A comparison of back-calculated lengths of silver carp derived from bony structures. Journal of Fish Biology 59:1483-1493. Kaul, M., and K. K. Rishi. 1993. Exotic Chinese carps. Punjab Fisheries Bulletin 27:5356. Klumb, R. A. 2007. Shallow water fish communities in the Missouri River downstream of Fort Randall and Gavins Point dams in 2003 and 2004 with emphasis on Asian Carps, Pierre South Dakota. Kolar, C. S., D. C. Chapman, W.R. Courtenay Jr., C.M. Housel, J.D. Williams, and D.P. Jennings. 2007. Bigheaded Carps: a biological synopsis and environmental risk assessment. American Fisheries Society Special Publication 33, Bethesda, Maryland. Krykhtin, M. L. and E. I. Gorbach. 1981. Reproductive ecology of the grass carp, Ctenopharyngodon idella, and the silver carp, Hypophthalmichthys molitrix, in the Amur Basin. Journal of Ichthyology 21:109-123. Kuronuma, K. 1968. New systems and new fishes for culture in the Far East. FAO Fisheries Report 5:123.

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Lazarro X. 1987. A review of planktivorous fishes: their evolution, feeding behaviors, selectivities and impacts. Hydrobiologia 146:97-167. doi:10.1007/BF00008764 Leventer, H. 1987. Contribution of silver carp (Hypophthalmichthys molitrix) to the biological control of reservoirs. Mekorth Water Company, Jordan District, Nazareth, Israel. Lohmeyer A. M. and J. E. Garvey 2009. Placing the North American invasion of Asian carp in a spatially explicit context. Biological Invasions 11:905-916. doi:10.1007/s10530-008-9303-5 Marciak, Z. and E. Bogdan. 1979. Food requirements of juvenile stages of grass carp, Ctenopharyngodon idella Val., silver carp, Hypophthalmichthys molitrix Val. and bullhead carp Aristichthys nobilis Rich. EMS Special Publication 4:139-157. Matthews, W. J. 1988. North American prairie streams as systems for ecological study. Journal of North American Benthological Society 7:387-409. Matyas, K., I. Oldal, J. Korponai, T. Tatrai and G. Paulovitsl. 2003. Indirect effect of different fish communities on nutrient chlorophyll relationship in shallow hypereurtrophic water quality reservoirs. Hydrobiologia 504:231-239. Morrison, B. J., J. C. Casselman, T. B. Johnson and D. L. Noakes. 2004. New Asian carp genus (Hypophthalmichthys) in Lake Erie. Fisheries 29:6-7. Opuszynski, K. 1980. The role of fishery management in counteracting eutrophication processes. Pages 263-269 in J. Barica and L. R. Mur, editors. Hypertrophic ecosystems. Developments in hydrobiology 2. Dr. W. Junk Publishers, The Hague, The Netherlands.

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Papoulias, D. M., D. Chapman, and D. E. Tillitt. 2006. Reproductive condition and occurrence of intersex in bighead carp and silver carp in the Missouri River. Hydrobiologia 571:355-360. Peters, L. M., M. A. Pegg, and U. G. Reinhardt. 2006. Movements of adult radio-tagged bighead carp in the Illinois River. Transactions of the American Fisheries Society 135:1205-1212. Peterson, B. J. and B. Fry. 1987. Stable isotopes in ecosystem studies. Annual Reviews in Ecological Systematics 18:293-320. Pflieger, W. L. 1997. The fishes of Missouri, 2nd edition, Missouri Department of Conservation, Jefferson City. Poff, N. L. and J. D. Allan. 1995. Functional organization of stream fish assemblages in relation to hydrological variability. Ecology 76:606-627. Poff, N. L. and J. V. Ward. 1989. Implications of streamflow variability and predictability for lotic community structure: a regional analysis of streamflow patterns. Canadian Journal of Fisheries and Aquatic Sciences 46:1805-1818. Quist, M. C., W. A. Hubert, and F. J. Rahel. 2004. Relations among habitat characteristics, exotic species and turbid-river cyprinids in the Missouri River drainage of Wyoming. Transactions of the American Fisheries Society 133:727742. Radke R. J. and U. Kahl. 2002. Effects of filter-feeding fish [silver carp, Hypophthalmichthys molitrix (Val.)] on phyto- and zooplankton in a mesotrophic reservoir: results from an enclosure experiment. Freshwater Biology 47:23372344.

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Rahel, F. J. 2000. Homogenization of fish faunas across the United States. Science 288:854-856. Rasmussen, J. L. 2002. The Cal-Sag and Chicago Sanitary and Ship Canal: a perspective on the spread and control of selected aquatic nuisance fish species. U.S. Fish and Wildlife Service, Fisheries Resources Office, Rock Island, Illinois. Samson, F., and F. Knopf. 1994. Prairie conservation in North America. BioScience 44:418-421. Sampson, S. J., J. H. Chick and M. A. Pegg. 2009. Diet overlap among two Asian carp and three native fishes in backwater lakes on the Illinois and Mississippi rivers. Biological Invasions 11:483-496. Schrank, S. J., P. J. Braaten, and C. S. Guy. 2001. Spatiotemporal variation in density of larval bighead carp in the lower Missouri River. Transactions of the American Fisheries Society 130:809-814. Schrank, S. J. and C. S. Guy. 2002. Age, growth and gonadal characteristics of adult bighead carp, Hypophthalmichthys nobilis, in the lower Missouri River. Environmental Biology of Fishes 64:443-450. Schrank, S. J., C. S. Guy and J. F. Fairchild. 2003. Competitive interactions between age0 Bighead Carp and Paddlefish. Transactions of the American Fisheries Society 132:1222-1228. Shetty, H.P., M.C. Nandeesha, and A.G. Jhingran. 1989. Impact of exotic aquatic species in Indian waters. Pages 45-55 in S.S. De Silva, editor. Exotic aquatic organisms in Asia. Proceedings of the workship on introduction of exotic aquatic organisms in Asia. Asian Fisheries Society Special Publication 3, Makati City, Philippines.

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Spataru, P. and M. Gophen. 1985. Feeding behavior of silver carp Hypophthalymichthys molitrix Val. And its impact on the food web in Lake Kinneret, Israel. Hydrobiologia 120:53-61. Stancil, W. 2003. An evaluation of sampling techniques and life history information on Bighead Carp in the Missouri River, below Gavins Point Dam, South Dakota and Nebraska, Pierre, SD. Starling, F.L.R.M. 1993. Control of eutrophication by silver carp (Hypophthalmichthys molitrix) in the trophical Paronoa reservoir (Brasilia, Brazil): a mesocosm experiment. Hydrobiologia 27:143-152. Vander Zanden, M. J. and J. B. Rasmussen. 1999. Primary consumer delta C-13 and delta N-15 and the trophic position of aquatic consumers. Ecology 80:1395-1404. Vander Zanden, M. J., S. Chandra, B. C. Allen, J. E. Reuter and C. R. Goldman. 2003. Historical food web structure and restoration of native aquatic communities in the Lake Tahoe (California-Nevada) Basin. Ecosystems 6:274-288. Verigin, B.V, A.P. Makeyeva, and M.I. Zaki Mokhamed. 1978. Natural spawning of the silver carp (Hypophthalmichtys nobilis, the bighead carp (Aristichthys nobilis), and the grass carp (Ctenopharyngodon idella) in the Syr-Da’ya River. Journal of Ichthyology 18:143-146. Verigin, B.V. 1979. The role of herbivorous fishes at reconstruction of ichthyofauna under the condition of anthropogenic evolution of water bodies. Pages 135-145 in J.V. Shireman, editor. Proceedings of the grass carp conference, Gainseville, Florida, Aquatic Weeds Research Center, Institute of Food and Agricultural Sciences, University of Florida, Gainsvelle.

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Vitousek, P. M., C. M. Dantonio, L. L. Loope and R. Westbrooks. 1996. Biological invasions as global environmental change. American Scientist 84:468-478. Vogel, J. L. and D. A. Beauchamp. 1999. Effects of light, prey size, and turbidity on reaction distances of lake trout (Salvelinus namaycush) to salmonid prey. Canadian Journal of Fisheries and Aquatic Sciences 56:1293-1297. Vybornov, A. A. 1989. Effects of silver carp, Hypophthalmichthys molitrix, on production indices of phyto- and zooplankton under experimental conditions. Journal of Ichthyology 29:136-140. Wanner, G. A. and R. A. Klumb. 2009. Asian Carp in the Missouri River: analysis from multiple Missouri River habitat and fisheries programs. US Fish and Wildlife Service, Pierre, SD. Wiens, J. A. 2002. Predicting species occurrences: progress, problems and prospects. Pages 739-749 in J. M. Scott, and coeditors, editors. Predicting Species Occcurrences. Issues of Accuracy and Scale. Island Press, Washington D.C. Wilcove D. S., S. Rothstein, J. Dubow, A. Phillips, and E. Losos 1998. Quantifying threats to imperiled species in the United States. BioScience 48:607-615. Wilcove D. S. and M. J. Bean 1994. The Big Kill: Declining Biodiversity in America’s Lakes and Rivers. Environmental Defense Fund, Washington, D.C. 275 pp Williamson, C. J. and J. E. Garvey. 2005. Growth, fecundity and diets of newly established Silver Carp in the middle Mississippi River. Transactions of the American Fisheries Society 134:1423-1430.

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Winemiller, K. O., A. S. Flecker, and D. J. Hoeinghaus. 2010. Patch dynamics and environmental heterogeneity in lotic ecosystems. Journal of the North American Benthological Society 29:84-99. Woynarovich, E. and L. Horvath. 1980. The artificial propagation of warm-water finfishes- a manual for extension. FAO Fisheries Technical Paper 201.

Mean-back calculated length (mm)

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1000 900 800 700 600 500 400 300 200 100 0 1

2

3

4

5

6

7

Age

Figure 1.1 Mean back-calculated length (mm) by age of male (squares) and female (triagles) Bighead Carp collected in the lower Missouri River from May to August 1998 and January to May 1999. Circles represent mean back-calculated length of Bighead Carp stocked in Poland lakes. Taken from Schrank and Guy (2002).

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Table 1.1: Comparison of feeding habits of Asian Carps and native planktivores

Type of feeder Food items consumed

Bighead Carp Opportunistic planktivore

Silver Carp Planktivore, but opportunistic

Bigmouth Buffalo

Zooplankton; phytoplankton

Phytoplankton, zooplankton, detritus

Indiscriminant plankton

Morphological characteristics specific to feeding

Long, comb-like gill rakers coated with mucus. Many taste buds to aid detection of zooplankton

Consumption rate

High; voracious feeder

Feeding temperatures

Most active at 20-22°C Often at surface, but also feed throughout the water column

Ecological niche for feeding

Special filtering apparatus on gill bars allowing removal of small particles. Suprabranchial organ consolidates ingested materials by producing large amounts of mucus High, but widely variable Most active at 15-30°C Do not commonly feed directly on the surface

Planktivore

Small mouth and pharyngeal cavities for handling minute particles of food

Gizzard Shad Filter feeding phytoplanktivore Phytoplankton – organic detritus, algae, diatoms

Emerald Shiner Planktivore zooplankton

Long and numerous gill rakers

High

Bottom to near bottom

Mid column to bottom mostly, at surface sometimes

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CHAPTER TWO AN HISTORICAL WATERSHED LEVEL PERSPECTIVE ON THE PERSISTANCE AND OCCUPANCY OF FISHES IN PRAIRIE STREAMS Abstract.‒Prairies are one of the most endangered regions in North America. Though fishes in these regions are adapted to naturally variable conditions, their response to increased anthropogenic alterations such as hydrologic alterations, introduced species, and land-use changes is relatively unknown. Where anthropogenic disturbances are becoming more frequent, such as in South Dakota, it is important to document spatial and temporal trends in the environmental history of fish assemblages to better assess future impacts on the ecosystem. The objective of this study was to examine spatial and temporal patterns in fish assemblage structure (e.g., species richness, occupancy and persistence) in three eastern South Dakota prairie river basins (Big Sioux, James, Vermillion). Using a historical presence-absence dataset spanning over 70 years, we identified seventy-seven species across the study area with some species unique to each river basin. Persistence (P) was high (P > 0.75) in the James and Big Sioux rivers and moderate (P = 0.53) in the Vermillion River, indicating few colonization or extinction events. Common species exhibited 100% persistence across decades in the Big Sioux and James River basins and 89% persistence in the Vermillion River, likely due to their high tolerance attributes. Seventeen species were considered extirpated from the study area, whereas forty species were first detected after 1990. These recent detections can be attributed to many factors, including range expansion (southern redbelly dace) and stocking practices (e.g., northern pike, largemouth bass, and white bass). Two non-native species that are considered nuisance species, common carp and grass carp, also were

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collected in recent surveys, but have not negatively affected fish assemblages. The most recent invaders, bighead carp and silver carps (detected after this database was compiled) also are nuisance species with the potential to restructure aquatic ecosystems and food webs. Despite the dynamic nature of the streams and anthropogenic alterations within the watersheds, the fish assemblages within this study area appear to be persistent across a broad timescale, characteristic of the biotic resistance hypothesis. New invaders will test the resistance of this system to yet another disturbance. As land-use alterations continue to increase on a global and local scale, and as introduced and invasive species continue to increase dramatically in novel systems, it is important to establish historical patterns in baseline data to document potential current and future impacts of these disturbances.

Keywords.‒ Prairie streams, fish assemblage structure, historical data, occupancy, species richness, persistence, anthropogenic modifications, nonnative species, South Dakota

Introduction There is a worldwide conversion of land from undisturbed natural states to disturbed unnatural states, which has impacted river ecosystems worldwide (Allan, 2004). As such, agriculture dominates in many developed watersheds throughout the world (Allan, 2004) covering approximately 40% of the world’s land surface (Ramankutty and Foley, 1999). In North America, agricultural land use was 66% in the Upper Mississippi basin (Benke and Cusing, 2004) and was as high as 95% in the Minnesota River basin (Allan, 2004). Additionally, crop production has increased since

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the 1940’s in the Great Plains (Parton et al., 2007). Lotic ecosystems are strongly influenced by the surrounding watershed at multiple spatial and temporal scales (e.g., Allan, 2004; Gido et al., 2010; Mullen et al., 2011). Agricultural land can affect stream integrity (Allan et al., 1997) through declines in water quality, habitat, and biological assemblages (Roth et al., 1996; Allan et al., 1997; Foley et al., 2005). The primary mechanisms by which land use and in particular agricultural land use can influence streams are through increased sedimentation, nutrient enrichment, contaminant pollution, hydrologic alteration, riparian reduction, and loss of large woody debris (Allan, 2004). Matthews (1988) suggested that future studies be devoted to basic ecological description and comparison among streams, and in particular, prairie streams, to further assess future impacts on the aquatic assemblage resulting from anthropogenic alterations, such as land use change and also a more recent threat, nonnative or invasive species. Nonnative and invasive species are now common in a majority of North American freshwater environments and are expanding their ranges through natural and humanmediated dispersal (Vitousek et al., 1996; Rahel, 2000). Nonnative species are the second leading cause of biodiversity loss in the United States behind habitat degradation (agricultural non-point pollution; Richter et al., 1997; Wilcove et al., 1998). Invasive species have the ability to tolerate a wide range of environmental conditions (Ricciardi and Rasmussen, 1998; Scott and Helfman, 2001; Kolar and Lodge, 2002) and can affect native communities through direct and indirect biotic interactions (Vitousek et al., 1996; Mack et al., 2000; Bruno et al., 2005), and habitat degradation (Bain, 1993). The impacts of invasive species on native fishes and the ecosystems they invade are poorly understood

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(Ricciardi and Rasmussen, 1998; Sakai et al., 2001; Bruno et al., 2005), especially in such dynamic prairie communities as those in eastern South Dakota. Prairie stream habitats are considered to be within one of the most endangered regions in North America (Samson and Knopf, 1994; Dodds et al., 2004) and are subject to frequent disturbances caused by stress from extreme climatic and physiochemical conditions (Matthews, 1988). Fish in this region are evolutionarily adapted to these harsh natural conditions (Matthews, 1987; Poff and Ward, 1989; Matthews and MarshMatthews, 2003). However, their tolerance and response to increased anthropogenic alterations such as channel modifications (e.g., channelization and riparian degradation), hydrologic alterations (e.g., dams and water withdrawal), introduced species (e.g., Asian carps), land use changes (e.g., increased agriculture production, cattle grazing), and wetland drainage (Matthews, 1988; Rabeni, 1996; Henley et al., 2000; Hoagstrom et al., 2006; Hoagstrom et al., 2007b) is relatively unknown (Matthews, 1988) as impacts can occur on multiple spatial and temporal scales (Allan, 2004). Due to physiochemical and biological modifications of aquatic habitat (Quist et al., 2004; Heitke et al., 2006; Fischer and Paukert, 2008), the abundance and distribution of native fishes in prairie rivers has been declining (Dodds et al., 2004; Oakes et al., 2005; Hoagstrom et al., 2006). Information on the current state of prairie river assemblages and examination of historical spatial and temporal patterns is necessary (Oakes et al., 2005; Mullen et al., 2011) as little is known of the biotic adaptations of fishes to natural, harsh, fluctuating environments (Matthews, 1988; Dodds et al., 2004) and how these fishes will respond to more recent anthropogenic threats, such as continued land use alterations and introduction of nonnative species.

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In the face of increased anthropogenic disturbance, it is important to document spatial and temporal trends (Oakes et al., 2005; Mullen et al., 2011) in the environmental history (e.g., persistence) of aquatic communities to better assess future impacts (e.g., land use changes. invasive species) on the ecosystem (Mooney and Cleland, 2001). The objective of this study was to examine spatial and temporal patterns in fish assemblage structure (e.g., species richness, occupancy, and persistence) in three eastern South Dakota prairie river basins (Big Sioux, James, and Vermillion rivers) using a historical presence-absence dataset spanning over 70 years. Several hypotheses were developed about the fish assemblages in the Big Sioux, James, and Vermillion rivers. We hypothesized that the lower reaches of the each river would have the most similar assemblages and the highest species richness because habitat structure is generally heterogeneous at tributary confluences (Benda et al., 2004) and downstream reaches (Vannote et al., 1980). These reaches would also be highly influenced by the Missouri River in that these assemblages would contain many big river species (e.g., paddlefish, Polyodon spathula and blue sucker, and Cycleptus elongatus). Because fish in prairie streams are adapted to harsh environmental conditions, we believe that there would be minimal change in species richness and occupancy across time periods; however, when taking into account anthropogenic alterations such as channelization, agricultural practices and wetland drainage, we believe that habitat specialists (e.g., goldeye, Hiodon alosoides, blue sucker, flathead catfish, Pylodictis olivaris) and intolerant species (e.g., shovelnose sturgeon, Scaphirhynchus platorynchus, and blue sucker) may be negatively affected (Eitzmann and Paukert, 2010).

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Study Area Three major tributaries in South Dakota converge with an unchannelized section (Galat et al., 2005a) of the Missouri River just downstream of Gavins Point Dam: the Big Sioux, James, and Vermillion rivers. These warmwater rivers drain the Central Lowlands physiographic province in South Dakota (Galat et al., 2005b). This region is characterized by low relief with abundant lakes, wetlands, and low-gradient streams of glacial origin (Hoagstrom et al., 2007b). The study area is also characterized by alternating wet years of prolonged flooding and high discharge, and dry years of extended drought and intermittent flows (Shearer and Berry, 2003). The Big Sioux River (basin area = 23,325 km2) extends 470 Rkm from the Prairie Coteau of northeastern South Dakota to its confluence with the Missouri River at the South Dakota-Nebraska-Iowa border near Sioux City, Iowa (Figure A). The river contains three low-head dams (2-5m high) and a natural barrier (a series of waterfalls in Sioux Falls, South Dakota) that impedes fish movements (Galat et al., 2005a). Land use within this region is comprised of 62% agriculture, 27% grassland/pasture, 3% forest shrub, 7% wetlands/river and 1% urban (Galat et al., 2005a). In the early 1960’s, this river was considered to be one of the most polluted rivers in the United States due to municipal waste and agricultural non-point source pollution (USEPA, 1978; Galat et al., 2005b), though the passing of the Clean Water Act in 1972 eventually improved water quality (Dieterman and Berry, 1998). The river is relatively natural as only 3% has been altered and only 30% of wetlands have been drained (Galat et al., 2005a). The James River (basin area = 57,000 km2) extends 760 Rkm from southeastern North Dakota through eastern South Dakota to its confluence with the Missouri River

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near Yankton, South Dakota (Berry et al., 1993; Figure A). The upper portion has a gradient of about 0.02 m/km, and the lower portion has a gradient of about 0.05 m/km (Owen et al., 1981), making this river one of the lowest gradient rivers in the United States (Benson, 1983). The dominant land use in the basin is row-crop agriculture that covers the uplands, and livestock grazing (Owen et al. 1981) that is directly adjacent to the streams in this basin (Wall and Berry, 2006). There are approximately 230 low-head dams and rock crossings (Shearer and Berry, 2003) that potentially limit fish migration and reproductive success during low hydrologic periods. However, most of these barriers are passable during flood events (Berry et al., 1993). The Vermillion River, the smallest basin in this study (basin area = 5,800 km2), extends 243 Rkm from the confluence of West and East Fork Vermillion rivers to its confluence with the Missouri River near Vermillion, South Dakota (Schmulbach and Braaten, 1993; Figure A). The basin topography is characterized by undulating uplands and hills bordering the stream valleys, but the river channel has been straightened, dredged, and restrained by levees (Schmulbach and Braaten, 1993; Johnson 1997). However, few tributaries have been impounded (Johnson and Higgins, 1997). There are two periods of flooding (spring snowmelt and early summer associated with heavy rainfall) in the annual hydrologic cycle (Schmulbach and Braaten, 1993). Flooding in this basin is the result of low stream gradient and high storage potential of the valley and is also influenced by the water levels of the Missouri River (Johnson, 1997). Many wetlands have been drained within the basin with a loss of 6% during a five year period in the 1980’s, which contributes to low flows in late summer and drought years (Johnson, 1997). Seventy to 90% of the watershed is dominated by row crop agriculture, replacing

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the historical tall and mixed grass prairies characteristic of this region (Johnson, 1997). Channelization and sedimentation have caused degradation of water quality in this basin (Underhill, 1959; Johnson, 1997). For spatial comparisons, each river basin (i.e., James, Vermillion and Big Sioux rivers) was further divided into 8 digit hydrologic units (HUs) derived from Seaber et al. (1987; Figure A). Hydrologic units represent standard geographic and hydrologic boundaries that correspond to a principal hydrologic feature within the unit (Seaber et al., 1987). Additionally we examined patterns between tributary and mainstem systems.

Methods We calculated the percentage of land that was devoted to cropland (i.e., corn, soy, wheat, hay, small grains) within our hydrologic units between 1850 and 2007 using datasets and the methods described by Gutman (2005). Small grains consist of oats, barley, flaxseed and rye. We examined patterns in fish assemblage structure of eastern South Dakota tributaries to the Missouri River using historical (1934-2005) survey data (Smith et al., 2002; Hoagstrom et al., 2007a). Early surveys date back to 1934; however, the first comprehensive inventory was compiled by Bailey and Allum (1962). The database was divided into five time intervals (pre 1970, 1970-1979, 1980-1989, 19901999, 2000-2005) to encapsulate environmental variability (e.g., flood, drought, and dams; Figure 2) and account for non-standardized sampling (N = 607 sampling occasions; Table 1). Early surveys were conducted with sodium cyanide and electric seining, changing to a combination of backpack electrofishing, hoop netting, fyke netting, and seining in more recent surveys (Sinning, 1968; Schmulbach and Braaten, 1993;

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Dieterman and Berry, 1998). To offset the bias associated with gear capture efficiency, size selectivity, and sampling effort, we focused on species presence-absence data (Schmulbach and Braaten, 1993), as interpretation of abundance data can be confounded by the extreme variability (Patton et al., 1998) of prairie streams. Our main focus was to assess spatial and temporal changes in assemblage composition (e.g., persistence), so presence-absence data were appropriate (Connel and Sousa, 1983; Schmulbach and Braaten, 1993; Patton et al., 1998). Persistence is a qualitative measure of continued presence of a species, assemblage or group of species (Ross et al., 1985; Matthews, 1986; Matthews, 1988) over time and it can measure extinction and colonization rates relative to how a system responds to high disturbance periods (e.g., drought and flood; Connel and Sousa, 1983). In addition, persistence of populations can be influenced by deterministic processes (e.g., local biotic interactions, microhabitat availability, flow regime; Grossman, 1982; Ricklefs, 1987; Poff, 1997; Nui et al., 2012). Deterministic processes (e.g., spatial position and fine substrates) explain more variation in species richness and assemblage structure than stochastic processes (e.g., hydrologic cycle) in prairie streams (Mullen et al., 2011; Matthews et al., 2013). Assemblage structure was evaluated for spatial (i.e., river basin, HU, and tributary vs mainstem) and temporal (across time intervals) patterns by calculating species richness, persistence, and occupancy. A species was considered missing if it was not collected in recent surveys (post 1990’s), declining if there was a decrease in occupancy between pre-1970 and 2000’s time periods, and common if it was detected at more than 50% of sites. Persistence was measured by the index of species turnover (modified from

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Diamond and May 1977) which is represented by the formula T = (C + E)/(S1 + S2) where C and E are the number of taxa that colonized or went extinct between the two time periods respectively, and S1 and S2 represent species richness during time intervals 1 and 2 (Meffe and Minckley, 1987; Meffe and Berra, 1988; Przybylski, 1994). This measures species turnover and ranges from 0 – 1. Persistence (P) was then calculated as 1-T, and ranges from zero (no persistence) to one (complete persistence) with values 0.7 indicating highly persistent assemblages between time periods and values in between considered moderately persistent (Meffe and Minckley, 1987; Patton et al., 1988; Meffe and Berra, 1988; Przybylski, 1994; Shearer and Berry, 2003). These values are also used for other similarity indices (e.g., Jaccard’s coefficient; Matthews et al., 1988; Phillips and Johnson, 2004; Kwak and Peterson, 2007). Occupancy was calculated as the number of sites a species was detected divided by the total number of sampling occasions. Patterns in species richness and effort were compared among HUs, basins, and time period using analysis of variance (ANOVA). In addition, discharge data was downloaded from seven USGS (2014) gaging stations along the James River and a mean annual discharge value was calculated. An ANOVA was used to characterize the hydrologic variability among decades. We assumed the James River hydrologic variability to be reflective of the Vermillion and Big Sioux rivers, as discharge data was sparse in the other two basins for the dataset time period. All analyses were conducted in SAS software (SAS Institute 2003) and an alpha level < 0.05 was considered significant.

Results

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There were dramatic increases in the amount of land (acres) converted to cropland beginning in the 1880’s which continued through 1910. Since 1910, between 70 % and 80 % of land has been devoted to cropland in all three river basins. In 2007, the Vermillion had had the most percentage of land devoted to crops (81%) followed by 73% of land in crops in the lower Big Sioux Basin. The upper Big Sioux Basin had the least amount of land devoted to cropland with 59%. In all basins, corn was the most frequent in 2007, except in the upper Big Sioux River which was largely in soy production. Percentages of cropland have undulated across the years, but there is a general decreasing trend across years beginning in 1910 (Figure B). Additionally, there was a transition from small grains to soy and corn production in most basins (Figure C). Discharge was significantly different (F4, 345 = 20.72, P < 0.0001) among decades with the 1990’s having the highest mean discharge (27.30 cubic meters/second), and the 1970’s having the lowest mean discharge (6.32 cubic meters/second; Figure D). Seventy-seven species were collected from the Big Sioux, James, and Vermillion rivers over the entirety of the dataset (Table 2). Thirteen species were unique to the Big Sioux River, four to the James River, and seven to the Vermillion River (Table 2). Species richness did not vary between basins (F2, 573 = .09, P = 0.91), but did vary between tributary and mainstem sites for the James River (F1 = 41.94, P < 0.0001), with mainstem sites having more species. Species richness differed among decades for the Big Sioux River, with the 1970’s and 1980’s being less speciose than the other three decades (F4, 277= 10.28, P < 0.0001), but did not differ for the James or Vermillion rivers among decades (Figure E; F4, 220 = 0.16, P < 0.18, and F3, 65 = 3.84, P < 0.01, respectively). Overall mean persistence was high in the James (P = 0.84) and Big Sioux

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(P = 0.75) rivers and 0.53 in the Vermillion river (Table 3). There was also a steady increase in the number of nonnative species to inhabit these ecosystems although the Big Sioux River has not followed this trend (Figure E). Fifty-three species were collected in the James River throughout the duration of this dataset. Six species (fathead minnow, Pimephales promelas, black bullhead, Ameiurus melas, sand shiner, Notropis stramineus, common carp, cyprinus carpio, orangespotted sunfish, Lepomis humilis, and red shiner, Cyprinella lutrensis), were detected at more than 50% of sites sampled (i.e., common) and exhibited 100% persistence across all five sampling periods. Seven species (silverband shiner, N. shumardi, plains minnow, Hybognathus placitus, pallid sturgeon, Scaphirhynchus albus, longnose dace, Rhinichthys cataractae, grass carp, Ctenopharyngodon idella, chinook salmon, Oncorhynchus tshawytscha, and bluntnose minnow, P. notatus) were collected only once (Table 2). Two species were not detected in recent surveys, one species was considered declining, and seven species, three of which are non-native, were first collected in recent surveys (Table 3). Species richness was variable across time intervals, increasing from 33 in the first time interval to 43 in the 2000’s. Upper watersheds were consistently less speciose and more persistent (P = 0.69), than the lower watersheds (P = 0.63; Figure F). Lower and middle watersheds were more consistent in occupancy across decades than the upper watersheds (Figure G). Additionally, there was an increasing trend in occupancy for common species at mainstem sites while occupancy of tributary sites has declined steadily since the 1980’s (Figure H). This is due in part to the decrease in red shiner occupancy. Despite an increase in the number of nonnative species over the course of this dataset (Figure E), the percentage of sites occupied by them remains low

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(e.g., < 0.50) exhibiting a decreasing pattern across decades (Figure I). Persistence of all species combined ranged from 0.79 and 0.86 across time periods (Table 3) with a mean value of 0.84 (SE = 0.01) indicating high persistence within this drainage. Forty-two species were collected in the Vermillion River throughout the duration of this dataset. Nine species (fathead minnow, sand shiner, creek chub, Semotilus atromaculatus, red shiner, common shiner, Luxilus cornutus, green sunfish, L. cyanellus, white sucker, Catostomus commersonii, Topeka shiner, N. topeka, and black bullhead) were collected in more than 50% of collections (N = 71) and exhibited high persistence (P = 0.89) from time period one to time period five. Occupancy of these species was variable across decades ranging from 0.14 before 1970 and increasing in the more recent decades (Figure G). Additionally, mainstem sites consistently had higher occupancy than tributary sites (Figure H). Fourteen species (yellow perch, Perca flavescens, white bass, Morone chrysops, smallmouth buffalo, Ictiobus bubalus, silver lamprey, Ichthyomyzon unicuspis, shovelnose sturgeon, sauger, Sander canadense, paddlefish, Iowa darter, Etheostoma exile, golden shiner, Notemigonus crysoleucas, finescale dace, Chrosomus neogaeus, flathead chub, Platygobio gracilis, blacknose dace, R. atratulus, bigmouth buffalo, I. cyprinellus, and American eel, Anguilla rostrata) were detected only once (Table 2). Ten species were not detected in recent surveys and 22 species, five of which are non-native, were first collected in recent surveys (post 1990; Table 4). Species richness was variable across time intervals, ranging from 28 species in early surveys (pre 1970’s), 42 species in the 1990’s, and 23 in the 2000’s (Figure E). Despite an increase in the number of nonnative species over the course of this dataset (Figure E), the percentage of sites occupied by these species remains consistently low (e.g., < 0.30; Figure I).

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Persistence of all species was also variable among time periods (Table 3) with a persistence value of 0.38, indicating low persistence within this drainage. Sixty-four species were collected in the Big Sioux River throughout the duration of this dataset. Six species (white sucker, fathead minnow, sand shiner, creek chub, common shiner, and johnny darter, E. nigrum) had 100% persistence across time periods and occupied at least 50% of sampling sites. Occupancy between middle and lower watersheds was fairly stable across decades with fish occupying fewer sites in the lower watershed (Figure G). Seven species (silver lamprey, plains minnow, mooneye, H. tergisus, golden shiner, flathead catfish, European rudd, Scardinius erythrophthalmus, southern redbelly dace, P. erythrogaster) were only detected on one sampling occasion. Five species were missing from recent collections, four species were considered declining, and 17 species (two of which are non-native), were first collected in more recent surveys (post - 1990; Table 4). Species richness was variable across time intervals ranging from 43 in the pre -1970’s collections to 54 species in the 1990’s and 37 species in the 2000’s. The upper Big Sioux watershed was consistently less speciose and more persistent between the first time period and the last time period (P = 0.87) than the lower portion of the watershed (P = 0.63; Figure E). Occupancy between the mainstem and tributary sites was similar and remained relatively consistent across decades (Figure H). Despite an increase in the number of nonnative species over the course of this dataset, the percentage of sites occupied by them remains low (e.g., < 0.50) with a decreasing trend across decades between the middle and lower watersheds (Figure I). Persistence of all species was also variable among time periods ranging from 0.71 to 0.78 (Table 2) with a

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mean persistence value of 0.75 (SE = 0.02), indicating relatively high persistence within this drainage.

Discussion The Big Sioux, James, and Vermillion rivers have diverse assemblages, each with fauna typical of prairie streams (e.g., plains topminnow, Fundulus sciadicus, Topeka shiner, and red shiner). The fish assemblage has remained persistent over the dataset timeframe, especially in the James River, which is consistent with studies over shorter time periods in the Big Sioux and James river basins (Braaten and Berry, 1997; Dieterman and Berry, 1999; Shearer and Berry, 2003). High persistence within HUs and basins indicates that localized colonizations and extinctions are rare (Meffe and Berra, 1988). Conversely, there was a 22% decrease in native fish richness and a 13% increase in non-native fish richness in the entire study area. This finding however, is less than what has been documented in other regions. For example, 67% of fish in the Illinois River have declined or been extirpated (Karr et al., 1985) and all native fishes in the Colorado River drainage are considered threatened, endangered or extinct (Frissell, 1993). Additionally, despite an increase in nonnative species richness, occupancy of these species was variable throughout the study area and across time periods, but generally remained low (i.e., < 50%), especially in the Big Sioux and Vermillion rivers. In fact, common carp occupancy decreased in the Big Sioux River by 30% in the lower watershed and 43% in the middle watershed. Similarly, McClelland et al. (2012) found a long-term decline in relative abundances of common carp and other nonnative species in the Illinois River, despite increases in species richness, and attributed this decline to

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rehabilitation efforts along the river. As a result of the Clean Water Act of 1972, water quality in the Big Sioux River has improved, particularly in the middle and upper reaches (Dieterman and Berry, 1998) which concurrently had the steepest declines in land devoted to cropland. The lower persistence of the fish composition in the Vermillion River may be the result of low sampling effort (N = 6) in the pre-1970’s time period, as persistence between 1990-1999 and 2000-2005 was moderately high (P = 0.68). In addition, persistence values for all rivers may be biased low from the numerous rarely captured species among basins (e.g., gizzard shad, Dorosoma cepedianum, freshwater drum, Aplodinotus grunniens; Meffe and Berra, 1988). Gizzard shad and freshwater drum rarely occurred, likely because they are considered large river species and are abundant in the Missouri River (Berry and Young, 2004). Lower persistence can also be attributed to species appearing and disappearing between time periods (Meffe and Berra, 1988). For example, smallmouth bass, Micropterus dolomieu, were collected every other sampling period in the James River Basin (P = 0.01). Common species (those detected in over 50 % of collections) were numerous across basins and were completely persistent across decades within the James and Big Sioux rivers and had high persistence in the Vermillion River (P = 0.89). The high occupancy of many common species is indicative of a highly persistent assemblage (Connel and Sousa, 1983). Some common species (e.g., Topeka shiner, black bullhead, orangespotted sunfish, and red shiner) are considered tolerant in prairie streams (Matthews, 1985; Wall and Berry, 2004) that are hydrologically extreme, turbid, and sometimes hypoxic (Matthews 1988). Prairie fish assemblages are typically comprised of

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species that are habitat, trophic, or reproductive generalists (Bramblett et al., 2005), as it would be evolutionarily disadvantageous to specialize in stochastic prairie streams (Braaten and Berry, 1997). As such, many of the common species within this dataset are considered generalists (e.g., green sunfish, creek chub; Shearer and Berry, 2002). The Topeka shiner, which is listed as federally endangered due to an 80% reduction in its distribution as a result of land use alterations (Wall and Berry, 2004), actually increased in the number of sites occupied across time intervals (see also Wall and Berry, 2006). The persistence of the Topeka shiner suggests that the habitats within this region have experienced less recent land-use alterations, and stream modifications than more southern regions (Shearer and Berry, 2003; Wall and Berry, 2006) or other prairie streams. For example, only 35% of wetlands have been drained in South Dakota (Wall and Berry, 2006) compared to Iowa or Kansas that have lost 89% and 48% of wetlands, respectively (Dahl, 1990). Wetlands provide important fish and wildlife habitats, maintain and recharge ground water supplies and water quality, provide erosion control, control and store floodwater, trap sediment, and act a nutrient sinks (Dahl, 1990; Murkin, 1998). The stochastic hydrology of prairie streams may reduce residency and delivery times of agricultural chemicals (Bramblett et al., 2005), thus reducing the impacts of agricultural inputs. Additionally, Iowa has converted 72% of land to crops in comparison to the 41% of South Dakota currently in crops (Heitke et al., 2006) although between 59% and 81% of land in eastern south Dakota is devoted to cropland. The high persistence of the Topeka shiner and other fishes in this dataset may be attributed to the relatively low impact of anthropogenic alterations compared to other Midwestern streams

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which have in turn seen more declines in fish species richness (Karr et al., 1985; Larimore and Bayley, 1996; Patton et al., 1998; Shearer and Berry, 2003; Johnson, 1995). Despite moderate to high persistence in assemblage composition over the study time period and increased sampling effort, numerous native species were not detected in recent surveys and are considered extirpated or “missing” from the region (e.g., northern hogsucker, Hypentelium nigricans and suckermouth minnow, Phenacobius mirabilis), consistent with previous findings of faunal change on a larger spatial and broader time scale (Hoagstrom et al., 2007b). Within eastern South Dakota disturbances such as long term land-use practices (e.g., row crop agriculture, livestock grazing, and wetland drainage) that have led to increased sedimentation and erosion, and low head dams that have led to fragmentation, may have influenced the extirpation or decline in occupancy of some intolerant fishes (Wood and Armitage, 1997; Henley et al., 2000; Hoagstrom et al., 2006; Slawski et al., 2008). These disturbances have led to decreased water quality (e.g., Big Sioux River) and increased erosion and flooding (e.g., James and Vermillion rivers) that resulted in the decline or extirpation of fishes in other regions (Matthews, 1988; Henley et al., 2000; Gido et al., 2006; Hoagstrom et al., 2007b). For example, the northern hogsucker, was last collected in the Big Sioux River in 1970. This fish exhibits seasonal movements between slower, deeper habitat with smaller substrates, to faster, shallow habitat with larger substrates (Matheney and Rabeni, 1995). Although water quality improved in the Big Sioux River since the early 1960’s (Dieterman and Berry, 1998), and dominant substrates within the basin are sand and silt with rare rocky outcrops, habitat diversity likely decreased as a result of increased erosion from land-use practices that have increased sedimentation (Dieterman and Berry,

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1998). Sedimentation increases instream turbidity, scouring, and abrasion, causes bottomup effects in food webs, covers gills and respiratory organs, and degrades fish habitat (Wood and Armitage, 1997; Henley et al., 2000; Allan, 2004). Additionally, the golden shiner was moderately common in historical collections (detected at 22% of sites in pre1970) but has not been captured in the James River where it is native since these early surveys. The typical habitats of the golden shiner include low gradient streams that are characteristic of the James River basin with clear water and abundant vegetation (Scott and Crossman, 1973) that is now rare as a result of turbidity and siltation from land-use practices (Owen et al., 1981; Trautman, 1981; Hoagstrom et al., 2006; Wall et al., 2006). Of the 20 species listed by SDGFP (SDGFP, 2012) as species of greatest conservation need, eight were present within eastern South Dakota Rivers, and of these, three (i.e., Carmine shiner, N. percobromus, finescale dace, paddlefish) were not collected in recent surveys (1990’s and early 2000’s; also see Hoagstrom et al., 2007b). The finescale dace, a state-threatened fish, is at the edge of its range in South Dakota and is believed to be impacted by habitat loss brought about by a decrease in beaver populations (Isaak et al., 2002) and habitat degradation, as they prefer cool spring fed streams (SDGFP, 2006) that are rare in eastern South Dakota. The Carmine shiner, formerly called the rosyface shiner, N. rubellus, is also at the edge of its range in South Dakota (Wood et al., 2002; Hayer et al., 2006). They are intolerant of turbidity and flow fluctuations, and occupy a narrow habitat niche of shallow riffles with clear water (COSEWIC, 2006). The disappearance of this species is probably the result of water pollution and increased sedimentation within the Big Sioux River basin.

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There were numerous species that were first detected in recent surveys (post1990) that could be the result of high discharge, range expansion, increased sampling effort and efficiency, stocking or invasive species. Some species (e.g., blue sucker and western silvery minnow, H. argyritis) may have been incidentally collected as a result of high discharge in the 1990’s, allowing them to migrate freely into the tributaries (i.e., Big Sioux James, and Vermillion rivers) from the Missouri River (Bailey and Allum, 1962; Berry et al., 1993). For example, the blue sucker, western silvery minnow and typical large river fishes (e.g., stonecat, Noturus flavus, quillback, Carpiodes cyprinus, smallmouth buffalo; Shearer and Berry, 2003), were first detected in the 1990’s, during high discharge events (Shearer and Berry, 2003) that allowed them to colonize the tributaries (Shearer and Berry, 2003). The James River may provide natural free-flowing conditions characteristic of the Missouri River before damming (Berry and Galat, 1993; Berry et al., 1993). Large river fishes (e.g., gar, Lepisosteus sp, emerald shiner, N. atherinoides, smallmouth buffalo, white bass, blue sucker, and paddlefish) are often only found in the lower portion of the James River (Berry et al., 1993). Hydrology strongly influences stream fish assemblages (Nui et al., 2012) in that there are typically more species present during high discharge periods (Angermeier and Schlosser, 1989; Oberdorf et al., 1997; Xenopoulos and Lodge, 2006; Nui et al., 2012), which supports our findings of higher species richness in the 1990’s. Range expansion potentially played a role in the collection of southern redbelly dace as it was not previously reported in South Dakota (Hayer et al., 2006), although they occur in the Big Sioux River basin in Minnesota (Bailey and Allum, 1962; Lee et al., 1980). Another possible reason for increased richness in recent collections is increased

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sampling effort and efficiency (Shearer and Berry, 2003; Hayer et al., 2006). Sampling effort was highest in the 1990’s, coinciding with high discharge periods and higher species richness. More effort in this time period may have been the result of increased conservation efforts for native species, and in particular the federally endangered Topeka shiner (Wall and Berry, 2004). Additionally, there may have been more habitats to sample as a result of high discharge that likely facilitated colonization by non-natives. Many non-native species (e.g., yellow bullhead, A. natalis, northern pike, Esox lucius, largemouth bass, M. salmoides and smallmouth bass, white bass; Schmulbach and Braaten, 1993) were present in recent surveys, some as a result of stocking in the numerous small impoundments within each basin (Chipps et al., 2007). For example, Chinook salmon, and northern pike were stocked in Lake Oahe on the Missouri River in the 1980’s (Barnes, 2007), and smallmouth bass were stocked in numerous northeastern South Dakota lakes in the 1980’s that now support self-sustaining populations (Barnes, 2007). Additionally, golden shiners were stocked across the state beginning in the late 1970’s (Barnes, 2007) and white bass were introduced and have become established in the Missouri River and dispersed into the Vermillion River (Schmulbach and Braaten, 1993). The James River has been stocked with smallmouth bass, northern pike, walleye, crappie (Pomoxis spp.), largemouth bass, yellow perch, and black bullhead (Berry et al., 1993). Two non-native species that were detected in eastern South Dakota rivers, common carp, and grass carp, are designated as aquatic nuisance species (Burgess and Bertrand, 2008; Gozlan, 2010). Common carp are considered to be tolerant and widespread within the state and compete with native fishes, especially benthic fishes

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(Burgess and Bertrand, 2008), although to our knowledge, there have been no studies in South Dakota that have documented negative impacts in lotic systems. However, they have caused increases in turbidity (Fletcher et al., 1985; King et al., 1997), destruction of aquatic vegetation (Fletcher et al., 1985; Roberts et al., 1995), modification of aquatic invertebrate assemblages (Robertson et al., 1997) and changes in abundance of native fish species (Gehrke and Harris, 2001) in other regions. Common carp were persistent across all decades in all three basins, though there was a decreasing trend in occupancy in the Big Sioux River, and relative abundance remains low in eastern South Dakota Rivers (Braaten and Berry, 1997; Dieterman and Berry, 1998; Shearer and Berry, 2003; Wall and Berry, 2006). Common carp require vegetation for spawning in lotic systems (Balon, 1975) which, given the increased turbidity and unstable sediments from flooding in this region, may have decreased reproductive success by silting in spawning habitat (Resseguie and Kelsh, 2008). Additionally, grass carp were first detected in 2000 (Shearer and Berry, 2003), and their numbers have remained low within the Big Sioux, James, and Vermillion rivers (Hayer, unpublished data) as well as in the Missouri River (Wanner and Klumb, 2007). Grass carp can reduce and alter aquatic macrophytes, thus reducing habitat heterogeneity, degrade water quality, negatively affect aquatic fauna, decrease refugia for aquatic fauna, increase nutrient enrichment, disrupt food webs and trophic structure, and spread parasites and diseases (Bain, 1993; Cudmore and Mandrak, 2004; Conover et al., 2007; Dibble and Kovalenko, 2009). Despite documented negative impacts in other regions, grass carp do not seem to be having an impact on eastern South Dakota basins, likely due to their low abundances.

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Recent invaders, bighead carp, Hypophthalmichthys nobilis, and silver carp, H. molitrix, (Asian carps), are also considered nuisance species (Burgess and Bertrand, 2008) and have the potential to dramatically alter aquatic ecosystems and food webs (Kolar et al., 2007). These two species were first detected at the confluence of the James River and the Missouri River in 2007 (Klumb, 2007; Kolar et al, 2007), after this database was compiled. In other invaded regions, they have reduced water quality, altered the plankton community (Kolar et al., 2007), and reduced condition of native fish (Irons et al., 2007). In the La Grange reach of the Illinois River, silver carp were first detected in 1998 and have exhibited exponential population growth since 2000 (Irons et al., 2011). These new invaders will test the resistance of these prairie ecosystems to yet another disturbance. Continued monitoring of eastern South Dakota streams is necessary to document potential effects and invasion of Asian carps in South Dakota. The fish assemblages within this study appear to be relatively persistent across a broad timescale, despite harsh climatic and hydrologic regimes, watershed modifications from land use practices, and instream alterations such as low head dams or channelization. In addition, these river basins have exhibited characteristics of the biotic resistance hypothesis which states that native assemblages persist despite introduction of invasive species (e.g., common carp and grass carp; Moyle et al., 1982; Moyle and Vondracek, 1985; Baltz and Moyle, 1993) and continued landuse for crops. The stochastic nature of prairie streams in South Dakota may inhibit or reduce the establishment of some colonizing species (Gido et al., 2004). However, the potential negative impacts of a major disturbance such as an invasive species are difficult to predict (Ricciardi and Rasmussen, 1998; Ehrenfeld, 2010; Simberloff, 2011) as non-

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natives can accelerate fish assemblage changes by having assemblages with persistent natives and established non-natives (e.g., streams in eastern South Dakota) or having species poor assemblages dominated by non-natives (Gido and Brown, 1999). These data can provide a baseline to demonstrate future ecological impacts that invasive species and other anthropogenic alterations may have on the unique and persistent fauna of these prairie rivers. As land-use alterations continue to increase on a global and local scale, and introduced and invasive species continue to increase dramatically in novel systems, it is important to establish historical patterns in baseline data in order to document potential current and future impacts of these disturbances.

Acknowledgements.-- We thank the South Dakota Gap Analysis Project for providing the fish database used for these analyses and specifically the South Dakota Cooperative Fish and Wildlife Research Unit that is jointly sponsored by South Dakota Department of Game Fish and Parks, Wildlife Management Institute, U.S. Fish and Wildlife Service, South Dakota State University, and U.S. Geological Survey. We would like to recognize J. VanDeHey for earlier review of the manuscript.

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Figure 2.A. 8-digit hydrologic unit (HUs) in the James, Vermillion, and Big Sioux rivers in eastern South Dakota.

63

Figure 2.B. Percent of land as cropland (includes corn, soy, wheat, hay and small grains) in eastern South Dakota river basins from 1900 to 2007. Hydrologic units include Upper Big Sioux, middle Big Sioux, lower Big Sioux, upper James, middle James, lower James, and Vermillion.

64

Figure 2.C. Percentage of small grains, hay, wheat, soy, and corn devoted to landuse between 1950 and 2007.

65

Time Period

Figure 2.D. Mean James River Basin discharge (m3/s) and one standard error by time period. The solid line represents the total mean discharge for the basin and the dotted lines represent 95% confidence intervals. Means with the same letter are not significantly different (P>0.05).

66

Time Period

Figure 2.E. Nonnative species richness across time periods in the James, Vermillion, and Big Sioux rivers in eastern South Dakota.

67

1.0

James River

0.8

0.6

0.4

0.2

0.0 1.0

Big Sioux River

Occupancy

0.8

Upper Middle Lower

0.6

0.4

0.2

0.0 1.0

Vermillion River

0.8

0.6

0.4

0.2

0.0

e Pr

19

70 19

9 -1 70

79 19

9 -1 80

89 19

9 -1 90

99 20

0 -2 00

05

Figure 2.F. Common species occupancy (species that occupied > 50% of sites sampled across the entirety of the dataset) between tributary and mainstem sites.

68

1.0

James River

0.8

0.6

0.4

0.2

0.0 1.0

Big Sioux River

Occupancy

0.8

0.6

Mainstem Tributary

0.4

0.2

0.0 1.0

Vermillion River

0.8

0.6

0.4

0.2

5

20

00

-2

00

9 99

9 90

-1

98 -1 19

19

80

-1 70 19

Pr

e

19

97

70

9

0.0

Figure 2.G. Common species occupancy (species that occupied > 50% of sites sampled across the entirety of the dataset) among hydrologic units (upper, middle, lower) for the James, Big Sioux and Vermillion rivers in eastern South Dakota.

69

1.0

James River

0.8

0.6

0.4

0.2

0.0 1.0

Big Sioux River

Occupancy

0.8

0.6

Upper Middle Lower

0.4

0.2

0.0 1.0

Vermillion River

0.8

0.6

0.4

0.2

5 -2

00

9 00 20

19

90

-1

99

9 98 -1 80

-1 19

70 19

Pr e

19

97

9

70

0.0

Figure 2.H. Nonnative species occupancy among hydrologic units (upper, middle, lower) for the James, Big Sioux and Vermillion rivers in eastern South Dakota.

70

Table 2.1. Number of sampling occasions per Hydrologic Unit and location within the basin for eastern South Dakota.

Basin Big Sioux Big Sioux Big Sioux James James James Vermillion

HU Location Lower Middle Upper Lower Middle Upper Total

Pre1970 18 21 3 9 7 2 6 66

19701979 6 2 2 15 14 22 1 62

19801989 0 0 3 9 36 6 0 54

19901999 103 89 7 29 43 1 45 317

20002009 23 23 2 27 11 3 19 108

Total 150 135 17 89 111 34 71 607

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Table 2.2. Persistence values and total mean value across decades for the Big Sioux, James and Vermillion River basins between decades. Values in parenthesis are one

Vermillion

James

Big Sioux

standard error. Decade 1970’s 1980’s 1990’s 2000-2005 Mean Pre1970’s 0.77 0.78 1970’s 1980’s 1990’s 0.71 0.75 (0.02) Pre-1970 0.86 0.79 0.79 0.82 1970’s 0.89 0.87 0.87 1980’s 0.84 0.83 1990’s 0.86 0.84 (.01) Pre1970’s 0.55 0.68 1970’s 1980’s 1990’s 0.38 0.53 (0.08)

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Table 2.3. Fish species collected, number of sites collected at (number of sites) and proportion of sites collected (occupancy; O) for the Big Sioux, James, and Vermillion rivers since 1934. Big Sioux Common Name

Scientific Name

American eel

Anguilla rostrata

Bigmouth buffalo

Ictiobus cyprinellus

Black buffalo

Ictiobus niger

Blackside darter

Number of sites

O

James Number of sites

Vermillion O

Number of sites

O

1

0.01

68

0.29

1

0.01

5

0.02

Percina maculata

17

0.06

Bigmouth buffalo

Ictiobus cyprinellus

17

Bigmouth shiner

Notropis dorsalis

138

0.46

35

0.15

26

0.37

Black bullhead

Ameiurus melas

140

0.46

158

0.68

36

0.51

Black crappie

Pomoxis nigromaculatus

12

0.04

67

0.29

6

0.08

Blacknose dace

Rhinichthys atratulus

51

0.17

1

0.01

Bluegill

Lepomis macrochirus

4

0.01

14

0.06

10

0.14

Bluntnose minnow

Pimephales notatus

85

0.28

1

0.00

Brassy minnow

Hybognathus hankinsoni

56

0.19

47

0.20

35

0.49

Brook stickleback

Culaea inconstans

57

0.19

11

0.05

2

0.03

Blue sucker

Cycleptus elongatus

8

0.03

2

0.01

4

0.06

Channel catfish

Ictalurus punctatus

65

0.22

67

0.29

16

0.23

Central mudminnow

Umbra limi

3

0.01

0.06

73

Chinook salmon

Oncorhynchus tshawytscha

1

0.00

Common shiner

Luxilus cornutus

167

0.55

38

0.16

43

0.61

Common carp

Cyprinus carpio

105

0.35

156

0.67

30

0.42

Creek chub

Semotilus atromaculatus

179

0.59

68

0.29

48

0.68

Emerald shiner

Notropis atherinoides

46

0.15

58

0.25

5

0.07

European Rudd

Scardinius erythrophthalmus

1

0.00

Flathead catfish

Pylodictis olivaris

1

0.00

10

0.04

3

0.04

Fathead minnow

Pimephales promelas

210

0.70

159

0.68

55

0.77

Finescale dace

Chrosomus neogaeus

1

0.01

Flathead chub

Platygobio gracilis

1

0.01

Freshwater drum

Aplodinotus grunniens

Goldeye

Hiodon alosoides

Golden shiner

Notemigonus crysoleucas

Grass carp

Ctenopharyngodon idella

Green sunfish

Lepomis cyanellus

Gizzard shad

Dorosoma cepedianum

Iowa darter

Etheostoma exile

Johnny darter

Etheostoma nigrum

Largemouth bass

Micropterus salmoides

6

0.02

64

0.27

5

0.07

12

0.04

43

0.18

3

0.04

1

0.00

3

0.01

1

0.01

1

0.00

79

0.26

92

0.39

41

0.58

6

0.02

67

0.29

5

0.07

51

0.17

12

0.05

1

0.01

162

0.54

56

0.24

31

0.44

3

0.01

16

0.07

13

0.18

74

Longnose dace

Rhinichthys cataractae

1

0.00

Longnose gar

Lepisosteus osseus

3

0.01

4

0.06

Logperch

Percina caprodes

2

0.01

Mooneye

Hiodon tergisus

1

0.00

Northern hog sucker

Hypentelium nigricans

2

0.01

Northern Pike

Esox lucius

99

0.33

69

0.29

7

0.10

Northern redbelly dace

Phoxinus eos

3

0.01

Orangespotted sunfish

Lepomis humilis

79

0.26

138

0.59

45

0.63

Paddlefish

Polyodon spathula

1

0.01

Pallid sturgeon

Scaphirhynchus albus

Plains minnow

Hybognathus placitus

Plains topminnow

Fundulus sciadicus

Quillback

Carpiodes cyprinus

17

Red shiner

Cyprinella lutrensis

River carpsucker

Carpiodes carpio

River shiner

Notropis blennius

Carmine shiner

Notropis percobromus

3

0.01

Smallmouth buffalo

Ictiobus bubalus

2

0.01

Sauger

Stizostedion canadense

2

0.01

1

0.00

1

0.00

7

0.03

2

0.03

0.06

4

0.02

2

0.03

139

0.46

135

0.58

47

0.66

26

0.09

62

0.26

17

0.24

4

0.06

1

0.01

1

0.01

1

0.00

15

0.06

75

Sand shiner

Notropis stramineus

Spotfin shiner

Cyprinella spiloptera

Shortnose gar

Lepisosteus platostomus

Shorthead redhorse

Moxostoma macrolepidotum

Shovelnose sturgeon

Scaphirhynchus platorynchus

Silver lamprey

Ichthyomyzon unicuspis

Silverband shiner

Notropis shumardi

Smallmouth bass

Micropterus dolomieu

Stonecat

Noturus flavus

Silver chub

Macrhybopsis storeriana

Central stoneroller

Campostoma anomalum

Spottail shiner

197

0.65

157

0.67

52

0.73

3

0.04

7

0.02

45

0.19

5

0.07

74

0.25

54

0.23

14

0.20

1

0.01

1

0.01

1

0.00 1

0.00

6

0.02

4

0.02

48

0.16

3

0.01

9

0.13

5

0.02

103

0.34

30

0.13

17

0.24

Notropis hudsonius

4

0.01

4

0.02

Suckermouth minnow

Phenacobius mirabilis

8

0.03

Tadpole madtom

Noturus gyrinus

61

0.20

30

0.13

7

0.10

Topeka shiner

Notropis topeka

47

0.16

37

0.16

36

0.51

Trout-perch

Percopsis omiscomaycus

13

0.04

Walleye

Stizostedion vitreum

50

0.17

59

0.25

2

0.03

White bass

Morone chrysops

19

0.06

4

0.02

1

0.01

White crappie

Pomoxis annularis

9

0.03

64

0.27

8

0.11

76

White sucker

Catostomus commersonii

213

0.71

80

0.34

Western silvery minnow

Hybognathus argyritis

2

0.01

2

0.01

Yellow bullhead

Ictalurus natalis

10

0.03

23

0.10

Yellow perch

Perca flavescens

49

0.16

20

0.09

36

0.51

1

0.01

77

Table 2.4. Status of missing species (not detected recently), declining species (by percent decrease in sites detected), and species additions to a system (first detected in recent collections). An * indicates a non-native species to that basin and ** indicates the species was only collected from one sampling occasion. Missing

Big Sioux River

Species

Declining Last detected

Species

Additions

% decrease in occupancy

Species

Year first detected

Black buffalo

1971

Bigmouth buffalo

0.09

Blue sucker

1994 **

Carmine shiner

1956

Emerald shiner

0.24

Bluegill

1994

Mooneye

1970

Quillback

0.13

European rudd *

1997 **

Northern hogsucker

1970

White crappie

0.10

Flathead catfish

1994

Suckermouth minnow

1971

Golden shiner

2003

Largemouth bass

1993

Logperch

1996 **

Plains minnow

2003

River carpsucker

1994

Sauger

1994

James

78

Golden shiner

1952

Spottail shiner

1979

Iowa darter

0.36

Shortnose gar

1994 **

Silver chub

1994 **

Silver lamprey

1997

Smallmouth bass *

1993

Southern Redbelly dace

2003

Western silvery minnow

1997

Yellow bullhead

1994

Chinook salmon *

2002 **

Grass carp *

2000 **

Longnose dace *

1999 **

Longnose gar

1990

Silverband shiner

1997 **

Stonecat

1999

Western silvery minnow

1997

79

Missing

Species

Vermillion

American eel

Declining Last detected

Species

Additions

% decrease in occupancy

Species

Year first detected

pre 1959

Bigmouth shiner

1991

Bigmouth buffalo

1934

Black crappie *

1991

Brook stickleback

1979

Bluegill

1991

Finescale dace

pre 1959

Brassy minnow

1991

Flathead chub

pre 1959

Central stoneroller

1991

Paddlefish

pre 1959

Channel catfish

1991

1965

Common shiner

1991

Plains topminnow Smallmouth buffalo

pre 1959

Golden shiner

1992 **

Shovelnose sturgeon

pre 1959

Iowa darter

1998 **

Silver lamprey

pre 1959

Largemouth bass *

1991

Northern Pike *

1992

Quillback

1999

Sauger

1991

Shorthead redhorse

1991

Spotfin shiner *

1991

80

Stonecat

1991

Tadpole madtom

1991

Walleye

1991

White bass

1991 **

White crappie *

1991

White sucker

1991

Yellow perch

1999 **

81

CHAPTER THREE: ADULT, JUVENILE, AND YOUNG-OF-YEAR BIGHEAD, HYPOPHTHALYMICHTHYS NOBILIS, AND SILVER CARP (H. MOLITRIX) RANGE EXPANSION ON THE NORTHWESTERN FRONT OF THE INVASION IN NORTH AMERICA

Introduction Invasive species are becoming a worldwide epidemic and are triggering changes in the structure and function of invaded ecosystems (Ricciardi et al. 2000). One pair of global invaders, bighead, Hypophthalmichthys nobilis, and silver carp, H. molitrix, (collectively referred to as Asian carps) have been introduced intentionally and unintentionally throughout the world, mostly for aquaculture purposes (Kolar et al. 2007) as they are the most important aquaculture species in Asia and east-central Europe (Lieberman 1996; Penman et al. 2005). Bighead carp have invaded 74 countries and are reproducing in 19 and the silver carp has invaded 88 countries and are reproducing in 23 (Kolar et al. 2007). Both species of Asian carps are currently reproducing in the United States (Papoulias et al. 2006; DeGrandchamp et al. 2007; Lohmeyer and Garvey 2009; Deters et al. 2013). Asian carps were originally introduced into the southern United States in aquaculture ponds in the early 1970’s where both species subsequently escaped and began their migration up the Mississippi River and into associated tributaries (e.g., Missouri River, Illinois River, Ohio River; Kolar et al. 2007). The northern front of the Asian carps invasion is in the Mississippi River basin where they threaten to invade the Great Lakes through the Chicago Shipping Canal (e.g., Lake Michigan) and the Illinois River, a tributary to the Mississippi River (Kocovsky et al. 2012). The eastern front of the

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invasion includes the Ohio River where they threaten to enter the Great Lakes (e.g., Lake Erie) through the Wabash River, a tributary to the Ohio River, and the Maumee River, the second largest tributary to Lake Erie (Kocovsky et al. 2012). On the northwestern front, Asian carp threaten to invade three tributaries (e.g., James, Vermillion, Big Sioux rivers) to the Missouri River below Gavins Point Dam, that serve as important fish habitat for many threatened and endangered fishes (Berry et al. 2007). These tributaries already face disturbance from current and present landuse practices and natural physiochemical and hydrologic fluctuations characteristic of prairie streams (Matthews 1988; Poff and Ward 1989; Hayer et al. 2014a). Prior to this study, distribution of Asian carps in three Great Plains tributaries in South Dakota was unknown. However; one bighead carp was caught by an angler in the middle James River near Mitchell, South Dakota in 2008 and one silver carp was caught by an angler in the Big Sioux River near Canton, SD in 2004 (Kolar et al. 2005). The goal of this paper is to document the northwestern invasion front of adult, juvenile, and young-of-year Asian carps in three prairie tributaries and other South Dakota waters (e.g., lakes) by providing the northern most latitude detection locations.

Methods Three prairie tributaries in South Dakota converge with an unchannelized section (Galat et al. 2005a) of the Missouri River just downstream of Gavins Point Dam: the James, Vermillion and Big Sioux rivers. These warmwater rivers drain the Central Lowlands physiographic province in South Dakota (Galat et al. 2005b) and are characterized by low gradient streams of glacial origin (Hoagstrom et al. 2007). The

83

James River (watershed area = 57,000 km2) extends 760 Rkm from southeastern North Dakota through eastern South Dakota to its confluence with the Missouri River (Berry et al. 1993; Figure 1). The Vermillion River, the smallest basin (watershed area = 5,800 km2) extends 243 Rkm from the confluence of West and East Fork Vermillion rivers to its confluence with the Missouri River (Schmulbach and Braaten 1993; Figure 1). The Big Sioux River (watershed area = 23,325 km2) extends 470 Rkm from the Prairie Coteau of northeastern South Dakota to its confluence with the Missouri River at the South Dakota-Nebraska-Iowa border (Figure 1). Standardized boat electrofishing occurred between 2009 and 2012 at five sites on the James River, two sites on the Vermillion River, and three sites on the Big Sioux River (Figure 1), all within South Dakota. Sampling consisted of three 10 minute electrofishing runs which generally covered three river kilometers. Sampling occurred once during each of three seasons: spring (May – June), summer (July – August), and fall (September – October). Additional non standardized boat electrofishing occurred at various sites on the James River in North and South Dakota. Adult carps were considered to be greater than 600 mm TL (age 3+), juveniles were between 300 and 600 mm TL (ages 1 and 2), and age-0 were less than 300 mm TL. Life stage ages were verified by analyzing otoliths (Hayer et al. 2014b).

Results Silver carp have invaded as far as they can naturally as three significant barriers, a 110 foot high dam on the James River in North Dakota, a 41 foot high dam on the Vermillion River in South Dakota, and a series of natural falls on the Big Sioux River in

84

South Dakota are the most upstream locations within these basins (Figure 1) where bighead and silver carps have been collected (Figures 2 - 4).

Adults One adult silver carp was collected as far upstream as Milltown on the James River (TL = 752, 4958 g) on August 2, 2012 (Figure 2) and have not been collected above the confluence of the Vermillion River with the Missouri River where nine silver carp were last collected on August 16, 2012 (TL = 681 – 784 mm; 3203 – 7285 g; Figure 2). Silver carp have also only been collected in the Big Sioux River at the confluence with the Missouri River and were last collected on August 17, 2012 (N = 5; TL = 634 – 783 mm; 3260 – 6123 g; Figure 2). One adult bighead carp was collected near Mitchell (August, 31 2010) on the James River (TL = 1001, 9072 g) and no adults have been collected in the Vermillion or the Big Sioux Rivers (Figure 2).

Juveniles Twelve juvenile silver carp were collected from Shue Creek, a tributary to the James River north of Huron on September 27, 2011 (TL = 409 – 507; 768 – 1474 g; Figure 3). One juvenile silver carp (450 mm, 1088 g) was subsequently collected in North Dakota on the James River in the Jamestown Reservoir tailrace (Figures 3, 4) on October 12, 2011. This is the first record of silver carp in North Dakota. Fifteen silver carp (TL = 311 – 369 mm, 299 – 656 g) were collected from the Vermillion River below East Lake Vermillion dam on August 11, 2011 and nineteen were collected in the Big

85

Sioux River at the confluence with the Missouri River (TL = 384 – 474 mm; 510 – 1191; Figure 3) on August 17, 2012. Six juvenile bighead carp (415 – 509 mm TL, 822 – 1389 g) were first collected from Firesteel Creek, a tributary to the James River near Mitchell on September 29, 2011 and were last collected in the Big Sioux River (N = 5; TL = 349 - 470 mm; 288 – 1389 g) on October 6, 2011 at the confluence with the Missouri River (Figure 3). Four bighead carp (309 – 372 mm TL, 299 – 565 g) were first collected from the Vermillion River below the East Vermillion Lake Dam on August 11, 2011 (Figure 3).

Age-0 Two age-0 silver carp were collected on Sept 29, 2011 in Shue Creek at the confluence with the James River (TL = 170, 174 mm; 44, 51 g; Figures 5,6). One youngof-year bighead carp was collected on October 20, 2010 in the Big Sioux River below the falls (TL = 298 mm, 301 g; Figure 6). Reproduction has been reported in the lower Missouri River (Shrank et al. 2001; Klumb 2007; Stukel et al. 2007), but as of the writing of this chapter has not been reported in the Missouri River in South Dakota or Nebraska.

Lakes Two bighead carp and one silver carp (TL approximately between 400 and 460 mm) were collected by a commercial fisherman with a seine pole on November 2, 2012 from Lake Byron which is the first confirmed lentic record in South Dakota (South Dakota Game Fish and Parks, personal communication).

86

Discussion The James River has over 230 low head dams which may impede movement during normal or low water years (Berry et al. 1993; Shearer and Berry 2003); however, record discharge and flooding in all basins in late 2010 and early 2011 (United States Geological Survey 2012) may have allowed for, and facilitated their unimpeded movement and colonization upstream and into normally unconnected lakes. Additionally, these records not only represent a range expansion, but they also represent invasion into new habitats (e.g., smaller watersheds, limited backwaters and floodplain lakes) which are atypical for established populations elsewhere (e.g., Illinois River, Ohio River, middle Mississippi River; Tucker et al. 1996; Kolar et al. 2007; DeGrandchamp et al. 2008). It is difficult to determine where these Asian carp populations are in the invasion process (e.g., dispersal, colonization, establishment, self-sustaining) as we have not confirmed reproduction in these basins; however, the presence of young-of-year silver carp in the middle to upper James River suggests reproduction may be occurring within the James River. As a direct result of the rapid expansion and increasing abundance of Asian carps in South Dakota, the South Dakota Game, Fish and Parks Department issued an emergency regulation in 2012 that closed these tributaries to commercial or recreational harvest of all bait fish in order to prevent further spread of Asian carps. South Dakota Missouri River tributaries support valuable fisheries and provide habitat for several threatened and endangered species (Berry et al. 2007). Continued monitoring and research on this newly invading population of Asian carps will provide invaluable insight into complex invasive species, assist with understanding Asian

87

carp population dynamics during an invasion, and expose the negative impacts Asian carp may be having on prairie stream ecosystems.

Acknowledgements We would like to thank the South Dakota Game Fish and Parks for their continued support of this project and specifically M. Smith and G. VanEeckhout for continued updates on Asian carp within the states. Additionally we would like to thank J Howell for assistance with field work.

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References Berry, CR Jr., Duffy WG, Walsh R, Kubeny S, Schumacker D, Van Eeckout G (1993) In: Hesse LW, Stalnaker CB, Benson NG, Zuboy JR (eds), The James River of the Dakotas. Restoration Planning for the Rivers of the Mississippi River Ecosystem, Biological Report 19, National Biological Survey, Washington, D.C. pp70-86 Berry CR., Higgins KF, Willis DW, Chipps SR (eds) (2007) History of fisheries and fishing in South Dakota. South Dakota Game, Fish and Parks, Pierre, South Dakota DeGrandchamp KL, Garvey JE, Csoboth LA (2007) Linking adult reproduction and larval density of invasive carp in a large river. Transactions of the American Fisheries Society 136:1327-1334. doi:10.1577/T06-233.1 DeGrandchamp, KL, Garvey JE, Colombo RE (2008) Movement and habitat selection by invasive Asian carps in a large river. Transactions of the American Fisheries Society 137:45-56. Galat, DL, Berry CR Jr., Peters EJ, White RG (2005a). Missouri River. In: Benke AC, Cushing CE (eds). Rivers of North America. Elsevier, Oxford, pp 427-480. Galat, DL, Berry CR Jr., Gardner WM, Hendrickson JC, Mestl GE, Power GJ, Stone C, and Winston MR (2005b). In: Rinne JN, Hughes RM, and Calamusso R, (eds), Spatiotemporal patterns and changes in Missouri River fishes. Historical changes in fish assemblages of large American Rivers. American Fisheries Society Symposium.

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Hayer C-A, Howell J, Graeb BDS, Bertrand KN (2014a) An historical watershed level perspective on the persistence and occupancy of fishes in prairie streams. American Midland Naturalist Hayer C-A, Bertrand KN, Graeb BDS (2014b) Population dynamics of bighead and silver carp on the northwestern front of their North American invasion. Aquatic Invasions, accepted. Hoagstrom CW, Wall SS, Kral JG, Blackwell BG, Berry CR (2007) Zoogeographic patterns and faunal change of South Dakota fishes. Western North American Naturalist 67:161-184. doi:10.3398/1527-0904(2007)67 Klumb, R. A. 2007. Shallow water fish communities in the Missouri River downstream of Fort Randall and Gavins Point dams in 2003 and 2004 with emphasis on Asian carps, Pierre South Dakota. Kocovsky PM, Chapman DC, McKenna JE (2012) Thermal and hydrologic suitability of Lake Erie and its major tributaries for spawning of Asian carps. Journal of Great Lakes Research 38:159-166. doi 10.1016/j.jglr.2011.11.015 Kolar, CS, Chapman D, Courtenay W, Housel C, Williams J, and Jennings D (2007) Bigheaded carps: a biological synopsis and environmental risk assessment, volume Special publication 33. American Fisheries Society, Bethesda. Kolar, CS, Chapman D, Courtenay W, Housel C, Williams J, and Jennings D (2005) Asian carps of the genus Hypophthalymichthys (Pisces, Cyprinidae) – a biological synopsis and environmental risk assessment. Report to the U.S. Fish and Wildlife Service. U.S. Geological Survey, LaCrosse, Wisconsin. 184 p

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Lieberman DM (1996) Use of Silver carp (Hypophthalmichthys molitrix) and Bighead carp (Aristichthys nobilis) for algae control in a small pond: changes in water quality. Journal of Freshwater Ecology 11:391-397. doi:10.1080/02705060.1996.9664466 Lohmeyer, AM, Garvey JE (2009) Placing the North American invasion of Asian carp in a spatially explicit context. Biological Invasions 11:905-916. doi:10.1007/s10530008-9303-5 Matthews WJ (1988) North American prairie streams as systems for ecological study. Journal of the North American Benthological Society 7:387-409. Penman, DJ, Gupta MV, Dey MM (eds.) (2005) Carp genetic resources for aquaculture in Asian. WorldFish Center Technical Report 65, 152 p Pflieger WL (1997) The fishes of Missouri. Second Edition. Missouri Department of Conservation, Jefferson City, Missouri. Papoulias, DM, Chapman D, Tillitt DE (2006) Reproductive condition and occurrence of intersex in bighead carp and silver carp in the Missouri River. Hydrobiologia 571:355-360. doi:10.1007/s10750-006-0260-7 Poff, NL, Ward JV (1989) Implications of streamflow variability and predictability for lotic community structure: a regional analysis of streamflow patterns. Canadian Journal of Fisheries and Aquatic Sciences 46:1805-1818. doi:10.1139/f89-228 Ricciardi A, Steiner WWM, Mack RN, Simberloff D (2000) Toward a global information system for invasive species. BioScience 50:239-244.

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Schrank, SJ, Braaten PJ, Guy CS (2001) Spatiotemporal variation in density of larval bighead carp in the lower Missouri River. Transactions of the American Fisheries Society 130:809-814. Shearer, JS, Berry, CR Jr. (2003) Fish community persistence in eastern North and South Dakota rivers. Great Plains Research 13:139-159. Stukel, S, Kral J, LaBay S (2007) 2006 Annual Report Pallid Sturgeon Population Assessment and Associated Fish Community Monitoring for the Missouri River: Segment 7. South Dakota Game Fish and Parks, Yankton, South Dakota. Prepared for the U.S. Army Corps of Engineers, Missouri River Recovery Program. March 2007. Tucker JK, Cronin FA, Hrabik RA, Peterson MD, Herzog DP. (1996) The bighead carp (Hypophthalmichthys nobilis) in the Mississippi River. Journal of Freshwater Ecology 11:241 – 243. United States Geological Survey. 2012. USGS current water data for the nation. URL: http://waterdata.usgs.gov/nwis/rt?

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Figure 3.1. Study area spanning North and South Dakota with the James, Vermillion, and Big Sioux Rivers in the eastern part of the states. Large barriers represent dams that presumable Asian carp cannot naturally pass. Standardized sampling sites are located throughout each river basin.

92

93

Figure 3.2. Northern most collections of silver and bighead carp adults. Adults are 3 years old or more and greater than 600 mm TL.

93

94

Figure 3.3. Northern most collections of silver and bighead carp juveniles. Juveniles are 1 - 2 years old and between 300 and 600 mm TL.

94

95

Figure 3.4. Juvenile silver carp collected in North Dakota on the James River below Jamestown Dam.

95

96

Figure 3.5. Northern most collections of age-0 silver and bighead carp. Age-0 carp were born the year of collection and are less than 300 mm TL.

96

97

Figure 3.6. Two age-0 silver carps collected in Shue creek, a tributary to the James River

97

98

Appendix 3.1. First collection records for adult, juvenile, and young-of year silver and bighead carps for North and South Dakota. Record Coordinates Life Stage

Species

Number Collected

State

Adult

Silver Carp

1

SD

Silver Carp

9

SD

Silver Carp

5

Bighead Carp Juvenile

Latitude (N)

Longitude (W)

Record Date

Reference

43.428253

-97.802089

2 August 2012

Current Study

Vermillion River confluence with the Missouri River

42.73362

-96.88918

16 August 2012

Current Study

SD

Big Sioux River confluence with the Missouri River

42.51066

-96.47855

17 August 2012

Current Study

1

SD

James River near Mitchell

43.70212

-97.96766

31 August 2010

Current Study

Silver Carp

12

SD

Shue Creek confluence with the James River

44.441792

-98.115286

27 September 2011

Current Study

Silver Carp

1

ND

James River below Jamestown Dam

46.931042

-98.708975

12 October 2011

North Dakota Game Fish and Parks

Silver Carp

15

SD

Vermillion River below East Lake Vermillion Dam

43.587875

-97.173394

11 August 2011

Current Study

Silver Carp

19

SD

Big Sioux River confluence with the Missouri River

42.51066

-96.47855

17 August 2012

Current Study

Bighead Carp

6

SD

Firesteel Creek confluence with the James River

43.70212

-97.96766

29 September 2011

Current Study

Bighead Carp

5

SD

Big Sioux River confluence with the Missouri River

42.51066

-96.47855

6 October 2011

Current Study

Location James River at Milltown

99

Age-0

Bighead Carp

4

SD

Vermillion River below East Lake Vermillion Dam

43.587875

-97.173394

11 August 2011

Current Study

Silver Carp

1

SD

Lake Byron

44.566761

-98.142442

2 November 2012

South Dakota Game Fish and Parks

Bighead Carp

2

SD

Lake Byron

44.566761

-98.142442

2 November 2012

South Dakota Game Fish and Parks

Silver Carp

2

SD

Shue Creek confluence with the James River

44.441792

-98.115286

29 September 2011

Current Study

Bighead Carp

1

SD

Big Sioux River below Sioux Falls

43.43759

-96.59552

20 October 2010

Current Study

100

CHAPTER FOUR: POPULATION DYNAMICS OF BIGHEAD AND SILVER CARP ON THE NORTHWESTERN FRONT OF THEIR NORTH AMERICAN INVASION

Abstract Invasive species are considered the second largest threat to native biodiversity, and ecosystem function and services. One pair of global invaders, bighead, Hypophthalmichthys nobilis, and silver carp, H. molitrix, (collectively referred to as Asian carps) have been introduced throughout the world, and threaten to invade three prairie stream tributaries to the Missouri River in the United States. Despite being one of the most endangered regions in North America, prairie ecosystems have historically shown signs of the biotic resistance hypothesis in which strong biotic interactions between native and nonnative species and the establishment of a nonnative species is regulated by the biotic interactions of native assemblages. Two hypotheses (biotic and abiotic resistance) may play a role in the impending invasion and potential establishment of Asian carps in eastern South Dakota prairie streams. There is a paucity of knowledge and understanding about Asian carps population characteristics and biology in North America. As such, we documented spatial and temporal trends in population characteristics (i.e., density, size structure, age, growth and condition) of Asian carps in three tributaries: Big Sioux, James and Vermillion. Finally, negative effects of Asian carps on native planktivores (i.e., gizzard shad, Dorosoma cepedianum, bigmouth buffalo, Ictiobus cyprinellus, and emerald shiner, Notropis atherinoides) were examined using condition (Fulton’s K). Three significant barriers were the most upstream locations within each river where Asian carps were collected and should act as barriers to further

101

natural spread. Overall, 469 silver carp and eight bighead carp were collected using boat electrofishing and mean catch-per-unit-effort of silver carp increased annually. The three rivers’ populations were similar in length frequencies. Silver carp growth was initially fast and slowed at later ages and overall was slower than Middle Mississippi River populations. Recruitment of silver carp was erratic with the 2010 year class dominating 91% of catches. Silver carp condition was also similar across rivers, seasons, and years. South Dakota silver carp were predicted to be lighter than the Gavins Point reach population of the Missouri River, and heavier than both the middle Mississippi and the Illinois River populations. Additionally, mean catch-per-unit-effort for bigmouth buffalo and emerald shiner decreased over the study period. Continued monitoring and research on this newly invading population of Asian carps will provide additional invaluable insight into complex invasive species, assist with understanding Asian carps population dynamics during and after an invasion, and expose the potential negative impacts Asian carps may be having on prairie stream ecosystems.

Key words: biotic resistance, environmental resistance, growth, invasion phase, planktivores, prairie streams, recruitment

Introduction Globally, invasive species are considered the second largest threat to native biodiversity (behind habitat loss; Wilcove et al. 1998) and ecosystem function and services (Vitousek et al. 1996; Sala et al. 2000; Sakai et al 2001; Pasari et al. 2013). Invasive species are also considered a significant component of global environmental

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change (Vitousek et al. 1996) as most countries have documented 100 – 10,000 nonnative species (Lodge 1993). Invasive species are common and increasing throughout a majority of North American freshwater environments and are expanding their ranges through natural and human-mediated dispersal (Lodge 1993; Vitousek et al. 1996; Rahel 2000) at an unprecedented rate (Pimm et al. 1995). For example, global trade and tourism expansion over the past 100 years has caused a 63% increase (approximately 60 new exotics) in invasive species in the Great Lakes over the past 60 years and the rate is accelerating (Vander Zanden 2005). Additionally, the common carp, Cyprinus carpio Linnaeus, 1758, is one of the most invasive freshwater fish species (Stuart and Jones 2006) and has established nuisance populations in all lower 48 United States (USGS 2010). The consequences of range expansion are poorly understood for most invasive species and the novel ecosystems they invade (Mitchell and Knouft 2009) as much research is conducted on a single species, usually after they have established and altered ecosystems (Kolar and Lodge 2002). Numerous studies have generalized about the negative impacts invasive species can have on native species which include, but are not limited to, competition, predation, and hybridization (Case 1990; Mooney and Cleland 2001; Gurevitch and Padilla 2004). Much research has been dedicated to predicting patterns of invasive species life history traits (e.g., recruitment) as well as the physical and biological traits of the invaded environment (e.g., species richness) to help explain the invasion potential of a species or ecosystem (e.g., Gido and Brown 1999; Kolar and Lodge 2002; Meador et al. 2003; Marchetti et al. 2004a, b). For example, species richness of an invaded ecosystem has been investigated as early as 1958 (Elton 1958) as a potential predictor of susceptibility

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to invasion, yet results remain contradictory (e.g., Gido and Brown 1999; Meador et al. 2003; Mitchell and Knouft 2009). These along with numerous other studies of invasive species were conducted on small geographic scales and were based on limited field data (Gurevitch and Padilla 2004). However, examining community and assemblage structure across larger geographic areas (Jeschke and Strayler 2005) is critical to identifying impacts of invasive species (Mack et al. 2000; Scott and Helfmann 2001; Jeschke and Strayler 2005; Simberloff et al. 2005; Poff et al. 2007; Leprieur et al. 2008) on aquatic ecosystems. Two global invaders, bighead, Hypophthalmichthys nobilis Richardson, 1845, and silver carp, H. molitrix Valenciennes, 1844, (collectively referred to as Asian carps) have been introduced intentionally and unintentionally throughout the world, mostly for aquaculture purposes (Kolar et al. 2007) as they are the most important aquaculture species in Asia and east-central Europe (Lieberman 1996). Bighead carp have invaded 74 countries and are reproducing in 19 and silver carp have invaded 88 countries and are reproducing in 23 (Kolar et al. 2007). Both species of Asian carps are currently reproducing in the United States (Papoulias et al. 2006; DeGrandchamp et al. 2007; Lohmeyer and Garvey 2009; Deters et al. 2013). Asian carps were originally introduced into the southern United States in aquaculture ponds in the early 1970’s where both species subsequently escaped and began migrating up the Mississippi River and into associated tributaries (e.g., Missouri River, Illinois River, Ohio River; Kolar et al. 2007). On the northwestern invasion front, Asian carps threaten to invade three tributaries to the Missouri River that support populations of many threatened and endangered fishes (Berry et al. 2007).

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Prairie ecosystems face disturbance from historic and present landuse practices and natural physiochemical and hydrologic fluctuations (Matthews 1988; Poff and Ward 1989). Despite being one of the most endangered regions in North America (Samson and Knopf 1994; Dodds et al. 2004), prairie ecosystems have historically shown signs of the biotic resistance hypothesis (Baltz and Moyle 1993). The biotic resistance hypothesis predicts that less speciose native communities have a higher probability of invasion than already saturated native communities (Herbold and Moyle 1986; Lodge 1993). Both common carp and grass carp Ctenopharyngodon idella Valenciennes, 1844, two invasive species, were historically, and still are, low abundance inhabitants of prairie rivers in eastern South Dakota and have not caused any measureable negative ecological impacts (Moyle et al. 1982; Moyle and Vondracek 1985; Hayer et al. unpublished data) despite causing significant harm in other regions (e.g., Fletcher et al. 1985; Roberts et al. 1995; King et al. 1997; Robertson et al. 1997; Gehrke and Harris 2001). A second hypothesis, the environmental resistance hypothesis, states that nonnative species are limited by their ability to tolerate the abiotic conditions of the novel ecosystems they invade (Moyle and Light 1996). As invasions are extremely unpredictable, it is unknown whether potential invaders (i.e., Asian carps) will invade eastern South Dakota tributaries and either become established in these already harsh prairie environments (Matthews 1988) and/or cause negative impacts to the native ecosystem. Both hypotheses (biotic and abiotic resistance) may play a role in the impending invasion and potential establishment of Asian carps in eastern South Dakota prairie streams. Limited data exists on the dynamic rate functions (i.e., recruitment, growth, and mortality) of Asian carps (Jennings 1988; Kolar et al. 2007), most often because

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estimating ages of these species is difficult (Schrank and Guy 2002; Kolar et al. 2007). Numerous hard structures (e.g., dorsal fin rays, pectoral fin rays, scales, saggital otoliths, cleithra, urohyal bones, and vertebrae; Kamilov 1985; Schrank and Guy 2002; Nuevo et al 2004; Williamson and Garvey 2005; Seibert and Phelps 2013) have been used and/or validated for accuracy and precision of age estimates. Regardless of method used to estimate ages, Asian carps can have rapid growth rates as bighead carp are capable of growing to 1800-2300 g in 4-5 years in Israel (Henderson 1978; Leventer 1987). Silver carp can reach 270 g (Waterman 1997) in their first year and can gain 45 g or more per month (Stone et al. 2000). In the Missouri River below Gavins Point Dam, just above the three prairie streams in this study, bighead carp ranged from 322 - 1200 mm total length and 400 - 1900 g (Wanner and Klumb 2009). Asian carps can also be long lived and have reached 8-10 years old in Lake Erie, Ontario (Johal et al. 2001; Morrison et al. 2004) and 7 years old in the lower Missouri River (Schrank and Guy 2002). There is a paucity of knowledge and understanding about Asian carp population characteristics and biology (Conover et al. 2007; Burgess and Bertrand 2008; Asian carp regional coordinating committee 2012; Coulter et al. 2013) in North America and many studies have focused on predicting future invasions (e.g., bioenergetics, ecological niche, habitat suitability; Chen et al. 2007; Herborg et al. 2007; Kocovsky, et al. 2012; Poulos et al. 2012), particularly into the Great Lakes (Cooke and Hill 2010; Kocovsky et al. 2012). As, such many are recognizing the potential contributions of basic research (e.g., behavioral, biological, ecological) to the study of invasive species (Sakai et al. 2001; Scott and Helfman 2001) and many state and government agencies are calling for general information about the population dynamics of Asian carps (e.g., Conover et al. 2007;

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Burgess and Bertrand 2008; Asian Carp Regional Coordinating Committee 2012). We documented spatial and temporal trends in population dynamics (i.e., density, size structure, age, growth and condition) of Asian carps. Finally, negative effects of Asian carps on native planktivores (i.e., gizzard shad, Dorosoma cepedianum Lesueur, 1818, bigmouth buffalo, Ictiobus cyprinellus Valenciennes, 1944, and emerald shiner, Notropis atherinoides Rafinesque, 1818; Mitchell and Knouft 2009) were examined using condition across time. This basic biological information will provide insights into two prolific invasive species and also provide awareness into what transpires during an active invasion phase. These data can help with predicting future invasions as eradication is still possible during the invasion phase but is unlikely or extremely difficult and costly after the establishment of a species (Kolar and Lodge 2002; Vander Zanden 2005).

Study Area The James, Vermillion, and Big Sioux rivers in South Dakota are three prairie tributaries which converge with an unchannelized section (Galat et al. 2005a) of the Missouri River just downstream of Gavins Point Dam (Figure 1). These warm-water rivers drain the Central Lowlands physiographic province in South Dakota (Galat et al. 2005b) and are characterized by low gradient streams of glacial origin (Hoagstrom et al. 2007). The upper portion of the James River has a gradient of about 0.02 m/km, and the lower portion has a gradient of about 0.05 m/km (Owen et al. 1981), making this river one of the lowest gradient rivers in the United States (Benson 1983). The James River (watershed area = 57,000 km2) extends 760 river kilometers (Rkm) from southeastern North Dakota through eastern South Dakota to its confluence with the Missouri River

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(Berry et al. 1993; Figure 1). In addition to one large impassable dam, approximately 230 low-head dams and rock crossings exist on the James River (Shearer and Berry 2003), potentially limiting fish migration and reproductive success during low hydrologic periods. However, most of these barriers are passable during flood events (Berry et al. 1993). The Vermillion River, the smallest basin (watershed area = 5,800 km2), extends 243 Rkm from the confluence of West and East Fork Vermillion rivers to its confluence with the Missouri River (Schmulbach and Braaten 1993; Figure 1). This river has one large dam (12.5 m high) on the East Fork Vermillion River, below East Lake Vermillion that acts as a barrier to fish passage. The Big Sioux River (watershed area = 23,325 km2) extends 470 Rkm from the Prairie Coteau of northeastern South Dakota to its confluence with the Missouri River at the South Dakota-Nebraska-Iowa border (Figure 1). This river has one natural set of waterfalls in Sioux Falls that may act as barriers to fish movement and 3 smaller low head dams (2.5 m high).

Methods Discharge Mean monthly discharge data (expressed as cubic meters per second, CMS) from 2008 – 2012 was downloaded from United States Geological Survey (United States Geological Survey 2013) gaging stations along the James (N = 5), Vermillion (N = 2) and Big Sioux rivers (N = 3) that corresponded with standard fish sampling locations used in this study (Figure 1). A one way analysis of variance (ANOVA) was used to characterize spatial (i.e., river: James, Vermillion, Big Sioux Rivers and site: N = 10) and temporal (i.e., year: 2008 – 2012 and season: spring, summer, fall) hydrologic variability in mean

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monthly discharge between 2008 and 2012. Additionally, a post-hoc Tukey’s t-test was used to illustrate which variables within the group that was being tested for differences (i.e., river, site, year, season) were different from one another. Mean annual discharge was also downloaded from the Huron gage on the James River to document current and historical trends (1949 – 2012). We assume that discharge at this gage is characteristic of all discharge at all gages within the study area.

Fish collection Standardized active and passive fish sampling gears were deployed seasonally (i.e., spring, summer, fall) from 2009 - 2012 at 10 sites on the James (N = 5), Vermillion (N = 2), and Big Sioux (N = 3) rivers (Figure 1). Multiple gears were selected to detect the entire fish assemblage (Drobish 2008; Guy et al. 2009). Additionally, many researchers have noted that Asian carps are extremely difficult to capture, (Stancil 2003; Williamson and Garvey 2005; Conover et al. 2007; Klumb 2007; Wanner and Klumb 2009) even in areas with high Asian carps concentrations. Therefore, multiple gears were deployed to determine effective sampling methods for Asian carps. Gears used included a combination of the following: pulsed DC boat electrofishing, hoop nets (hoop size: 1 meter; # hoops: 7; mesh size: 0.32cm x 0.61m), modified minifyke nets (4.5 m x 0.6m; 2 rectangular frames 1.2 m x 0.6m; 2 circular hoops 0.6 m diameter; 3mm ACE nylon mesh), and baited minnow traps (length = 42 cm; height = 19 cm; width = 22 cm; Purina dog food). All gears were deployed following the standard methods of Drobish (2008) and Guy et al. (2009). Standardized effort at each site included three ten minute runs for boat electrofishing, three hundred meter transects for backpack electrofishing, four hoop

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nets set overnight, four minifyke nets set overnight, and four minnow traps set overnight. All fish sampled were identified, counted, weighed (g) and total length was measured (mm). All fish species besides Asian carps and native planktivores (i.e., bigmouth buffalo, gizzard shad, and emerald shiner) were immediately released upon collection of measurements. After weighing and measuring fish, Asian carps were placed on ice and taken back to the laboratory at South Dakota State University where asterisci otoliths were extracted from each Asian carp. Asterisci otoliths (Seibert and Phelps 2013) were removed from 334 silver carp and 15 bighead carp using methods similar to those described by Schneidervin and Hubert (1986). Otoliths were cleaned using warm water, stored in plastic vials, and allowed to dry for at least two weeks prior to further processing. Additionally, we boat electrofished additional sites in response to angler and South Dakota Game, Fish, and Parks reports of Asian carps. These additional data were included in catch-per-unit-effort (CPUE) calculations and age and growth estimates for Asian carps.

Asian carp population dynamics Asian carps abundance was indexed using catch-per-unit-effort (CPUE), which was calculated as number per hour of electrofishing and number per net-night for passive gear. One-way ANOVA was used to test for differences in CPUE and percent of total catch by season (i.e. spring [May, June], summer [July, Augusta], fall [September, October]), year (i.e., 2009, 2010, 2011, 2012), river (i.e., James, Vermillion, and Big Sioux rivers), and site, and a post-hoc Tukey’s t-test was used to illustrate which variables within a group (e.g., season, year, river, and site) were different.

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Size structure of silver carp was examined with length-frequency histograms across time (i.e., seasonally). There were too few bighead carp collected to examine size structure. A nonparametric Kruskal-Wallis test for several samples was used to determine if length-frequency distributions for silver carp were different among rivers, sites, seasons, and years (Neuman and Allen 2007). Additionally, Fulton’s condition factor which relates fish weight to length was used to examine condition of Asian carps and native planktivores (bigmouth buffalo, emerald shiner, gizzard shad). Fulton’s condition factor (K) can be represented as: K = (W/L3)*105 where W equals weight (g) and L equals total length (mm). Condition was compared separately for silver carp and bighead carp, across rivers, sites, seasons and years using an ANOVA with a post-hoc Tukey’s t-test. After drying, one otolith from each individual was embedded in a two part epoxy mixture consisting of 5 parts Buehler® EpoxiCure® resin to one part Buehler® Epoxicure® hardener. One 0.5 mm cross section was cut out of each otolith using an Isomet (model # 11-1280-160) low speed precision saw mounted with a 0.3 mm diamond wafering blade (Buehler no. 11-4244). The otolith cross section was cut along a transect extending from anterior dorsal corner to the posterior ventral corner of the otolith. Each otolith thin-section was dried and glued to a glass microscope slide using cyanoacrylic cement. All otolith thin-sections were polished with 1,000 grit wetted sand paper and covered with immersion oil to enhance clarity prior to estimating ages. Otolith thinsections were viewed through a dissecting microscope (Olympus© SZX7 with a 1X-4 objective) using transmitted light. Ages were estimated from all thin-sectioned otoliths by

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an expert independent reader who has estimated ages from thousands of thin-sectioned and cracked otoliths from many different species. Silver carp growth was estimated using a von Bertalanffy growth curve (Isley and Grabowski 2007) represented by the equation: Lt = L∞[L-e –K(t-t0)], where Lt is length at time t, L∞ is the average maximum attainable size, K is the Brody growth coefficient, and t0 is the time coefficient at which fish length is theoretically 0 (von Bertalanffy 1938).

Potential deleterious effects Catch-per-unit-effort and Fulton’s condition factor (K) were also calculated for native planktivores (i.e., gizzard shad, bigmouth buffalo, emerald shiner) and compared among rivers, sites, years, and seasons with one-way ANOVA and post-hoc Tukey’s ttest to test for any differences that may be caused by Asian carp presence. To investigate differences in condition of native planktivores by river, linear regressions of weightlength data were compared after log10 transformation of both length and weight. The slope and intercept values were examined first, then the corresponding 95% confidence intervals for each river were inspected. Significant differences were identified by nonoverlapping 95% confidence intervals and slope and intercept values that fall outside the confidence intervals (Pope and Krus 2007).

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Results Mean annual discharge for the James River during the study period (2009 – 2012) consisted of the largest, second largest, and fourth largest discharge values on record dating back to 1949 (Figure 2). Again, we assume these patterns are consistent across the study area. Additionally, mean monthly discharge for all rivers combined differed annually (F4, 541 = 30.75; P < 0.0001) with 2010 (971.0 m3s) and 2011 (1100.4 m3s) being significantly larger than 2008 (170.7 m3s), 2009 (433.3 m3s) and 2012 (170.6 m3s; Figure 2). Mean monthly discharge was also different across all study rivers (F2,541 = 32.47, P < 0.0001), with the James River having the largest mean monthly discharge across the study period (881 m3/s) followed by the Big Sioux River (524 m3/s) and then the Vermillion River (109 m3/s; Figure 3). By river, discharge followed a longitudinal pattern, decreasing from upstream to downstream sites (F9, 541 = 17.19, P < 0.0001; Figure 4) as discharge was significantly different at each site within the respective watershed. Additionally, each season was unique (F3,538 = 8.49; P < 0.001) with spring having the highest mean discharge (965 CMS) followed by summer (708 CMS), and fall (377 CMS).

Northwestern distribution front Three significant barriers, a 33.5 meter high dam on the James River in North Dakota, a 12.5 meter high dam on the Vermillion River in South Dakota, and a series of natural falls on the Big Sioux River in South Dakota are the most upstream locations within these basins (Figure 1) where bighead and silver carps have been collected.

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Population dynamics A total of 21,409 fish belonging to 53 species were collected over the course of this study. Thirty-nine boat electrofishing hours were spent collecting fish between fall 2009 and fall 2012. Passive gear effort included 130 mini-fyke net nights, 130 minnow trap nights and 131 hoop net nights. No Asian carps were collected using passive gears, and all further bighead and silver carp results were from boat electrofishing catches. Only eight bighead carp were collected during the study interval using standardized methods (Figure 4) although our observations along with numerous anglers and the South Dakota Game Fish and Parks (SDGFP) personnel, suggest localized concentrations of bighead carp below the numerous low-head dams, which were too dangerous to sample. We did however, collect additional bighead (N=15) and silver carp (N=116) at non-standardized sites. Overall, 469 silver carp were collected using boat electrofishing (Figure 4). Boat electrofishing catches also included any Asian carps that may have jumped in the boat while electrofishing. Forty-nine percent of silver carp were collected from the James River followed by 32 % in the Vermillion River and 19% from the Big Sioux River. Mean CPUE for silver carp did not differ spatially (i.e., river) or seasonally (i.e., spring, summer, fall) although the highest mean CPUE occurred in the middle-lower James River (87 silver carp per hour) in 2012 at a non-standardized site. However, mean CPUE was highest at the lower James site and the Vermillion confluence with the Missouri River (31 and 24 silver carp per hour, respectively) across all years. Mean CPUE also increased each year at most sites (Table 1) with highest mean CPUE in 2012 (Figure 5; 36 silver carp/hour). Bighead

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carp CPUE did not change across years as catches remained small throughout the study time frame (Figure 5). The most bighead carp (N = 3) were collected at the James River confluence with the Missouri River. Ages were estimated from 334 of 469 silver carp collected and 15 of 23 bighead carp collected. Silver carp age estimates ranged from 0 – 5 and bighead carp age estimates ranged from 0 – 3. The largest silver carp (5 years old) was 823 mm and weighed 8219 g and the largest bighead carp was 1001 mm and weighed 8219 g. Silver carp length frequency histograms revealed the longest fish occurred in summers and mean length in 2009 was the longest (838 mm) followed by 2012 (507 mm; Figure 6). Only three adult silver carp were collected from the James River confluence in 2009. Length-frequency distributions were similar among rivers (F 2, 394 = 0.66, P = 0.52). Further examination of size frequency histograms illustrated fast growth of silver carp seasonally, even growing in the winter months (Figure 6). Silver carp growth was initially fast reaching almost 400 mm by the end of their second growing season (Figure 7). This initial fast growth was followed by slower growth at later ages (Figure7). Growth rates of silver carp were fastest at the confluence sites, and declined longitudinally within each river (F12, 384 = 8.75, P < 0.0001). Growth rates of silver carp in South Dakota were faster than the Amur River and Gobindsagar Reservoir but slower than the middle Mississippi River (Figure 8).

The 2010 year class dominated most of the silver carp catches (Figure 9) indicating they may have erratic recruitment. In fact, the 2010 year class increased in density annually and comprised 91% of silver carp collected (Figure 9). There were

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numerous missing year classes including ages 1-5 in 2010 and age-0 in 2012 (Figure 9). Additionally, numerous weak year classes were present in 2011 and 2012. Silver carp condition (Fulton’s K) was relatively stable across rivers (Figure 10), season, and years. Additionally, 95% confidence intervals of length-weight regression parameters (slope and intercept) overlapped, indicating similar condition among rivers (Table 2). Based on length-weight regressions, an 800 mm silver carp in the Big Sioux River is predicted to be smaller (6150 g) than the Gavins Point reach population of the Missouri River (6628 g; Wanner and Klumb 2009), and larger than both the Middle Mississippi (5477 g) and the Illinois River populations (5856 g; Williamson and Garvey 2005; Irons et al. 2011; Figure 11).

Potential deleterious effects Overall planktivore CPUE ranged from 0 – 179/hour, gizzard shad ranged from 0 – 1786/hour, bigmouth buffalo ranged from 0 – 36/hour, and emerald shiner ranged from 0 – 129/hour. There were no differences in CPUE across rivers (F2, 66 = 0.84; P = 0.44) or annually (F3,65 = 2.36; P = 0.08) for native planktivores (Figure 12). Mean CPUE for bigmouth buffalo and emerald shiner was highest in 2009 and decreased each subsequent year (Figure 14) as silver carp populations were increasing. Gizzard shad exhibited a significant difference in condition among seasons (F2, 649 = 3.05; P = 0.048) with highest values in the spring (K = 1.18) and there were no differences in bigmouth buffalo or emerald shiner condition through the study period or region.

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Discussion During this study, the James River recorded three of the top five highest mean annual discharge events since 1949 (United States Geological Survey 2013). These flood conditions overtook most of these lowhead dams and may have allowed for and facilitated the unobstructed movement and colonization upstream of silver carp into the James, Vermillion, and Big Sioux rivers. Furthermore, silver carp were documented to migrate ~700 km from the confluence of the James River and the Missouri River over most all of the lowhead dams with little resistance to the Jamestown Dam in North Dakota. During these high discharge events, the James River may have provided natural free-flowing conditions characteristic of the Missouri River prior to damming (Berry and Galat 1993; Berry et al. 1993) and as such in these high water years, the James is utilized by large riverine and migratory fish (e.g., blue sucker Cycleptus elongatus Lesueur, 1817, western silvery minnow Hybognathus argyritis, paddlefish Polyodon spathula Walbaum, 1792; Shearer and Berry 2003). In fact, paddlefish were seen in the middle James River near Mitchell, SD and a few blue suckers were collected in the middle to lower James River near Olivet, SD in 2010. Historically, the blue sucker and western silvery minnow were detected in the James River in the early 1990’s, a time period of high discharge events (Morey and Berry 2003; Shearer and Berry 2007). Additionally, gizzard shad, another large river fish, were collected in the middle-upper watershed of the James and had high densities in 2011 (e.g., CPUE = 345 per hour). The high discharge events between 2009 and 2011 likely facilitated the invasion of silver carp further upstream into the three Missouri River tributaries as lowhead dams were over-topped.

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Density (i.e., CPUE) and percent of silver carp catches increased to 45% across years as we expected based on the population dynamics of other invasive species (e.g., zebra mussel Dreissena polymorpha Pallas, 1771) as well as Asian carp (Irons et al. 2011). For example, the Illinois River showed an exponential increase in silver carp catches during invasion (Chick and Pegg 2001) and percent of silver carp catches reached 51% (Sass et al. 2010). Other invasive populations seem to peak in abundance after a rapid and/or exponential increase in density or biomass (Irons et al. 2007; Valery et al. 2008; Irons et al. 2011) which is followed by a leveling off corresponding to some carrying capacity (Roughgarden 1998). However, the silver carp in rivers in this study may not have peaked in abundance as abundance was still increasing during the final year of this study. As such, they can be considered to still be in the invasion/colonization stage. Additionally, 91% of silver carp collected were from the 2010 year class indicating that natural reproduction is not augmenting the population in these rivers. Instead, silver carp must be continually immigrating from the Missouri River into the tributaries. The majority of the population of these tributaries may become reproductively viable in 2013, as age at maturity in other invasive populations is 3+. As a result there may be an augmentation in the population in 2013 or 2014 not only from immigration, but from reproduction. Erratic reproduction is not only characteristic of these rivers, but of others (e.g., Illinois River) as recent declines in Asian carps populations is attributed to sporadic reproductive events (Irons et al. 2011) which occurred in 2000 and 2003 in the Illinois River. We used asteriscus otoliths, which in cyprinids are the largest of the three pairs of otoliths (Brown et al. 2004; Seibert and Phelps 2013) for estimating ages of both bighead

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and silver carp and got clear images; however, there is no standard or validated structure that has been shown to be the best for estimating ages of Asian carps. Some studies used otoliths to estimate ages of Asian carps, but were unsuccessful in estimating ages for bighead carp as the image was considered cloudy and unsuitable (Kamilov 1985; Nuevo et al. 2004). Other studies however used saggital otoliths and were able to age silver carp larvae and adults successfully (Schrank et al. 2001; Seibert and Phelps 2013). A validation study on another large cyprinid, the common carp, concluded that otoliths are suitable to determine age (Brown et al. 2004). Additional structures have also been used to estimate ages of Asian carps (Johal et al. 2001; Schrank and Guy 2002; Williamson and Garvey 2005). For example, the pectoral fin ray had 60% agreement between two readers on Middle Mississippi River silver carp (Williamson and Garvey 2005) and urhoyal bones, cleithra, and scales were suitable structures for estimating ages in Gobindsagar Reservoir, India (Johal et al. 2001). Structures may vary geographically as is apparent with the numerous different structures used across a broad geographic range. Finding a suitable structure by experimenting with numerous structures and methods of structure preparation to estimate ages is extremely important in understanding changes in population dynamics, such as growth rates and mortality, over time (Irons et al. 2011). Growth rates for South Dakota populations were faster for all ages than two nonNorth American populations (Amur River; Nikolskii 1961; and Gobindsagar Reservoir; Tandon et al. 1993; Williamson and Garvey 2005), especially the Amur River in Russia which is considered a native population. There could be numerous abiotic or biotic reasons for this discrepancy, but we think that it could suggest three things. First, there may be less competition from native planktivores in South Dakota rivers (e.g., bigmouth

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buffalo, emerald shiner, gizzard shad) than in other North American populations and in their native range (sharpbelly, Hemiculter leucisculus Basilewsky 1855; Zhang et al. 2007) as abundance of native planktivores, especially the potentially large bigmouth buffalo are low (Hayer unpublished data). Second, larger fish could be remnant of a typical invasive species population that has faster growth in the novel environment than in the native environment (Kolar and Lodge 2002). Lastly, there may be unlimited food resources for the newly invading Asian carp in South Dakota as they are early in the invasion and may not yet be abundant enough to deplete the plankton resources and alter the fish community. South Dakota fish, however, are smaller than the Middle Mississippi River silver carp out to age six (Williamson and Garvey 2005). The oldest fish we caught in this study was five, so making inferences past this age is based on the growth of smaller fish. Larger fish tend to slow in growth (Williamson and Garvey 2005). Smaller fish length-atage indicates biological or environmental resistance in South Dakota rivers that could be limiting their growth. Perhaps, the high discharge throughout most of the study time frame limited growth, or the typical harsh environmental conditions of the prairie stream (e.g., high turbidity, variable discharge, temperature fluctuations) itself are limiting growth and population size. There were differences in growth rates and condition among, the Gavins Point reach of the Missouri River, Middle Mississippi River, Illinois River, and South Dakota Missouri tributaries silver carp populations (Williamson and Garvey 2005; Irons et al. 2011); however these populations were in different invasion stages. The Illinois River predicted length-weight regression included silver carp collected before and after the

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peak in population biomass (i.e., 2000–2006) and the Missouri River regression represented an already established population with successful recruitment. The James River fish are still in the invasion stage as their densities continue to increase, reproduction has not been confirmed, and recruitment is erratic. The Missouri River tributary fish may be encountering environmental resistance from flooding or biotic resistance in finding novel habitats and additional energy spent on colonization and migration which all may weaken their condition. The Illinois River populations were probably undergoing intense intraspecific competition for food resources or density dependence during peak abundances (Schrank et al. 2003; Sampson et al. 2009; Irons et al. 2011).

Potential deleterious effects Emerald shiner displayed marked declines in density across the short study period. Emerald shiner is an important forage fish for a popular sportfish, the walleye. Emerald shiner is a planktivore that consumes algae at small sizes and consumes copepods and cladocerans at larger sizes (Fuchs 1967) which may have led to competition with planktivorous Asian carps (Irons et al. 2007; Kolar et al. 2007). However, there was no indication of reduced condition in emerald shiner across the time period. Additionally, some emerald shiner populations are characterized by high mortality and fluctuating populations (Fuchs 1967). On the other hand, emerald shiner could have been flushed down into the Missouri River during high discharges between 2009–2011 (US Geological Survey 2013), however this species in known to be plastic and adaptable to a variety of ecological conditions (Cambell and MacCrimmon 1970).

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Thus, future monitoring should examine emerald shiner populations to see if declines continue, or if they are a result of variable recruitment and/or movement. Bigmouth buffalo abundance also decreased during the study period. Bigmouth buffalo are not a common species to these rivers (Dieterman and Berry 1998; Shearer and Berry 2003) so any adverse impacts contributed to Asian carps could be detrimental to the population. However, condition did not change indicating that there may be a lag time between first colonization of Asian carps and measurable deleterious effects on native species as a result of invader characteristics and resistance and/or characteristics of the novel community (Valery et al. 2008). Further monitoring is necessary to quantify any negative impacts, such as reduced condition, growth, and increased mortality, attributed to Asian carps.

Summary Asian carps have invaded and become widely distributed in three northern prairie streams that are tributaries to the Missouri River. It is difficult to diagnose the stage of the Asian carps invasion in South Dakota (e.g., dispersal, colonization, establishment, self-sustaining), because silver carp density (silver carp) was still increasing annually at the end of this study. We have not confirmed reproduction in these basins; however, the presence of age-0 silver carp in the middle to upper James River, approximately 375 Rkm from the confluence of the James River with the Missouri River, suggests reproduction may be occurring within the James River. However, recruitment was erratic with the 2010 year class dominating each population, the 2011 year class was weak, and the 2012 year class was missing. 2010 recorded the second highest mean annual discharge on

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record since 1949. This year also had numerous peaks in discharge across the year which silver carp may use as a spawning cue (Schrank et al. 2001; Lohmeyer and Garvey 2009). As a result of the high discharge, low head dams in the James and Big Sioux Rivers were over-topped facilitating upstream migration. This year produced the only successful silver carp year class during this study. Recruitment and reproduction factors are not well understood across their range, native or nonnative (Irons et al. 2011); however it is suspected that discharge plays a major role in the survival of larvae (Costa-Pierce 1992; Schrank et al. 2001; Lohmeyer and Garvey 2009). The majority of fish from the 2010 year class should be reaching reproductive potential in 2013 (e.g., age 3+; Williamson and Garvey 2005). If suitable spawning habitat exists in these rivers, we should expect a pulse of reproductive effort, both from immigration and reproduction, assuming abiotic and/or biotic resistance does not overshadow the reproductive capabilities of Asian carps. Asian carps have the classic characteristics of an invasive species and this study is one of the few that has captured the leading edge of an invasion. Despite the harsh conditions in prairie streams, silver carp density increased; however, bighead carp densities have not. This may suggest some biotic or abiotic resistance to bighead carp invasion that may not affect silver carp. The James River appears to have suitable habitat for sustaining silver carp populations. For example, low velocity areas above lowhead dams may facilitate the upstream migration of silver carp by enhancing phytoplankton production and increasing foraging opportunities (Williamson and Garvey 2005). These shallow prairie streams may not be suitable habitat for bighead carp as they prefer deeper water (Kolar et al. 2007). Biotic resistance was able to maintain low abundance of

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common carp and grass carp, however, these two invasive species combined with bighead and silver carp, along with record high discharge may have undercut the environmental resistance of these ecosystems allowing for silver carp colonization. Continued monitoring and research on this newly invading population of Asian carps will provide additional invaluable insight into complex invasive species, assist with understanding Asian carps population dynamics during and after an invasion, and expose the potential negative impacts Asian carps may be having on prairie stream ecosystems. In conclusion, we have documented an invasion of silver carp into eastern South Dakota Missouri River tributaries from the onset. Each year silver carp densities increased significantly, yet bighead carp numbers remained consistently low. Both silver and bighead carp could be encountering abiotic and biotic resistance that may limit the impact these species will have on these naturally harsh prairie ecosystems. For example, there was only one strong year class suggesting abiotic resistance against reproduction and recruitment. As a result of both abiotic and biotic resistance, bighead carp may exhibit the same population characteristics of the common carp and grass carp and have minimal impacts on the native ecosystem in eastern South Dakota prairie streams.

Acknowledgements We thank the South Dakota Department of Game, Fish and Parks and the United States Fish and Wildlife Service Missouri River Research Office for their continued support of this project and specifically M Smith and G VanEeckhout for continued updates on Asian carps within North and South Dakota. Additionally we thank J Howell for assistance with field work. Funding for this research was provided to KN Bertrand,

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BDS Graeb, RA Klumb, and C-A Hayer through Study 1516 of the Dingell-Johnson Sportfish Restoration Act administered through the South Dakota Department of Game, Fish and Parks. Our good friend, colleague and mentor, Rob Klumb, was an integral part of this project since its inception, but he was tragically killed while working for the United States Fish and Wildlife Service.

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Figure 4.1. Study area in eastern South Dakota showing ten standardized sites on the James, Vermillion, and Big Sioux Rivers.

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Figure 4.2. Mean annual discharge (CMS) between 1949 and 2012 with the 25th and 75th percentiles, the median, and the study period marked. Discharge is displayed in ascending order and was downloaded from USGS site # in Huron, South Dakota.

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Figure 4.3. Mean monthly discharge (CMS) downloaded from the nearest US Geological Survey water gage to a sampling site for the James (A), Vermillion (B), and Big Sioux (C) rivers in eastern South Dakota between 2008-2012.

Number collected

142

180 160 140 120 100 80 60 40 20 0

Bighead carp

Silver carp

F09 SP10 SU10 FA10 SP11 SU11 FA11 SP12 SU12 FA12 Sampling period Figure 4.4. Bighead and silver carp catches in South Dakota tributaries across sampling seasons, spring (SP), summer (SU), fall (FA) from 2009-2012.

CPUE (#/hour)

143

50 45 40 35 30 25 20 15 10 5 0 2009

2010

2011

2012

Year Figure 4.5. Mean (±1 standard error) catch-per-unit-effort (#/boat electrofishing hour) for silver carp between 2009 and 2012 in the Big Sioux, James and Vermillion rivers, South Dakota.

144

Figure 4.6. Length-frequency distribution for silver carp across seasons for 2011 and 2012.

145

Total length (mm)

Mean length at age 900 800 700 600 500 400 300 200 100 0

lt = 1441.7[1-e-0.1032(t + 2.1008)] F3, 338 = 9781; P < 0.0001 0

1

2

Age

3

4

5

Figure 4.7. Mean length-at-age (squares) and raw data (diamonds) for silver carp collected in South Dakota tributaries 2009-2012. A von Bertalanffy growth curve was fit to the data and the equation was significant.

Total length (mm)

146

Middle Mississippi River

South Dakota tributaries

Amur River

Gobindsagar Reservoir

1000 900 800 700 600 500 400 300 200 100 0

1

2

3

4

5 Age

6

7

8

9

Figure 4.8. Predicted length at age for South Dakota tributaries, Middle Mississippi River (Williamson and Garvey 2005), the Amur River, Russia (Nikolskii 1961), and Gobindsagar Reservoir, India (Tandon et al. 1993).

147

Figure 4.9. Age frequency histograms for silver carp collected in South Dakota tributaries 2010-2012 (A-C). Note that y-axis scales are not the same.

148

Figure 4.10. Condition (K) of native planktivores, bigmouth buffalo Ictiobus cyprinellus, gizzard shad Dorosoma cepedianum and emerald shiner Notropis atherinoides versus total length for the James (A), Vermillion (B) and Big Sioux (C) rivers.

149 25000

Weight (g)

20000

James

Vermillion

Big Sioux

Gavins Point

Interior Highlands

Illinois River

15000

10000

5000

0

0

200

400

600 800 Total length (mm)

1000

Figure 4.11. Length-weight regression of silver carp collected in South Dakota tributaries (James, Vermillion and Big Sioux rivers), the Missouri River below Gavins Point Dam, the Interior Highlands portion of the Missouri River, and the Illinois River.

1200

150

Figure 4.12. Mean catch-per-unit-effort (±1 standard error) for native planktivores, bigmouth buffalo Ictiobus cyprinellus, gizzard shad Dorosoma cepedianum and emerald shiner Notropis atherinoides, across years (2009-2012) in South Dakota tributaries.

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Table 4.1. Mean catch-per-unit effort by year (2009-2012) for standardized boat electrofishing on the James, Vermillion, and Big Sioux rivers. Numbers in parentheses represent standard deviations. River James

Site Location

2010

2011

2012

Confluence

0

1.36 (2.72)

2.88 (0.81)

23.14 (13.13)

Lower-middle

0

.

11.65 (12.53)

76.73 (17.54)

Middle

0

0

22.00

2.62

Middle-upper

0

0

0

3.07(2.79)

Upper

0

0

0

.

0

0

22.50 (7.77)

58.43 (48.24)

Middle

.

.

.

.

Confluence

0

0.44 (.76)

20.62 (27.31)

38.01 (30.46)

Middle

0

1.26 (1.79)

0

0.99 (1.40)

Middle-upper

0

0

0

.

Vermillion Confluence

Big Sioux

2009

152

Table 4.2. Length-weight regression for silver carp collected in the James, Vermillion, and Big Sioux Rivers in eastern South Dakota along with r2 and 95% upper and lower confidence intervals (CI) around the parameter estimates. Length-weight equations were also gathered for the Missouri River proper, the Middle Mississippi River proper and the Illinois River, a tributary to the Mississippi River. Predicted fish weights (g) for fish that are 450 mm and 800 mm are also given in the table for comparison.

River

L-W Regression Equation

Intercept

Slope 95%

450

800

95% CI

CI

mm

mm

r2

Reference

Missouri River Tributaries log10 weight = -5.26 + 3.11(log10 James

length) log10 weight = -4.82 + 2.90(log10

Vermillion

length) log10 weight = -5.53 + 3.21(log10

Big Sioux

length)

0.96 5.43

-5.10

3.05

3.17

981

5869

-4.47

2.82

3.07

748

3971

-5.14

3.07

3.36

970

6150

0.98 5.16 0.98 5.91

Missouri River log10 weight= -6.92 + 3.7 (log10 Gavins Point

length)

Wanner and Klumb 0.97 0.21

0.59

788

6628

2009

153

Interior

log10 weight = -5.35 + 3.13 (log10

Highlands

length)

Wanner and Klumb 0.93 0.76

0.21

900

5453

2009

Mississippi River log10 weight = -5.29 + 3.11(log10 Middle

length)

Williamson and Garvey 0.81

915

5477

2005

0.99

972

5856

Irons et al. 2011

log10 weight = -5.29 + 3.12(log10 Illinois River

length)

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CHAPTER FIVE: ASSEMBLAGE AND TROPHODYNAMIC STRUCTURE OF THREE PRAIRIE RIVERS DURING THE INITIAL INVASION OF BIGHEAD HYPOPHTHALMICHTHYS NOBILIS AND SILVER CARP H. MOLITRIX

Abstract: Invasive species are key effectors of global change by altering community composition and the interactions and processes that occur within recipient ecosystems. Knowledge of impacts on community structure and food webs across wide spatiotemporal scales is essential to recognize invasion effects on aquatic communities. The northern Great Plains of the United States, one of the most endangered regions in North America, represents the northwestern front of a pair of global invaders, bighead carp Hypophthalmichthys nobilis and silver carp, H. molitrix. Post-invasion research in the Mississippi River Basin suggests that their expansion into the Missouri River and its tributaries could have wide ranging ecological and economic consequences by altering aquatic communities, particularly native planktivores through middle out and cascading effects. This study examines assemblage structure and food web consequences of bighead carp and silver carp invasion at the onset in three Great Plains rivers. We quantified invasive species-driven ecological changes in assemblage structure and tested whether those changes corresponded to changes in fish trophic position and river food web structure using stable isotope and diet analysis. We show that silver carp used a variety of food resources which may enhance their invasion success as they adapt to the fluctuating abiotic and biotic conditions of these harsh prairie streams. Isotope analysis and diet analysis reveal trophic overlap between bighead and silver carp and native planktivores. Results from this study indicate negative impacts may already be occurring between

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silver carp and emerald shiner, and there is potential for trophic competition between silver carp and gizzard shad. Keywords: stable isotope analysis, Great Plains streams, diet analysis, food web dynamics, invasive species, native species

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Introduction Biological invasions can cause profound economic and ecological damage to recipient systems. They are key factors of global change as they can modify not only community composition, but also ecosystem processes (e.g., competition, nutrient cycling; Vitousek 1990; Sala et al. 2000; Sakai et al. 2001). Herein we refer to invasive species as a species that has colonized, increased to become abundant in the novel ecosystem and has had deleterious effects on the recipient community in other geographic ranges where it is non-native. Invasive species effects are particularly most severe in freshwater environments (Moyle 1999; Ricciardi and MacIsaac 2000) as aquatic species are disproportionately more threatened than terrestrial species, because of the intensive use of water by humans and environmental degradation and modification (Moyle and Light 1996; Rahel 2002; Dudgeon et al. 2006). As a result of natural and humanmediated dispersal, invasive species are rapidly expanding their ranges (Vitousek et al. 1996; Moyle 1999; Rahel 2000). The consequences of range expansion for the individuals, populations, communities and ecosystems they invade (Nilsson et al. 2012; Bohn et al. 2004) are poorly understood for most invasive species, however they have been linked on the individual and population level to the decline or extinction of many native species (Wilcove and Bean 1994; Vitousek et al. 1997; Wilcove et al. 1998) within invaded areas as well as to the alteration of energy flow and nutrient fluxes (Simon and Townsend 2003) at the ecosystem level. Information on the influence of invasive species on community structure and food webs across wide spatial and long temporal scales is necessary (Lazzaro 1987; Ibanez et al. 2014) for increased understanding into dynamic and diverse invasion ecology.

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Invasive species individual, community and ecosystem effects have most commonly been studied post-invasion, after recipient ecosystems have changed (Williamson 1999; Bohn et al. 2004). Researching an invasive species at the edge of range expansion permits real-time analysis of the direction and rate of change in energy fluxes as a result of the new invaders and may help to improve our ability to counteract invasions and associated negative impacts on aquatic ecosystems (Mooney and Hobbs 2000; Bohn et al. 2004; Ibanez et al. 2014). The northern Great Plains of the United States represents the northwestern front of two global invaders, the bighead carp Hypophthalmichthys nobilis Richardson, 1845, and silver carp, H. molitrix Valenciennes, 1844, hereafter referred to as Asian carps. The Great Plains region is one of the most endangered in North America (Matthews 1988; Samson and Knopf 1994; Dodds et al. 2004) as it faces disturbance from historic and present landuse practices as well as natural physiochemical and hydrologic fluctuations (Matthews 1988; Poff and Ward 1989; Hayer Chapter 2). Asian carp populations were first detected in the Missouri River below Gavins Point Dam in 1998 and 2003, (Klumb 2007; Kolar et al. 2007) respectively and their colonization and establishment into Missouri River tributaries below this barrier was unknown prior to this study. Extensive research on Asian carp populations has been conducted in their native range (e.g., rivers in west Asia and Amur River in China), where both species are prized food fish, and the results of this research, conducted in enclosure experiments, aquaculture ponds and in the wild, serve as a predictive framework for their potential effects in novel communities and ecosystems (Chen et al. 2007; Poulos et al. 2012). Asian carps are filter feeders (Lazarro 1987; Fukushima et al. 1999; Xie and Yang 2000).

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Bighead carp are predominantly zooplanktivorous and silver carp are predominately phytoplanktivorous although both species exhibit prey switching (e.g., zooplankton to phytoplankton, phytoplankton to zooplankton) when resource biomass shifts (Bitterlich and Gnaiger 1984; Spataru and Gophen 1985; Opuszynski et al. 1991). Asian carps, depending on their biomass, can have a profound effect on the structure and composition of the plankton communities (Domaizon and Devaux 1999; Xie and Yang 2000). For example, in enclosure experiments, zooplankton was reduced in high silver carp biomass treatments, which resulted in a decrease in zooplankton grazing on nanoplankton and caused a subsequent or cascading increase in the production of small algae (Opuszynski 1979; Domaizon and Devaux 1999). Asian carps can also alter water quality, although effects are often contradictory (Kolar et al. 2007). Silver carp presence in experimental ponds or cages has increased, decreased, or had no effect on nutrient concentrations (Opuszynski 1980; Starling 1993; Henderson 1978; Matyas et al. 2003), decreased or increased algal biomass (Leventer 1987; Lieberman 1996; Lu et al. 2002; Opuszynski 1981; Spataru et al. 1983; Milstein et al. 1985), and decreased dissolved oxygen in the water (Vybornov 1989). Discrepancies in water quality results may be the result of the ability of Asian carps to acclimate to novel surroundings (e.g., Mississippi River basin) as they have thrived (e.g., high population densities) in conditions different than those found in their native waters (Kocovsky et al. 2012; Coulter et al. 2013). Although Asian carps are considered a riverine species they and are colonizing and surviving in a plethora of habitats (e.g., lotic, lentic, reservoir; Spataru and Gophen 1985; Domaizon and Devaux 1999; Cooke and Hill 2010; Tan et al. 2011).

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Post-invasion research on Asian carps in parts of its introduced range, the Mississippi River Basin suggest that their expansion into the middle Missouri River and its tributaries could have wide ranging ecological and economic consequences by changing aquatic communities, particularly native planktivores (i.e., bigmouth buffalo, Ictiobus cyprinellus, emerald shiner, Notropis atherinoides, gizzard shad, Dorosoma cepedianum and paddlefish, Polyodon spathula; Schrank et al. 2003; Williamson and Garvey 2005; Irons et al. 2007; Sampson et al. 2009). Additionally, competition with native planktivores (Lohmeyer and Garvey 2009), especially gizzard shad and emerald shiner, could affect the forage base quality and/or quantity for adult walleyes. The subsequent competition could result in a decrease in the population density of walleye recruitment and/or condition and lead to an adverse effect on an important recreational fishery (Groen and Schmulbach 1978; Kocovsky et al. 2012). Interspecific competition between Asian carps and native planktivores has been well documented in the Mississippi River basin (Schrank et al. 2003; Irons et al. 2007; Sampson et al. 2009). For example, gizzard shad and bigmouth buffalo body condition declined after Asian carps invaded the Illinois River (Irons et al. 2007), paddlefish displayed decreased growth in the presence of bighead carp in experimental ponds (Schrank et al. 2003) and there was dietary overlap between gizzard shad and Asian carps in the Illinois and Mississippi rivers (Sampson et al. 2009). Thus through middle out and cascading effects (Lu et al. 2002), Asian carps could restructure Missouri River tributary community composition and structure of food webs. Conducting a “natural experiment” (Diamond 1983; Sakai et al. 2001) on a broad spatiotemporal scale provides insights into how a new species invades novel ecosystems

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and what initial impacts they may have on ecosystem functioning (Lodge et al. 1998; Chen et al. 2007; Ibanez et al. 2014). By researching a population at the edge of an invasion range during the early phase of the invasion process, information can be gained about future invasion processes and characteristics such as: time to establishment and sustainability, propagule pressure necessary for increasing abundance or rate of spread, novel ecosystem characteristics and vulnerability (e.g., resiliency), characteristics of the novel biotic community (e.g., species diversity), lag times to perceived negative impacts or abrupt abundance increases, and species invasiveness (Kolar and Lodge 2001; Sakai et al. 2001; Marchetti et al. 2004; Vander Zanden 2005; Poulos et al. 2012). Depending on which research is read, there are three phases in the invasion process: an initial “establishment phase” with small populations, little spread and little to no intraspecific competition (e.g., Great Plains streams), an “expansion phase” with exponentially increasing populations, increased spread rates and intraspecific competiton begins to intensify, and a “saturation phase” with reproduction occurring to augment the population until a carrying capacity is reached (e.g. Illinois River; Bohn et al. 2004; Arim et al. 2006; Ibanez et al. 2014). These three phases can usually be followed by an “impact phase” (Kolar and Lodge 2002) where there is a replacement of native species (Allendorf and Lundquist 2003) and a shift in the food web of the novel invaded community. This shift can take years or even decades to occur (e.g. time lag) before deleterious effects become visible and measureable (Sakai et al. 2001). We believe that Asian carps in the middle Missouri River tributaries are in the initial phases of “establishment” and “expansion” as reproduction has not been confirmed (Hayer et al. 2014; Chapter 4).

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During these initial phases, invasive populations exhibit population growth (Sakai et al. 2001) characteristic of the silver carp within this region (Hayer et al. 2014 Chapter 4). Competition among Asian carps and native planktivorous fishes in tributaries to the middle Missouri River tributaries could result in altered food webs. This could become apparent through a change in trophic position, smaller trophic niches, a reduction in trophic feeding levels or a reduction in trophic diversity, all most likely resulting from resource partitioning (Schoener 1974; Werner 1986; Baltz and Moyle 1993; Bohn and Amundsen 2001). Support for interspecific competition will be evident when two species have high dietary overlap in allopatry and low dietary overlap in sympatry (Schoener 1970; Werner 1986). The onset of competition with other planktivorous fishes, such as native planktivores and larval fishes (e.g., walleye), for planktonic resources could be traced using stable isotopes of nitrogen and carbon through time (Lu et al. 2002; Williamson and Garvey 2005; Chen et al. 2007; Irons et al. 2007; Kolar et al. 2007). We expect high dietary overlap and low competition among native planktivores and Asian carps during the initial invasion; however, as Asian carps become established and increase in density, we expect dietary overlap to decrease as native fishes shift to other food resources. For example, specialist species tend to be affected by invasive species more than generalist species as specialists experience more competition during disturbance (e.g., invasion; Clavel et al.2011). This results in an even smaller trophic niche for the specialist species (Clavel et al.2011). Lag times may also exist across time between invasion (acute phase) and realized impacts (chronic phase; Crooks 2005; Strayer et al. 2006). This lag phase may be shorter (days versus years) in short-lived species (e.g., minnows) as more generations are subjected to Asian carps presence. Also,

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depending on a species’ trophic position, effects or rates of change resulting from silver carp invasion may differ. For example, piscivores or top-level species are less at risk to disturbance than intermediate or lower level species (Mandrak 1989; Parent and Schriml 1995). Competition, either direct or indirect could be the driving mechanism in structuring Great Plains streams in eastern South Dakota as a result of the invading Asian carps. Differences among recipient ecosystems mediate effects of invasive species. The James, Vermillion and Big Sioux Rivers in South Dakota differ in basin size, water chemistry and underlying geology (Berry et al. 1993; Schmulbach and Braaten 1993; Galat et al. 2005). Asian carps invasion rates and invasion effects could vary between species and among the three rivers. Further, we expect within river differences: effects of Asian carps on food webs are likely to be strongest at confluence sites, since densities are highest here, in all rivers and decrease upstream coinciding with an abundance gradient of Asian carps. Any seasonal effects would manifest in the planktivore-plankton matchmismatch (Beaugrand et al. 2003; Wojtal-Frankiewicz 2012) plankton community composition where density varies predictably throughout the year (Berry et al. 1993; Wetzel 2001). Additionally, phytoplankton is less diverse and abundant in rivers than other habitats (e.g., lentic) as a result of greater turbulence and turbidity (Hynes 1970; Zale et al. 1989). Low abundance of resources for Asian carps may accentuate earlier impacts associated with Asian carp presence and abundance. Finally, in order to trace a middle out or a cascade effect we would examine all ecological levels of the food web, including primary producers (i.e., phytoplankton), primary consumers (i.e., zooplankton), secondary consumers (i.e., zooplanktivores) and

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top predators (i.e., walleye) for changes in the food web as related to trophic position and uptake of energy sources. A food web effect (e.g., decreased trophic niche, change in trophic position, decreased trophic diversity) at each level would suggest that Asian carps are keystone species (Bohn and Amundsen 2001). For example, if a trophic cascade existed, secondary consumers such as zooplankton might be negatively affected by silver carp ravenous foraging and feeding (Kolar et al. 2007) which would reduce condition of top level predators such as walleye or reduce abundance of walleye larvae and fry, and young-of-year that depend on zooplankton to survive (Jackson et al. 1992). Thus, walleye trophic diversity and trophic niche might shift and/or become smaller and the zooplankton community may also change in composition and become smaller. Trophic cascades created by Asian carps can affect the bottom of the food web by altering phytoplankton composition or the top of the food web by altering walleye prey (e.g., emerald shiner, zooplankton, invertebrates) and those levels in-between. The impact of Asian carps at this early stage of invasion could be insignificant as exploitative competition may be taking place where the impact of negative interactions takes time to develop (i.e., time lag) through gradual shared resource depletion with other fishes dependent on plankton during some stage in their life histories (Bohn and Amundsen 2001). As density of Asian carps increases, so will competition for resources which would result in resource limitation and a gradual reduction in niche overlap between Asian carps and other native fishes, particularly planktivores. Thus, we expected that ecological consequences would intensify through time. This study examined the assemblage structure and food web consequences of Asian carps (i.e., bighead carp, silver carp) invasion at the onset in North American Great

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Plains streams. We selected to sample three neighboring tributaries to the Missouri River (i.e., James, Vermillion and Big Sioux rivers) in South Dakota and multiple sites within each tributary. We quantified invasive species-driven ecological changes in assemblage structure and tested whether those changes corresponded to changes in fish trophic position and river food web structure. We expected Asian carps to have food web effects on the native pelagic planktivore functional feeding guild (i.e., bigmouth buffalo, emerald shiner, gizzard shad). Furthermore, we postulated that increases in Asian carps abundance would shift native pelagic planktivore trophic position away from Asian carps trophic position or shrink their trophic niche as competition may be an underlying mechanism of resource partitioning. Additionally, an increase in Asian carps abundance would lead to an increased reliance of native planktivores on benthic prey and thus lead to a reduction or change in food source from pelagic to benthic sources by native planktivores. These data can be used as a baseline for future studies on the potential negative ecological impacts (e.g., competition with resources) of Asian carps and other anthropogenic alterations (e.g., hydrologic alteration) within these and other aquatic ecosystems facing invasion.

Methods Study Area

The Missouri River is the longest river in North America (3,768 km). The stretch of river for this study includes 127 km below the lowermost impoundment, Lewis and Clark Lake (Gavins Point Dam). Three major tributaries in South Dakota enter the Missouri River

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within this reach: James, Vermillion and Big Sioux rivers (Fig. 1). The Big Sioux (watershed area = 23,325 km2), James (watershed area = 57,000 km2) and Vermillion (watershed area = 5,800 km2) rivers are warmwater streams characterized by low relief with abundant lakes, wetlands, and low-gradient streams of glacial origin (Hoagstrom et al. 2007). Upper sites tended to have higher turbidities, conductivity and TDS than lower sites within each of the three watersheds. Dissolved oxygen ranged from 3.17 (J01) to 9.24 (JC) and chl-a values were the lowest in the lower James sites (JC, J01, J03) ranging from 3.74 – 7.61 (Table 1). Wetland drainage, channelization and low-head dams (Sinning 1968; Owen et al. 1981) are common in these three rivers and land use such as range and agriculture dominate the watersheds (Galat et al. 2005; Hayer Chapter 2). These watersheds are also characterized by alternating wet years of prolonged flooding and high discharge, and dry years of extended drought and intermittent flows (Shearer and Berry 2003). In fact, mean annual discharge for the James River during the study period (2009 – 2012) consisted of the largest, second largest, and fourth largest discharge values on record dating back to 1949 (Table 2).

Field and laboratory methods Ten sites were selected from the James, Vermillion and Big Sioux rivers, South Dakota, for detailed assemblage and food web structure (Fig. 1). Food webs were analyzed using stable isotopes of carbon (δ13C/ δ12C) to assess energy sources and stable isotopes of nitrogen (δ15N/ δ15N) to assess trophic position (Layman et al. 2012) of Asian carps, native planktivores (i.e., bigmouth buffalo, emerald shiner, gizzard shad) and other species collected. Fish, crayfish, common aquatic and terrestrial vegetation and algae

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were collected. Fish were collected using a variety of gears (i.e., boat electrofishing, minifyke nets, hoop nets, minnow traps) to detect as much of the fish assemblage and all size classes possible (Drobish 2008; Guy et al. 2009). For fish >100 mm total length (TL), white muscle tissue was taken from just below the dorsal fin each species per site and total length was recorded. For juvenile and small-bodied fish (0.0072) with the VC site (1044 ppm) having the highest TDS and all three sites on the Big Sioux River having the lowest TDS (Table 2; mean = 621, 580, 544 ppm). Additionally, turbidity was significantly different among sites (F9,38=2.19; P>0.0400) with the J02 site having the highest mean turbidity (Table 2; mean = 88 NTU) and the V02 site having the lowest mean turbidity (Table 2; mean = 6.93 NTU). There was no longitudinal or basin pattern water quality variables. We collected 22,536 individual fish representing 53 species, 15 families and six trophic guilds (Appendix I). The piscivore trophic guild was occupied by 12 species, omnivore by 7 species, insectivore by 27 species, benthivore by 1 species, planktivore by 6 species and herbivore by 1 species. Insectivores comprised 54% of total catches, omnivores comprised 19%, piscivores comprised 10%, planktivores comprised 26%, herbivores comprised 0.02% and benthivores comprised 1% of total catches, respectively (Appendix I). Black bullhead, Ameiurus melas (30%), gizzard shad (13%) and emerald shiner (12%) comprised the highest percentage of catches throughout the study period and region (Appendix I). Skipjack herring, Alosa chrysochloris and plains topminnow, Fundulus diaphanus, were unique to the Vermillion River, bigmouth shiner, Notropis dorsalis, rock bass, Ambloplites rupestris, spottail shiner, N. hudsonius, trout-perch, Percopsis omiscomaycus and western blacknose dace, Rhinichthys obtusus, were unique

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to the Big Sioux River, and paddlefish, which were not targeted, was unique to the James River although they were observed on one occasion jumping at the Vermillion confluence site (VC). Skipjack herring were only collected in 2012 and rock bass, spottail shiner, plains topminnow, bigmouth shiner and fathead minnow, Pimephales notatus, were only collected in 2010. Silver carp comprised a total of 1.9% of catches throughout the study period; however there was an increase in percent catches each year from 0.00% in 2009, 0.70% in 2010, 8.90% in 2011 and 45.00% in 2012 (Hayer et al. 2014 Chapter 3). Of the planktivores, gizzard shad were sympatric with silver carp at six of ten sites, bigmouth buffalo at eight sites, and emerald shiner and bighead carp at five sites. Native planktivore distributions overlapped with invasive planktivores, being detected at almost all the same sites, except for bigmouth buffalo and emerald shiner which were found at an additional site on the Big Sioux River (i.e., BS03) where Asian carps were not detected. Overlap in distribution among planktivores allows for potential competition of food and habitat resources. According to Morisita’s Index, all confluence sites (JC, VC, BSC) had similar fish species compositions (MI>0.80; Table 3). The JC site had the highest mean similarity to other sites and J01 had the lowest mean similarity to other sites (Table 3). Species composition similarity was not related to longitudinal position within each river basin (Table 1). NMDS (stress value = 0.15) indicated that the VC and BSC were the most similar in species composition and all other sites were least similar in species composition (Fig. 2). Environmental and water quality variables (e.g., temperature, conductivity, watershed area) indicated that the distance to the confluence with the

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Missouri River is the strongest driver of the community structure at each site (Fig. 2) and was the only significant variable explaining community structure (r2=0.80; P>0.02).

Ecological changes in food web structure We collected 1241 fish dorsal muscle samples which included 388 silver carp, 13 bighead carp, 54 gizzard shad, 47 bigmouth buffalo and 10 emerald shiner samples from the James, Vermillion and Big Sioux rivers, 2009 – 2012 (Appendix I). The difference between the highest (25.25) and lowest (2.87) δ15N value was 22.38 ‰. Alternately, δ13C values ranged from -25.99 to -9.59, a difference of 16.40 ‰ (Fig. 3). The planktivore biplot revealed a 3.41‰ mean difference between silver carp and bigmouth buffalo (Fig. 4). Additionally, silver carp and bigmouth buffalo energy sources were the most depleted when compared to other planktivores (Fig. 4). There was trophic overlap between bighead carp, bigmouth buffalo, and emerald shiner and trophic overlap between silver carp and gizzard shad (Fig. 4). There was a 3.42‰ spread in mean energy source values of planktivores (Fig. 4). There was a significant correlation between δ13C and fish total length of the herbivore guild (r2=0.96, P=0.0015), and a significant correlation between δ15N and fish total length for the omnivore (r2=0.22, P=0.0204) and piscivore (r2=0.15, P=0.0304) guilds. James River fishes were generally less enriched in δ15N (mean=15.56, SE=0.25) than the Big Sioux (mean=16.01, SE=0.26) and Vermillion Rivers (mean=16.60, SE=0.29), which were similar. Additionally, fishes sampled in 2012 were the least enriched in δ15N (mean=15.00, SE=0.34) and different from those sampled in all other years. Additionally, the James and Vermillion rivers had similar δ13C values (mean=-

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27.31, SE=0.32; mean=-26.81, SE=0.38, respectively) and the Big Sioux River fishes (22.70, SE=0.33) were on average 4.36‰ more enriched in δ13C than those sampled from the James and Vermillion rivers. Post hoc analysis also revealed that δ13C values in 2011 were different from all other years and had the least enriched values (mean=-27.55, SE=0.53). Bigmouth buffalo trophic position overlapped with the omnivore guild and insectivores were almost completely within the range of piscivore trophic positions (Appendix I). Planktivore species were the most variable with a half a trophic level difference within the guild (Fig. 5). Silver carp, gizzard shad and grass carp occupied the lowest trophic positions and flathead catfish, longnose gar, Lepisosteus osseus, and walleye, Sander vitreus, occupied the highest trophic positions (Appendix I). Within planktivores, silver carp and gizzard shad occupied the lowest trophic positions and native planktivores, emerald shiner and bigmouth buffalo occupied the highest trophic positions (Fig 4). The mixed model testing for differences in trophic position (TP; dependent variable) among guilds, rivers, and years was significant among guild, rivers, and years (F5, 1230=53.23, P