MIAMI UNIVERSITY the Graduate School Certificate for Approving the Dissertation

MIAMI UNIVERSITY The Graduate School

Certificate for Approving the Dissertation

We hereby approve the Dissertation of Andrew J. Tucker

Candidate for the Degree: Doctor of Philosophy

______Advisor (Dr. Craig E. Williamson)

______Committee Member (Dr. Thomas O. Crist)

______Committee Member (Dr. Mark H. Olson)

______Committee Member & Reader (Dr. James T. Oris)

______Committee Member/Graduate School Representative (Dr. A. John Bailer) ABSTRACT

THE ROLE OF ULTRAVIOLET RADIATION IN MEDIATING WARMWATER FISH INVASION IN TRANSPARENT LAKES

by Andrew J. Tucker

Considerable uncertainty remains regarding what the impacts of sustained or changing ultraviolet radiation (UV) stress may be for aquatic ecosystems. This dissertation represents the first attempt to explore the role of UV in a biological invasion context. Chapter 1 serves as a broad introduction to the role of UV in lakes. It emphasizes the complex interactive and indirect effects of UV on aquatic ecosystems and in particular the implications of UV for disease dynamics, contaminant toxicity, carbon cycling, and biodiversity in lakes. The remaining chapters elaborate on the topic of UV and biodiversity by developing the idea that UV can regulate warmwater fish invasion in transparent lakes. Chapter 2 provides the foundation for this idea. The results from an in situ incubation experiment that manipulated incident UVR exposure of larval bluegill (Lepomis macrochirus), and an assessment of UVR exposure levels in nearshore habitats using DNA dosimeters are reported. The results demonstrate that UVR mediates habitat invasibility in a large sub-alpine lake (Lake Tahoe CA/NV USA). In chapter 3 an invasion-window model is developed which predicts that largemouth bass (Micropterus salmoides) establishment in Lake Tahoe is controlled by the ability of these non-native fish to cope with UV and temperature stress along a UV-temperature stress gradient. In situ incubation experiments in contrasting high and low-UV nearshore habitat corroborated model predictions. In situ incubation experiments with native Lahtonan redside minnows (Richardsonius egregius) showed that the native species overlaps broadly with the range of ambient environmental conditions in the Tahoe littoral zone. The final chapter suggests a novel approach for managing warmwater fish invasion in lakes. A UV Attainment Threshold, which is a target value for water transparency based on 1) incident solar UV exposure levels during peak spawning season, and 2) experimentally derived UV exposure levels lethal to larval warmwater fish, was developed. Taken together these chapters provide novel insights that could improve our ability to predict and manage aquatic species invasion, especially in transparent lakes. THE ROLE OF ULTRAVIOLET RADIATION IN MEDIATING WARMWATER FISH INVASION IN TRANSPARENT LAKES

A DISSERTATION

Submitted to the Faculty of

Miami University in partial

fulfillment of the requirements

for the degree of

Doctor of Philosophy

Department of Zoology

Andrew J. Tucker

Miami University

Oxford, OH

2011

Dr. Craig E. Williamson, Miami University, Chair, Major Advisor

Table of Contents

LIST OF TABLES iii

LIST OF FIGURES iv

ACKNOWLEDGEMENTS vi

GENERAL INTRODUCTION 1

CHAPTER 1: LAKES IN A NEW LIGHT: DIRECT AND INDIRECT EFFECTS OF ULTRAVIOLET RADIATION

CHAPTER 2: ULTRAVIOLET RADIATION AFFECTS INVASIBILITY OF LAKE ECOSYSTEMS BY WARMWATER FISH 46

CHAPTER 3: THE INVASION-WINDOW FOR WARMWATER FISH IN A LARGE SUB- ALPINE LAKE: THE ROLE OF UV RADIATION AND TEMPERATURE 72

CHAPTER 4: A UV ATTAINMENT THRESHOLD FOR THE PREVENTION OF WARMWATER AQUATIC INVASIVE SPECIES

104

CONCLUDING REMARKS 133

List of Tables

TABLE 2.1. ATTENUATION DEPTHS FOR 320-NM UV AT 9 SITES IN LAKE

TAHOE...... 65

TABLE 3.1. SURVIVAL OF LARGEMOUTH BASS AND LAHONTAN REDSIDE LARVAE

AFTER IN SITU INCUBATION……………………………………………………….94

TABLE 4.1. UV TRANSPARENCY OF NEARSHORE SITES IN LAKE

TAHOE……………………………………………………………………………...….126

TABLE 4.2. ATTAINMENT/NON-ATTAINMENT STATUS FOR 11 NEARSHORE SITES

BASED ON UVAT……………………………………………………………………..127

TABLE 4.3. TEMPEARTURE (°C) AT 1 METER FOR 11 NEARSHORE

SITES……………………………………………………………………….…………..128

iii

List of Figures

FIGURE 1.1. DIRECT AND INDIRECT EFFECTS OF UV ON LAKE BIOTA……..………41

FIGURE 1.2. VERTICAL PROFILES OF UV FROM LAKES AROUND THE WORLD…..42

FIGURE 1.3. THE ROLE OF UV IN REGULATING PATHOGENS IN LAKES…………...43

FIGURE 1.4. THE ROLE OF UV ON DIVERSITY IN LAKES………………………………44

FIGURE 1.5. UV EFFECTS ON CARBON CYCLING IN LAKES………….……………….45

FIGURE 2.1. PHOTO OF NEARSHORE SITES IN LAKE TAHOE…………………………66

FIGURE 2.2. MAP OF LAKE TAHOE INDICATING LOCATION OF SAMPLE SITES…..67

FIGURE 2.3. RELATIONSHIP BETWEEN DNA DAMAGE AND UV EXPOSURE….……68

FIGURE 2.4. UV EXPOSURE, DNA DAMAGE, AND SURVIVAL OF BLUEGILL LARVAE IN EXPERIMENTAL MICROCOSMS…………………………………………………………69

FIGURE 2.5. DNA DAMAGE AND UV TRANSPARENCY RELATIONSHIP……………..71

FIGURE 3.1. CONCEPTUAL MODEL OF INVASION-WINDOW………………………….95

FIGURE 3.2. DIAGRAM OF THE METHOD USED TO PREDICTION INVASION POTENTIAL……………………………………………………………………………………..96

FIGURE 3.3. MAP OF LAKE TAHOE INDICATING LOCATION OF STUDY SITES…….97

FIGURE 3.4. EFFECT OF UV EXPOSURE ON REDSIDE AND BASS SURVIVAL………98

FIGURE 3.5. EFFECT OF TEMPERATURE ON HATCHING SUCCESS OF BASS EGGS..99

FIGURE 3.6. PREDICTED REPRODUCTIVE SUCCESS FOR BASS AT 3 SITES……….100

FIGURE 3.7. CONTOUR PLOT OF PREDICTED BASS REPRODUCTIVE SUCCESS…..102

FIGURE 3.8. PLOT OF MEAN UV AND TEMPERATURE CONDITIONS IN 11 NEARSHORE SITES…………………………………………………………………………..103

FIGURE 4.1. PHOTO OF YOLK-SAC LARVAE……………………………………………129

FIGURE 4.2. MAP OF LAKE TAHOE INDICATING LOCATION OF SAMPLE SITES…130

FIGURE 4.3. FREQUENCY PLOT OF 305 NM UV SURFACE EXPOSURE……………..131

FIGURE 4.4. EXPOSURE-RESPONSE CURVES FOR BASS, BLUEGILL, AND REDSIDE LARVAE……………………………………………………………………………………….132

Acknowledgements

I owe many thanks to my advisor! Thank you Craig for so generously sharing your brilliant ideas and for showing me what it looks like to love what you do. When you invited me to accompany you to Lake Tahoe back in 2006 I figured it would be no more than a one-time sampling trip. Somehow I think you knew it would be much more.

I’m also grateful to my committee members. Jim, Tom, John, and Mark thank you for your advice along the way. Your insights helped me to place this research in a relevant context and each of you contributed an expertise that improved this dissertation. A special thanks to Jim for his help in making connections with key people at Lake Tahoe.

Speaking of key people at Lake Tahoe… I’d like to thank our collaborators at the Tahoe Environmental Research Center and the University of Nevada-Reno. Anne Liston, Geoff Schladow, Brant Allen, Jill Falman, Marcy Kamerath, Sudeep Chandra, and Christine Ngai were especially helpful.

I’m also incredibly grateful to those that accompanied me on field trips and helped in the lab over the years. No one put in more hours with me than Michael Cohen. Thanks Mike for weekend fish feedings and for keeping me sane out on the lake. I apologize if I drove you insane! Erin Overholt, Kevin Rose, Jeremy Mack, Carrie Kissman, Ian Lizzadro-McPherson, Amanda Gevertz, and Sandi Connelly were all key contributors to this research as well. Thank you all for your help and for your friendship.

To my friends outside of the halls of Miami and to my family, thank you for your encouragement and support throughout this journey. Mom, Dad, and Melissa thank you for never doubting I could do this. Duane and Rachel thank you for always showing an interest in my work and for helping us to celebrate each milestone along the way. Rebekah I’d have given up a long time ago without you. And Owen, thank you for always making everything seem worth it!

General Introduction

Ultraviolet radiation in a biological invasion context

Ultraviolet radiation is a potent and ubiquitous environmental stressor but only in the last 20 to 30 years has the field of the ecology of ultraviolet radiation (UV) emerged to explore the important role that UV can play in regulating ecosystem structure and function. Early research in the realm of UV ecology tended to emphasize the direct negative effects of UV exposure for individuals or populations with particular emphasis on the implications of increasing UV exposure for primary producers in the world’s oceans (e.g. El-Sayed 1988). Over time the field of the ecology of UV has matured and we are beginning to appreciate the often complex indirect effects of UV on ecosystems and on lakes in particular. For example, advances in our understanding of the ecology of UV in lakes may explain some enigmatic patterns including, why zooplankton exhibit vertical migration in lakes where there are no visual predators to induce their migration and why declining water transparency seems to result in decreased species diversity among African cichlids (Williamson and Rose 2009). Furthermore, UV ecologists have identified potential benefits of UV exposure in lakes for both individuals and ecosystems (Williamson et al. 2001, Williamson and Rose 2010). One largely unexplored potential benefit of UV exposure in lakes is the possibility that UV could help maintain species diversity. In fact, UV is generally thought to reduce species diversity by limiting the number of UV intolerant taxa that occur in high UV lakes (e.g. Marinone et al. 2006). However taken in a different context the exclusion of certain UV sensitive taxa from clear waters could actually help to maintain the integrity of communities of native species and thus promote higher levels of diversity in lakes. For example, in Lake Tahoe CA/NV, USA introduced warmwater fish (bluegill and largemouth bass in particular) have virtually extinguished native minnow species (Lahontan redsides) in some nearshore habitats. The introduction of these warmwater fish coincided with the documented decline in water transparency in the lake. Furthermore, established populations of these warmwater non-natives appear to be limited to nearshore sites in the south end of the basin characterized by extensive development and in close proximity to some of the lake’s largest tributaries where water transparency tends to be relatively low (Kamerath et al. 2008). If the survival of these non-native species is related to UV exposure then changes in water transparency (and thus underwater UV levels) could have important implications for the fish community in Lake Tahoe.

This question of whether or not UV exposure is important in regulating the survival and establishment of non-native and potentially invasive species is one that to my knowledge has not been previously explored. In fact, the role of abiotic factors in general as regulators of biological invasion is often only explored at very broad spatial scales (e.g. using ‘climate envelope models’ to predict the potential distribution of a species in a novel habitat based on correlation of that species current distribution with abiotic variables across its native range). However, evidence is emerging that suggests a central role for abiotic factors in regulating invasions at a more local level (e.g. Holway et al. 2002, Dethier and Hacker 2005, Menke and Holway 2006, Gerhardt and Collinge 2007). A primary goal of my dissertation was to contribute to this body of literature by testing the role of UV in regulating warmwater fish invasion in transparent lakes using Lake Tahoe as a case study. The concept that UV regulates invasion extends disturbance/stress models, originally developed in plant ecology, to explain the limited invasiveness of warmwater fish in clear, high- UV environments. In Lake Tahoe, the warmer waters required for non-native fish spawning only occur in the shallow nearshore environment where exposure to UV is potentially high. Thus, UV and temperature interact to define a high stress environment accommodating a limited, vigorous native community but probably excluding many potential invaders in the absence of disturbance. Historical conditions in Lake Tahoe likely excluded warmwater fish from spawning due to the high levels of damaging UV radiation in the warmer, shallow, nearshore habitats. However, human and natural disturbance have expanded the physical habitat conditions to include warmer temperatures and lower UV transparency, and thus an “invasion-window” has opened where warmwater fish can successfully survive (a “survival window”) and reproduce (an “establishment window”). Changes in transparency are not uncommon in lakes and in fact, represent one of the more typical types of disturbance events in lakes (Mazumder et al. 1990, Edmondson 1991, Kaufman 1992, Seehausen et al. 1997). Thus, whether or not UV plays an important role in regulating biological invasion could have important implications for species invasion in lakes.

Objectives & Hypotheses

The overall goals of this dissertation were 1) to test the role of UV as it relates to warmwater fish invasion in transparent lakes using Lake Tahoe as a case study and 2) to apply 3

the findings in a management context to help lake managers answer the questions, “How do we predict where invasive species will occur and how do we prevent their establishment?” These overall goals were addressed by accomplishing the following objectives:

Objective 1: Test the concept that UV influences whether or not warmwater fish can establish in transparent lakes. Hypothesis 1: UV controls the suitability of nearshore habitats for the earliest life history stages of warmwater fish. Approach to test H1: Larval bluegill and largemouth bass were collected from nests in Lake Tahoe and used for in situ incubation experiments in a number of nearshore locations across a gradient of UV transparency. These in situ experiments were 4 day incubations at 0.5 to 1m depth, consistent with standard incubation depth and duration for warmwater fish. The incubations utilized UV blocking and UV transparent microcosms to isolate the effect of UV on larval survival. In the first round of experiments, dosimeters of raw DNA in solution were used as tools to assess potential UV effects on larval bluegill by relating DNA damage in dosimeters with larval fish mortality. DNA dosimeters were incubated with larval bluegill in contrasting high and low UV sites and DNA dosimeters alone were placed in 8 additional sites to estimate the potential for bluegill survival in those sites.

Objective 2: Develop an invasion-window model that predicts invasion potential (i.e. larval survival) for largemouth bass based on UV and temperature conditions in nearshore habitats. Hypothesis 2: UV will induce mortality in pre-feeding larval bass at natural UV exposure levels that are not lethal to native Lahontan redside minnows. Prediction 2.1: Reproductive success for non-native bass will vary as a function of UV transparency and temperature in nearshore locations. Approach to test H2: The UV tolerance of larval largemouth bass and Lahontan redside minnows collected from Lake Tahoe was tested under ambient UV conditions in a temperature controlled outdoor UV exposure chamber. Briefly, larvae were added, across treatments, to each of 12 treatment dishes and 3 control dishes. Neutral density stainless steel mesh screens were placed over treatment dishes to provide a range of UV exposure levels. Treatments were then

placed in the outdoor temperature controlled water bath and larval survival was monitored every 2 hours for up to 6 days. Temperature effects on bass survival were estimated from a published study on the effects of incubation temperature on hatching success of largemouth bass eggs. To test the prediction that bass reproductive success would vary with UV and temperature conditions, a simple model was developed based on the UV and temperature responses of bass larvae. The potential for larval bass survival in 11 sample sites over the range of UV exposure and temperature conditions observed from 4 years of sampling data was assessed. In situ incubation experiments in a subset of the 11 sites were conducted to test model predictions.

Objective 3: Develop a management tool that could be used to predict where warmwater fish establishment could occur. Approach: In addition to the UV tolerance tests for largemouth bass and Lahontan redside larvae described above as part of the approach to test H2, a UV tolerance experiment of the same design was conducted with larval bluegill collected from Lake Tahoe. Results from these UV tolerance tests were used to develop a UV Attainment Threshold (UVAT). The UVAT is a target value for water transparency based on, 1) incident solar UV exposure levels during peak spawning season, and 2) experimentally derived UV exposure levels lethal to larval warmwater fish. The UVAT is reported as the percent of surface 305-nm UV exposure that must penetrate to any given depth to prevent larval warmwater fish survival (i.e., result in mortality of 99% of the population):

UVAT = 100*(LE99 / E0) -2 where: LE99 = effective UV exposure selected to target bass mortality (kJ*m ) -2 E0 = cumulative 4 day 305-nm UV exposure at surface (kJ*m )

Organization of Dissertation

In addition to this general introduction this dissertation has been divided into 5 sections. Chapters 1 through 4 are written and conceived of as manuscripts for publication (chapter 2 has already been published). Chapter 1 serves as a broad introduction to the role of UV in lakes. Chapter 2 addresses objective 1 by testing the concept that UV controls the suitability of nearshore habitats for warmwater fish establishment. In chapter 3 the invasion-window model

relevant to objective 2 is developed. The final chapter introduces the UVAT, relevant to objective 3. The final section of concluding remarks includes a synthesis of how the results address the overall goals of the research and offers recommendations for future research. Tables, figures, and references appear at the end of each chapter.

Chapter 1

Lakes in a new light: Direct and indirect effects of ultraviolet radiation

To be submitted to: Global Change Biology

Abstract

Stratospheric ozone concentrations are not expected to recover to pre-ozone hole levels until the mid-21st century, and even after ozone recovery climate and other anthropogenic changes could alter ultraviolet radiation (UV) exposure regimes in aquatic ecosystems. Though our understanding of the ecology of UV continues to move towards a new paradigm that emphasizes complex interactive and indirect effects of UV on communities of organisms, rather than simple direct negative effects on individuals, considerable uncertainty remains regarding what the impacts of sustained or changing UV stress may be for aquatic ecosystems. In this synthesis we examine the importance of indirect UV effects for some key ecosystem level characteristics and processes in lakes. In particular we draw attention to the implications of UV for disease dynamics, contaminant toxicity, biodiversity, and carbon cycling in lakes. Though UV has strong lethal mutagenic and chronic physiological effects on organisms, we suggest that indirect effects pathways, including UV effects on animal behavior, food quality, and trophic interactions are likely to be especially important in determining ecosystem dynamics in lakes.

Introduction

Ultraviolet radiation (UV), high energy solar radiation at wavelengths between 280 and 400 nm in Earth’s atmosphere, is a potent and ubiquitous environmental stressor. While UV has been recognized as an environmental driver for some time, the discovery of the ‘ozone hole’ in the mid-1980s sparked more intense interest in the important role of UV in both natural and human-dominated ecosystems. During the period from 1990 to 2000 the frequency of publications concerning UV and freshwater ecosystems in the international scientific literature increased almost 300% (Bronmark and Hansson 2002). Many of these publications emphasized the detrimental role of UV for aquatic ecosystems and some foretold ecological disaster, prompting swift action by the international community to limit the use of ozone depleting substances. International agreements to curb ozone depletion led to the Montreal Protocol which has been effective in reducing levels of anthropogenic ozone depleting substances in the stratosphere (McKenzie et al. 2011). However ozone depletion continues and stratospheric ozone concentrations at mid-latitudes are not expected to recover to pre-1980 levels until the mid 21st century and perhaps even later in polar regions (Andrady et al. 2010, McKenzie et al. 2011). In the last decade we have developed a more nuanced understanding of the role that UV plays in aquatic systems and in lakes especially. This new perspective recognizes not only the detrimental effects of UV exposure but also the potential benefits of UV for individuals and whole lake ecosystems (Williamson et al. 2001a,Williamson and Rose 2010). Advances in our understanding of the ecology of UV have also helped to explain some enigmatic patterns in lakes related to animal behavior, phototoxicity, and species diversity (Williamson and Rose 2009). Our goals in this synthesis are to highlight these more subtle (often indirect) effects of UV on lakes and to draw attention to the implications of UV for disease dynamics, contaminant toxicity, biodiversity, and carbon cycling in lakes. We start by showing how indirect effects pathways can be as important as direct effects of UV on aquatic organisms in lakes (Fig. 1.1). We define indirect effects as having pathways where UV interacts with some component of the environment (abiotic or biotic) to induce an effect on an individual or population that may or may not otherwise be directly exposed to or affected by UV. For example, indirect UV effects could be realized through UV effects on

predators, competitors, animal behavior, or on organic or inorganic chemical compounds in the water. In subsequent sections we describe UV effects on disease dynamics and contaminant toxicity in lakes as two specific examples that demonstrate the strong role that indirect UV effects can play in regulating populations and communities of lake organisms. Finally, we discuss the role of UV in regulating species diversity and carbon cycling in lakes. We suggest that the strong direct and indirect effects of UV on these processes in individual lakes could have important implications for biodiversity and carbon balance on broader geographic scales. All of these UV mediated effects are dependent upon levels of surface and underwater UV exposure in lakes. Therefore, we begin with a brief discussion of the factors that regulate UV exposure in lakes.

UV attenuation in lakes

UV transparency varies substantially among lakes (Fig. 1.2). A useful tool for comparing

UV transparency is the 1% attenuation depth (Z1%, the depth to which 1% of surface irradiance of a given wavelength penetrates) for 320 nm UV. The 1% depth has no particular ecological significance but rather is used as a convenient metric that is consistent with the 1% depth of photosynthetically active radiation (PAR), which in turn represents the compensation depth where photosynthesis equals respiration. The 320 nm wavelength is used both because it is available on commonly used submersible radiometers and also is within the wavelength range wherein there is a maximum biological effect when one considers both the energy of the photons and the number of photons at a given wavelength in incident solar irradiance (Williamson et al.

2009). In some lakes this Z1% 320nm is less than a fraction of 1 m, whereas in other lakes it can be in excess of 20 m (Fig. 1.2). Lake location, watershed features including soil type and catchment slope, and atmospheric characteristics like precipitation and ozone layer all influence UV exposure levels in lakes. Lakes at high latitudes tend to have lower incident UV exposure than equatorial lakes (Rautio & Tartarotti 2010), although polar regions have experienced the most substantial declines in ozone concentration and thus the greatest increase in UV exposure relative to baseline conditions (ACIA 2005). UV exposure at the lake surface also tends to increase with increasing elevation.

For example, in the Central Alps UV flux increases by 11% per 1000 meters for 320 nm UV-A (Blumthaler et al. 1992). Below the surface of a lake dissolved and particulate substances determine the depth to which UV penetrates. Colored or chromophoric dissolved organic matter (CDOM), often measured as dissolved organic carbon concentration (DOC), is generally the most important determinant of UV transparency. Called the ‘ozone of the underwater world’ (Williamson and Rose 2010), DOC concentrations are a good predictor of underwater UV levels (Morris et al. 1995, Rae et al. 2001). Two important exceptions to this are very low DOC lakes in alpine regions (Laurion et al. 2000, Sommaruga and Augustin 2006), and relatively low DOC eutrophic lakes (Hodoki and Watanabe 1998) where phytoplankton play a more important role in regulating underwater UV exposure than DOC. While inorganic particulates are likely to be important in regulating UV transparency in highly turbid reservoirs or glacial systems, few data are available on the ecology of UV in these kinds of systems. The quality and quantity of UV regulating materials can vary seasonally as the physical and biological conditions in lakes change (Morris and Hargreaves 1997, Hayakawa and Sugiyama 2008). When combined with the variation in incident UV related to changes in sun angle and ozone these changes can result in strong seasonal variation in UV transparency as well. While stratospheric ozone concentrations begin to stabilize, moving atmospheric UV levels toward pre-ozone hole levels, the factors controlling UV attenuation in lakes remain highly variable. Many lakes in Europe and North America have shown increasing trends in DOC concentrations (Roulet and Moore 2006). The magnitude of the changes in DOC will vary with temperature and depend largely on precipitation (Pace and Cole 2002), which varies regionally and has been shown to affect DOC concentrations in lakes at a continental scale (Zhang et al. 2010). Anthropogenic acidification can also have strong effects on DOC quality and quantity and thus UV attenuation in lakes (Schindler et al. 1996, Williamson et al. 1996). In one lake, drought-induced acidification increased the estimated depth of UV-B penetration more than two- fold (Yan et al. 1996). There is some evidence that recovery of lakes from acidification can lead to a rebound in DOC concentration (Evans et al. 2006, Monteith et al. 2007). Elsewhere, decreasing DOC concentrations are still reported (Striegl et al. 2005, Monteith et al. 2007), with likely corresponding increases in UV transparency and underwater exposure levels. Eutrophication-driven changes in transparency to visible light have also been widely observed

(Rast and Lee 1978; Goldman 1988; Edmondson 1991). UV transparency is often a more sensitive sentinel of environmental change than visible transparency in lakes (Rose et al. 2009). This suggests that anthropogenic eutrophication, especially in low DOC lakes where phytoplankton play an important role in controlling UV attenuation, could also significantly reduce UV penetration in lakes.

Indirect effects of UV on lake organisms

Much of the initial information on the role of UV in lakes emphasized the direct negative impacts of UV at the organism level. This paradigm made sense at the time given the unknown effects of a large increase in levels of high energy radiation in the atmosphere in the early days of ozone depletion and the evidence for strong chronic (physiological stress, depression of physiological processes) and lethal (mutagenesis) effects of UV in a variety of organisms. These strong negative direct effects led to the suggestion in some early reports that elevated UV could contribute to wholesale collapse of aquatic food webs by direct effects on primary producers (El- Sayed 1988). This fear was largely rebuffed (e.g. Vincent and Roy 1993), but for some time now aquatic ecologists have recognized the potential for substantial, though more subtle and complex, impacts on lake ecosystems through indirect UV effects (e.g. Williamson 1995). The attention given to indirect effects of UV in lakes has generated important and surprising insights, suggesting even beneficial effects of UV on aquatic organisms. In this section we highlight some of the most important indirect effects of UV on phytoplankton, zooplankton, and fish in lakes. Whereas UV can directly suppress primary production through photoinhibition by limiting the photosynthetic capacity of phytoplankton (Olesen & Maberly 2001, Leavitt et al. 2003, Lorenzen 1979), indirect trophic-mediated UV effects can, in some cases, increase primary production. For example, in a stream where benthic consumers were more sensitive to UV exposure than sympatric algae, primary production under enhanced UV conditions actually increased (Bothwell et al. 1994). Assessments of UV effects based on changes in UV alone can also be misleading. It is important to consider the interaction of ‘whole-ecosystem’ processes like mixing, temperature, and nutrients when predicting the effect of UV on algal production. For example, the effect of UV on algal production can be mediated by nutrient limitation. In general, algal sensitivity to UV increases with lake nutrient content but declines again as light

becomes limiting at high algal biomass (Moeller 1994, Xenopolous et al. 2002). Furthermore, in light limiting conditions increases in surface irradiance may stimulate phytoplankton growth in spite of increases in UV exposure because light limitation can be more important than UV damage (Xenopolous et al. 2009). UV can also interact with temperature to affect phytoplankton communities. For one, UV and temperature interact to alter zooplankton grazing rates which can in turn change algal abundance (Williamson et al. 2010). These UV and temperature mediated grazing effects are species specific, suggesting that changes in UV and temperature may favor some algal species over others (Williamson et al. 2010). Moreover, thermal stratification can intensify the impact of UV on phytoplankton biomass and community composition (Xenopolous et al. 2000).

For zooplankton, the majority of studies on direct UV-induced damage report increased mortality (Zagarese et al. 2003, Rautio & Tartarotti 2010). Some direct sublethal effects, including reduced growth rates and decreased fecundity (De Lange et al. 1999, Huebner et al. 2006) as well as increased respiration rates (Fischer et al. 2006) have also been reported. In some cases indirect effects may also negatively impact zooplankton demography. For example, UV affects the morphology, biochemistry, and stoichiometry of phytoplankton (Hessen et al. 1997). Thus, UV irradiated algae often decrease zooplankton growth and fecundity through UV induced changes in phytoplankton fatty acid composition (DeLang and VanDonk 1997, De Lange and Van Reeuwijk 2003). On the other hand UV induced changes in fatty acid composition and decreased C:P and N:P ratios in Selenastrum spp had no significant effects on D. magna growth rate (Leu et al. 2006), suggesting that species specific differences in phytoplankton response to UV (and therefore impacts on grazers) are important.

The extent to which zooplankton grazers are influenced by UV may depend heavily on ecological tradeoffs. For example, it has been demonstrated that in thermally stratified clear lakes the negative phototaxis of zooplankton to UV light can constrain them to suboptimal temperatures that may compromise fitness (Cooke et al. 2008). Other studies have shown that high UV intensity may increase predator-prey overlap and thus increase predation pressure on zooplankton (Alonso et al. 2004; Boeing et al. 2004). The role of photoprotective compounds (PPCs) in zooplankton survival is an especially informative example of the kinds of UV induced

tradeoffs that exist for zooplankton. Zooplankton accumulate PPCs, like carotenoids or mycosporine amino-acids, as protection from UV induced DNA damage (Tartarotti et al. 2001, Persaud et al. 2007). Zooplankton do not appear to synthesize PPCs, and instead acquire them directly from their food (Carefoot et al. 2000, Moeller et al. 2005). An important exception is the production of melanin in some zooplankton, especially cladocerans (Hansson et al. 2007, Rautio et al. 2009), though this is quite costly energetically and is an important tradeoff in itself (Hessen 1996). UV exposure regulates PPC levels in phytoplankton (Rieger and Robinson 1997, Tartarotti et al. 2004). Thus UV effects on algal food quality, via changes in the concentration of PPCs, can increase UV tolerance in zooplankton. However, the incorporation of algae rich in photoprotective MAAs or carotenoids makes zooplankton more conspicuous to visual predators (Hairston 1976, Hansson 2004).

Some recent studies suggest that zooplankton defense mechanisms may be optimized in a way that minimizes the potential cost of UV induced tradeoffs. For example, the transparency gradient hypothesis (TGH), based on a survey of zooplankton distribution in alpine and sub- alpine lakes and experimental tests of PPC concentration and UV tolerance, suggests that zooplankton respond to UV and predators in a way that reduces the overall threat of both stressors (Kessler et al. 2008). The TGH asserts that UV will drive vertical migration in high UV conditions whereas fish predation is the most important driver of downward migration in low UV lakes. Another recent study, comparing population dynamics and reproduction of several zooplankton taxa in mesocosms, observed little difference in long-term zooplankton population dynamics under substantially different UV exposures (Hansson and Hylander 2010). This study suggests that effective defense mechanisms, like those underlying the TGH, might allow zooplankton population dynamics and community composition to persist relatively unchanged under changing UV conditions. On the other hand, this experiment (because it was conducted in small predator-free mesocosms) did not consider the full suite of UV mediated indirect effects, including UV induced changes on the feeding efficiency of zooplankton predators or the implications of UV-induced migration into sub-optimal habitats. Although there is evidence that zooplankton can optimize their UV response to minimize both direct (e.g. by incorporating/synthesizing PPCs) and indirect (e.g. optimized UV avoidance strategies) UV effects (Hansson and Hylander 2009), additional mechanistic studies that examine multiple UV

induced tradeoffs simultaneously are needed to verify whether the net response to UV exposure in zooplankton is generally positive or negative.

A number of strong direct negative UV effects have been demonstrated in fish. These include egg and larval mortality and ‘sunburn,’ which can initiate outbreaks of infectious disease (Zagarese and Williamson 2001). However, as with zooplankton and phytoplankton there are also some important indirect UV effects to consider. In particular, many fish exhibit strong behavioral responses to UV. For example, larval whitefish avoid high UV conditions in the surface waters of transparent lakes on sunny days (Ylönen et al. 2005). Spawning depths for yellow perch and bluegill are substantially greater in high UV lakes than in low UV lakes (Williamson et al. 1997, Gutierrez-Rodriguez and Williamson 1999). Also, while ambient UV levels can induce significant mortality for larval bluegill across the range of depths at which bluegill spawn, bluegill appear to select nest locations that limit UV exposure, and can even modify mean nest depth over the course of a season to limit UV exposure (Gutierrez-Rodriguez and Williamson 1999, Olson et al. 2008). For juvenile salmonids that can detect UV there is also a tendency to increase shade seeking behaviors in the presence of UV (Kelly and Bothwell 2002, Holtby and Bothwell 2008).

While these avoidance responses may alleviate the direct negative effects of UV on fish there are some important implications of UV avoidance. For example, the tendency for juvenile salmon to seek shade under high UV conditions may increase visual isolation from neighbors. This may in turn allow more fish to occupy the same habitat, resulting in density-dependent decreases in average size and ultimately reduced overwinter survival and smolt production (Holtby and Bothwell 2008). Likewise, the tendency to select deeper nest depths under high UV conditions may well impose thermal constraints that limit reproductive success (Huff et al. 2004). UV mediated habitat use may also generate changes in predator-prey overlap (Alonso et al. 2004) that could alter food availability, and thus growth potential and reproductive capacity, for planktivorous fish.

UV can influence disease dynamics in lakes

Pathogens, defined here simply as any infectious agent that causes disease to its host, have numerous impacts on lake structure and function. For example, pathogens can regulate community composition by eliminating susceptible species from lakes (Scudder 1983). Some bacterial parasites reduce population densities of important plankton grazers in temperate lakes (Duffy and Hall 2008). Parasites have even been known to induce changes in food chain length and increase omnivory in lake food webs (Amundsen et al. 2009). Perhaps one of the most noted pathogens, the chytrid fungus (Batrachochytrium dendrobatidis) has been implicated in the severe decline or extinction of nearly 200 species of frogs, threatening global biodiversity (Skerratt et al. 2007). Given the potential for such strong impacts it is worth asking how UV might interact with pathogens in lakes, and whether UV will generally alleviate or exacerbate pathogen effects. Indirect UV mediated trophic effects may be especially influential. Outcomes will depend on 1) the relative susceptibility of host and pathogen to UV, 2) the interplay of host immune function and pathogen susceptibility to UV, 3) the effect of UV on the kinds of available hosts (whether susceptible hosts or ‘diluters’), and 4) the kind of predator community (whether ‘sloppy’ predators or visually selective predators) (Fig. 1.3).

UV exposure can directly reduce pathogen infectivity. Recent UV exposure experiments with Cryptosporidium parvum showed that ambient levels of solar UV can rapidly inactivate oocysts and reduce infectivity (Connelly et al. 2007, King et al. 2008) although the effectiveness of UV disinfection in water varies with exposure time and water turbidity (Gomez-Couso et al. 2009). UV-B is also known to weaken disease defense systems (Orth et al. 1990, Tevini 1993). Studies that assess the interaction of UV induced changes in pathogen infectivity and host susceptibility to pathogens under natural UV levels are lacking. Also, some hosts are inherently more resistant to parasites than others. When parasite resistant hosts are present in large numbers, these so called ‘diluters’ can inhibit disease (Hall et al. 2009). We do not currently know if ‘diluters’ and susceptible hosts exhibit inherent differences in UV tolerance, but UV mediated shifts in host community composition are common (as discussed below for zooplankton) and changes in the relative abundance of susceptible versus resistant hosts could potentially influence disease dynamics in lakes.

UV mediated trophic interactions have perhaps the greatest potential to either exacerbate or ameliorate epidemics. For example, many fish possess UV photoreceptors (Leech and Johnsen 2003), which increase predation rates and foraging efficiency and enhance prey detection in fish from a variety of families, including Centrarchids, Percids, and Salmonids (Leech et al. 2009, Loew et al. 1993, Browman et al. 1994). In lakes selective predation and increased predation intensity both stem epidemics, by reducing host body size (competence) in the latter case (Hall et al. 2010, Hall et al. 2007) and by culling infected hosts in the former (Hall et al. 2005, Duffy et al. 2005). Consequently, UV-enhanced fish predation on zooplankton hosts would appear to be a viable mechanism whereby UV could inhibit disease. This hypothesis has not been explicitly tested. Nevertheless, results consistent with the idea that UV mediated trophic interactions stem disease have been reported. In mesocosm experiments yellow perch and bluegill preferred Daphnia infected with chytrid over uninfected individuals. However, in mesocosms with high concentrations of DOC (suggesting lower UV conditions) selective predation on infected Daphnia was eliminated. In this same study a comparative survey showed that lakes supporting chytrid infections had higher DOC concentrations and lower light penetration (Johnson et al. 2006). These results suggest a potential moderating influence of UV attenuation on disease incidence through either of two mechanisms. One is that infected Daphnia tend to occur at greater densities in environments that conceal their elevated visibility. The other is that UV has a stronger negative effect on the chytrid either due to lower UV tolerance, or due to the ability of Daphnia to behaviorally avoid UV damage by migrating down in the water column.

Whereas UV effects on selective vertebrate predators might inhibit disease outbreak in transparent lakes, UV transparent lakes with sizeable populations of ‘sloppy’ invertebrate predators could actually be more susceptible to parasite epidemics. While feeding ‘sloppy predators’ (e.g. Chaoborus) will often release spores contained in the prey they consume (e.g. infected Daphnia). The released spores, which are still infective, are thus re-suspended in the water column where they can be consumed by a new host. In the absence of ‘sloppy predators’ hosts dying of infection tend to sink to the lake bottom before releasing spores, effectively trapping spores in bottom sediments. Consequently, ‘sloppy predators’ tend to facilitate epidemics in lakes (Caceres et al. 2009). Elevated UV levels likely enhance the downward

migration of UV sensitive zooplankton to deeper depths where invertebrate predators reside (Boeing et al. 2004), and may therefore promote epidemics by increasing ‘sloppy predation.’

UV can influence contaminant toxicity in lakes

Organic and inorganic contaminants including PCBs, Hg, Ar, pesticides, PAHs are pervasive in lakes (Stahl et al. 2009), even in remote systems generally considered ‘pristine’ (Landers et al. 2010). UV interacts with these contaminants in several ways. For example, while UV can increase toxicity to aquatic organisms through the induction of phototoxic effects, UV can also mitigate the toxicity of mercury, one of the most problematic aquatic contaminants. DOC plays a pivotal role in these relationships as it can both decrease UV transparency of lakes, and also mitigate contaminant toxicity.

The absorption of a UV photon by a photosensitizing compound produces an excited state molecule that reacts directly or indirectly with a biological substrate (e.g. DNA) to produce an adverse phototoxic effect (Kochevar 1981). These phototoxic effects include covalent binding of photoaddition products to DNA and the formation of active oxygen species that produce photooxidized molecules (Kochevar 1981). Ultimately these phototoxic effects impair cellular function. Phototoxicity is a rather common phenomenon for compounds in aquatic environments that are not otherwise acutely toxic (Arfsten et al. 1996, Diamond et al. 2000) and it generally amplifies the negative impacts of contaminant exposure for aquatic organisms. For example, UV enhances the toxicity of some pesticides up to 10-fold (Zaga et al. 1998). For certain PAHs, UV increases toxicity by a thousand fold or more (Oris and Giesy 1985). Others have observed UV enhanced toxicity of selected antibiotics (Pandey et al. 2002), and there is evidence that some explosives, used in ammunition and then dumped in both lakes and oceans, are phototoxic (Dave et al. 2000). UV can also increase toxicity of heavy metals to a number of aquatic or semi- aquatic animals (Hansen et al. 2002, Preston 1999, Baud and Beck 2005). Sub-lethal effects have been documented for a few phototoxic compounds including a reduced ability to evade predators (Duesterloh et al. 2002), and depressed reproduction in zooplankton (Holst and Giesy 1989).

Nevertheless, the UV photochemical transformation of compounds or elements could actually reduce the negative impacts of some contaminants, like mercury. Mercury (Hg) is a pervasive threat to lakes, and is globally distributed and often transported in the atmosphere and then deposited far from emissions sources (Balcom et al. 2004). Coal combustion in electric power plants is the single largest source of Hg emissions to the air, though emissions from industrial boilers and the burning of hazardous waste are also significant sources (Lindberg et al. 2007). The effect of Hg in lakes largely depends on which form of Hg is present. For instance, methyl mercury (MeHg), the form of Hg that biomagnifies in food chains, is linked to immunologic and reproductive dysfunction in top predators (Hammershmidt et al. 2002, Drevnik et al. 2008). MeHg is also 10-100 times more toxic than the inorganic form, Hg (II) (Boening 2000). The negative effect of elemental mercury, Hg (0), in lakes is limited since it is the major component of dissolved gaseous mercury and can therefore readily volatilize from lakes.

UV plays an important role in regulating the speciation of Hg. For example, UV reduces soluble Hg (II) into Hg (0) and can therefore control evasion of Hg from lakes (Amyot et al. 1997). However, UV can also photooxidize volatile Hg (0) back to soluble Hg (II) (Lalonde et al. 2001). Consequently, the relative rates of UV mediated photoredox reactions will ultimately control Hg concentrations in lakes. Photoredox rates and the transformation of Hg may be strongly influenced by DOC concentration. For example, DOC acts as a substrate that when degraded provides photoreactive intermediates that can reduce Hg (II) to volatile Hg (0) (Garcia et al. 2005). This suggests that high DOC concentrations in lakes would increase Hg evasion and reduce in situ Hg levels. However, high DOC concentrations can also increase mercury methylation (Weber 1993), which could mean that the biological effect of Hg in high DOC lakes would be exacerbated even if overall mercury concentrations are reduced. Still other studies have suggested that DOC inhibits methylation, though the extent of methylation was dependent on both pH and salinity, with increased salinity and more neutral pH resulting in reduced rates of methylation (Barkay et al. 1997). Complicating the matter even more, a recent study in a relatively transparent lake showed that mercury flux (i.e. reduction of Hg (II) to volatile elemental Hg (0)) was driven primarily by UV exposure with little or no impact of DOC quantity or quality (Peters et al. 2007). Clearly, more research is needed to resolve whether the interaction of UV and DOC ultimately increases or decreases Hg concentrations in lakes, and to

elucidate to what extent pH, salinity, and other abiotic factors influence Hg evasion and methylation from one lake to another. Nevertheless, it appears that UV alters the bioavailability and cycling of Hg.

UV can regulate biodiversity in lakes

High levels of UV generally reduce biodiversity in lakes, but in some cases high UV might actually create conditions that promote or maintain biodiversity (Fig. 1.4). By and large, the effect of UV on biodiversity in lakes is dictated by interspecific differences in UV tolerance. For instance, zooplankton UV tolerance is strongly related to taxon and to a lesser extent body size, with rotifers and copepods usually more tolerant than cladocerans and smaller species more tolerant than large ones (Leech and Williamson 2000). These differences in taxon-specific UV tolerance have the potential to influence zooplankton community composition. For example, in Glacier Bay Alaska a study of lakes with varying degrees of terrestrial succession in their watersheds attributed differences in zooplankton species composition among lakes to the significant variation in UV attenuation from lake to lake (Williamson et al. 2001b). Likewise, in a survey of 53 temperate lakes along a UV exposure gradient, zooplankton species richness and Shannon-Weiner specific diversity were limited to minimum values when mean water column UV levels were high (i.e. average exposure levels were greater than 10% of 320 nm surface irradiance (Marinone et al. 2006). Similar results of depleted species richness and specific diversity with increasing UV exposure have been reported for artificial phytoplankton communities (Xenopolous and Frost 2003). Though the examples above demonstrate the potential for UV to reduce species diversity by limiting the number of UV intolerant taxa that occur in high UV lakes, taken in a different context the exclusion of certain UV sensitive taxa from clear waters could actually help to maintain the integrity of communities of native species and thus promote higher levels of diversity. For example, in Lake Tahoe, introduced warm-water fish (bluegill and largemouth bass in particular) have virtually extinguished native minnow species in some nearshore habitats. Experiments that isolated the effect of UV on survival of bluegill larvae demonstrated that high UV conditions may limit the susceptibility of nearshore habitat to warm-water fish invasion by preventing the survival of this critical life history stage (Tucker et al. 2010). In light of the

documented negative impact of warm-water fish on fish species richness and cyprinid richness in particular (Jackson 2002), the exclusion of warm-water fish invaders could help to maintain native biodiversity at least in more transparent lakes. UV might help to maintain species diversity in lakes through reproductive isolation as well. In Lake Victoria, decreases in UV exposure levels may help to explain the severe decline in cichlid diversity during anthropogenic eutrophication (Williamson and Rose 2009, Williamson and Rose 2010). Eutrophication in the lake has increased turbidity and seems to have interfered with mate selection by coloration and reflectance of the fish body, the mechanism that maintains reproductive isolation and thus species diversity for these cichlids (Seehausen et al. 2008). The role of UV in these declines has not been explicitly tested but many cichlid species of Lake Malawi, one of the clearest African rift lakes, do have UV photoreceptors that may function in mate selection (Jordan et al. 2006) as is the case for other freshwater fish species (Smith et al. 2002, Rick et al. 2005). Ultimately, any UV mediated changes in community composition, and in species diversity in particular, will be scale-dependent and habitat- and taxon-specific. In some transparent high elevation lakes where increased nitrogen deposition and advancing tree line point to potential declines in current underwater UV exposure (Sommaruga et al. 1999; Grabherr 1994), declining UV transparency could create a physical refuge for UV intolerant species that tend to thrive in ‘disturbed’ habitats. The replacement of highly UV tolerant endemic species with UV intolerant cosmopolitan species could subsequently reduce regional or even global diversity through biotic homogenization (McKinney & Lockwood 1999). In other lakes, like many in the Northeastern USA, precipitation is expected to increase and recovery from acidification continues (Driscoll et al. 2007, Hayhoe et al. 2007). These trends are likely to move lakes toward historical water clarity levels, decreasing underwater UV exposure and potentially increasing measures of richness and/or diversity as UV sensitive species are able to recolonize lakes in their native range. The extent to which UV mediated behavioral or mate selection pathways may influence diversity has been largely unexplored and could be a more general phenomenon with important implications for lake biodiversity worldwide.

UV can influence carbon dynamics in lakes

Exchange of carbon dioxide with the atmosphere in aquatic systems is a balance of uptake via photosynthesis and release through respiration or photodegredation of DOM into dissolved inorganic carbon. UV radiation can influence each of these processes. Because the UV mediated effects on these processes are often realized through photochemical transformation of organic material, UV will interact with DOM and may influence whether lakes are net sources

or sinks of CO2 (Fig. 1.5).

As discussed above, UV can depress algal photosynthetic capacity and overall primary production in lakes. Furthermore, UV is able to photobleach DOC, which could increase underwater UV levels, potentially exacerbating photoinhibition of algal primary production (Osburn et al. 2001, Helbling et al. 2001). On the other hand, UV wavelengths are the most effective wavelengths for converting the dissolved organic matter (DOM) pool into biologically available ammonium (Bushaw et al. 1996). In one study, UV transformed the pool of recalcitrant DOM into bioavailable forms at rates of 3% per day for nitrogen and 17% per day for phosphorus (Vahatalo et al. 2003). This means that in nutrient limited systems photochemical release of nutrients could appreciably enhance primary production. Moreover, some studies have shown that increased irradiance appears to increase the N:P ratio in phytoplankton and reduces herbivore efficiency (Dickman et al. 2006, Dickman et al. 2008). Although neither of these studies specifically assessed UV effects on phytoplankton stoichiometry or herbivore efficiency, the results are consistent with other studies demonstrating that UV can reduce P uptake in phytoplankton or increase carbohydrate storage (Hessen 1995; van Donk and Hessen 1995, though see Xenopolous et al. 2002). Low P:C ratios have also been associated with decreased activity of the enzymes that protect Daphnia from UV damage (Balseiro et al. 2009), suggesting stoichiometric constraints increase the susceptibility of this important phytoplankton grazer to UV.

Photochemical transformation of the DOM pool can also affect bacterial biovolume and overall microbial respiration. For example, although UV increases the bioavailability of nutrients for photosynthesizers, bacterial biovolume may increase by 70 % at the same time (Vahatalo et al. 2003). Bioassays showed that bacterial respiration increased in UV exposed (relative to non- exposed) treatments. This is consistent with other studies that show how photochemically

induced increases in DOM bioavailability often enhance microbial respiration (de Lange et al. 2003, Daniel et al. 2006). Thus, even if UV increases nutrient availability and subsequently enhances primary production, coincident increases in microbial respiration might still result in net carbon loss from lakes. However this may depend strongly on the source of DOM in lakes because UV-induced increases in bacterial production have been observed in lakes with more recalcitrant forms of DOM (i.e. allochthonous forms/humic lakes), but not in lakes with primarily autochthonous DOM (Tranvik & Bertilsson 2001).

UV can also convert carbon to CO2 through direct photochemical mineralization of DOM, though on a global scale this is not as significant a source of carbon flux from aquatic systems (Moran and Zepp 1997). In a humic lake UV mineralized approximately 22% of DOC to

CO2 in one week (Vahatalo 1999). However, DOC quality and quantity also matter for photomineralization. Less recalcitrant forms of DOC (i.e. autochthonous) appear to be less

photoreactive and therefore less likely to be mineralized to CO2 (Obernosterer and Benner 2004). Furthermore, UV photodegredation and the retention of carbon in lakes (i.e. storage in sediments or evasion to atmosphere) are most pronounced in low DOC systems (< 4 mg L-1; Molot and Dillon 1997). Whether carbon is stored in sediments or lost to the atmosphere is also strongly related to lake pH. In acidified lakes more carbon is evaded than stored (Anesio and Graneli 2004, Dillon and Molot 1997, Gennings et al. 2001, Reche et al. 1999), presumably because of oxidation with hydroxyl (·OH) radicals via the iron mediated photo-Fenton pathway (Voelker et al. 1997, Xie et al. 2004). In alkaline lakes it appears that non-·OH mechanisms operating in the longer wavelength UV-A region might be most important for the photo-oxidation of DOC (Molot et al. 2005).

Given the myriad complex interactions of UV and DOM, long-term field experiments that measure carbon flux under different light and DOM regimes (and that also consider additional factors like pH and the underlying mechanisms for photodegredation of DOM) could help to resolve whether UV exposure tends to promote the loss or storage of carbon in lakes. These are important questions to pursue because changes in lake carbon dynamics have the potential to influence the global carbon budget. Although they represent only a small portion of Earth’s continental land surface area (~3%, Downing et al. 2006, but see Seekell & Pace 2011

L&O 5:350) lakes appear to play an important role in carbon cycling. Freshwater aquatic systems receive 2.9 Pg C y-1 from land (Tranvik et al. 2009). Though seemingly insignificant when compared to gross carbon fluxes for the terrestrial biosphere or oceans, the emissions of inland waters (1.4 Pg C y-1) are on the same order of magnitude as carbon emissions caused by deforestation (1.6) and carbon uptake by oceans (2.6) (Burgermeister 2007). Lakes are also a substantial source of methane, one of the more potent greenhouse gases, and as such offset 25% of the estimated land carbon sink (Bastviken et al. 2011). Clearly increases or decreases in export or storage rates of carbon in freshwater could be a significant sink for anthropogenic carbon (Cole et al. 2007).

Conclusions

Ecology is rarely as simple as understanding the direct effect of one organism or factor on another. It should come as no surprise then, that some of the most interesting and important effects of UV on lakes are not the result of the direct effect of UV exposure on a single individual or group of organisms. Nor are UV effects limited to the direct negative effects on biota with which UV is so often associated. Rather, UV effects in lakes are often realized through indirect effect pathways and can induce a range of beneficial or detrimental impacts on lakes and the organisms therein. The role that UV plays in regulating animal behavior, contaminant toxicity, or food quality subsequently affects habitat selection, predator-prey interactions, disease incidence, and habitat invasibility among other things. Ultimately, these UV mediated effects influence the survival of organisms and the structure of lake communities. In some cases the interactions initiated by UV will impact biodiversity or carbon cycling, which could have important implications for these dynamics at broader scales.

Even with ‘normal’ levels of atmospheric UV (e.g. after recovery of stratospheric ozone depletion) there are still likely to be many fascinating interactive effects of UV on lakes and the organisms that reside in them. In fact, differences in the community composition of alpine and sub-alpine lakes and seasonal patterns in carbon retention in lakes are just two examples that suggest that many of these intriguing effects are already apparent across the strong

environmental gradients of UV in space and time. Climate change and other stressors are altering UV environments in lakes. For example, climate induced changes in DOC export to lakes and the eutrophication of once oligotrophic lakes by atmospherically deposited nutrients threaten to drastically alter underwater UV levels in lakes. These reductions in UV transparency may enhance invasion by warm-water species in cold clear lakes, especially in rapidly warming alpine and subalpine zones. Reductions in UV transparency also have the potential to contribute to an increase in pathogens, like Cryptosporidium, that are sensitive to UV and yet are not removed from municipal water supplies by traditional chemical water treatment methods such as chlorination. It is therefore imperative that we continue to pursue an understanding of the important direct and especially indirect effects of UV on lake ecosystems. Our ability to effectively forecast and mitigate the environmental impacts of climate change and other stressors on lakes depends on our efforts.

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Figure 1.1. Direct (bold text/solid lines) and indirect (italic text/dashed lines) effects of UV on the growth and survival of lake biota. Direct effects are only a small component of the many potential UV mediated effects on lakes. Indirect effects mediate population and ecosystem level processes through any of a number of primary and secondary filters that result in either positive or negative effects on individual growth and survival. For example, UV effects on animal behavior can influence habitat selection which can in turn induce trade-offs with predators, exposure to suboptimal temperatures, or other ecosystem components that may reduce growth and fecundity.

UV320nm as Percent of Surface 110100 0

Depth (m) Depth Lacawac (NE USA) Sparkling (N USA) 15 Alta (S New Zealand) Correntoso (W Argentina) Oesa (W Canada) Tahoe (W USA) 20

Figure 1.2. Vertical profiles of UV vs. depth showing variation in summer UV transparency among lakes around the world. Intersection of lines with the vertical axis indicate the depth to which 1% of surface irradiance penetrates, an often-used convenient metric of UV transparency. The non-log linear trends evident in some profiles indicate that transparency can vary with depth, perhaps owing to differences in photobleaching and particulate concentrations through the water column.

UV + +

+ SELECTIVE PREDATION SLOPPY PREDATION (‐) (‐)

HOST HOST COMPETENCE SUSCEPTIBILITY +

+ PATHOGEN INFECTIVITY (‐)

(‐)

DISEASE

Figure 1.3. Conceptual model of the role of UV in regulating pathogen effects in lakes. Arrows indicate effects pathways, (+) indicates a positive effect while (-) indicates a negative effect from one box to the next. UV can affect disease incidence directly, by reducing pathogen infectivity, or indirectly. Indirect effects include UV mediated changes in host susceptibility to pathogens or UV mediated effects on predation intensity by either selective or sloppy predators. UV increases feeding efficiency by selective predators which will generally reduce host body size, the major determinant of host competence as measured by potential spore production. Sloppy predators re- suspend infective spores when feeding on infected hosts. Thus, UV induced behavioral responses in hosts that increase their overlap with sloppy predators could increase disease incidence.

20 26/20 = 1.3 18 16 26/14 = 1.9 14 -div 2 =β 12 Richness 10 20 =α-div 26/8 = 3.3 8

Species 6 12 26/6 = 4.3 6 4 6 =α-div 2 UV TOLERANT 2 UV INTOLERANT 0 Low Moderate High Extreme UV transparency

Figure 1.4. Conceptual model showing species richness across a UV transparency gradient within a landscape. At the landscape level, species richness (i.e. γ- diversity) is 26 because there are 20 UV intolerant and 6 UV tolerant species. Within-lake richness (i.e., α-diversity) tends to decrease with increasing UV transparency as habitats become less suitable for otherwise competitively superior UV intolerant species. Along the UV transparency gradient the turnover in species composition among different lakes contributes additional diversity (i.e. β- diversity = γ/α) to the region so that the effective number of communities is greatest when all lake types are present on the landscape.

Figure 1.5. Conceptual model of possible effects of UV on carbon cycling in lakes. Arrows indicate effects pathways, (+) indicates an increase while (-) indicates a decrease from one box to the next (e.g., photobleaching of DOM increases photoinhibition, which in turn decreases photosynthesis). UV effects on carbon cycling are largely regulated by the DOM pool in lakes.

Chapter 2

Ultraviolet radiation affects invasibility of lake ecosystems by warmwater fish

As published:

Abstract

Predicting where species invasions will occur remains a substantial challenge in ecology, but identifying factors that ultimately constrain the distribution of potential invaders could facilitate successful prediction. Whereas ultraviolet radiation (UVR) is recognized as an important factor controlling species distribution and community composition, the role of UVR in a habitat invasibility context has not been explored. Here we examine how underwater UVR can regulate warm-water fish invasion. In Lake Tahoe, California and Nevada, USA, established populations of exotic bluegill sunfish (Lepomis macrochirus) are currently limited to turbid, low-UVR embayments. An in situ incubation experiment that manipulated incident UVR exposure of larval bluegill, combined with an assessment of UVR exposure levels in nearshore habitats around Lake Tahoe, demonstrates that UVR can mediate habitat invasibility. Our findings suggest that the susceptibility to invasion by UVR sensitive species may increase in transparent aquatic systems threatened by declining water quality, and they highlight the importance of abiotic factors as regulators of invasion risk in ecosystems.

Introduction The proliferation of invasive species is one of the most important anthropogenic impacts in freshwater systems (Naiman et al. 1995). The problem is largely a byproduct of human development, with its tendency to deconstruct biogeographic barriers (Rahel 2007) and fundamentally alter the biotic and abiotic components of environments that foster distinct populations of plants and animals and regulate the susceptibility of habitats to invasion. Consequently, habitat invasibility is generally thought to be high in areas characterized by extensive human impact. For example, among California, USA watersheds, the number of nonnative fish species is positively correlated with anthropogenic landscape-level changes related to watershed disturbance and altered hydrology (Marchetti et al. 2004). Reservoirs are also a notable example of how human activity may promote invasion (Havel et al. 2005). These examples highlight important factors that are likely to control invasibility in some habitats but they are driven by more traditional notions of human impact, such as the stabilization of flow regimes related to habitat alteration or the influence of a high degree of environmental variability through time. Whereas changes in water transparency with anthropogenic disturbance are widely recognized in aquatic habitats, little attention is given to how such disturbances can mediate exposure to damaging wavelengths of ultraviolet radiation (UVR). Here we demonstrate the potential importance of UVR exposure as a factor controlling habitat invasibility of a warmwater fish in Lake Tahoe. Lake Tahoe is a sub-alpine lake in the northern Sierra Nevada range spanning the California–Nevada border. The lake is renowned for its deep blue water and high transparency, afforded by the combination of great depth, small watershed-to-lake-area ratio, and granitic geology of the basin (Jassby et al. 1994). However, the transparency has decreased over time with the average annual Secchi transparency declining from 31 m in 1968 to 21 m by 1998 (Jassby et al. 1999). During this same 30-year interval, a number of nonnative warm-water fish species established populations in some portions of Lake Tahoe (Reuter and Miller 2000). The establishment of these warm-water species may be directly related to the significant changes in water transparency observed in recent decades (Fig. 2.1). For example, larval bluegill sunfish (Lepomis macrochirus) perish within a single day when exposed to incident UVR at the surface of transparent lakes (Williamson et al. 1999). Yet the requirement for warmer spawning

temperatures constrains bluegill nests to the shallow surface waters in the littoral zone of lakes and rivers (Kitchell et al. 1974). Thus the transparency of the water as well as the depth and location of nests are critical determinants of reproductive success in bluegill (Olson et al. 2006). Currently, the only well-established bluegill populations in Lake Tahoe are limited to sites in the southern end of the basin characterized by extensive development and in close proximity to some of the lake’s largest tributaries (Kamerath et al. 2008). Water transparency at these sites is low and may explain their suitability for the UVR-sensitive bluegill. Our primary aim was to explicitly test the hypothesis that UVR controls the suitability of nearshore habitats for the earliest life history stages of exotic bluegill. We were also interested in understanding what controls the UVR transparency of nearshore habitats in Lake Tahoe. The decline in visible light transparency has been attributed to increases in both biological (i.e., phytoplankton and detritus) and inorganic (i.e., terrestrial sediment) particulate matter (Swift et al. 2006) resulting largely from human impacts in and around the basin related to eutrophication (Goldman 1988) and stream bank erosion (Byron and Goldman 1989). However, the attenuation of UVR in freshwater lakes is strongly regulated by chromophoric dissolved organic matter (CDOM) (Morris et al. 1995, Williamson et al. 1996). CDOM may be especially important in nearshore habitats where fish spawning occurs, since CDOM inputs are likely to be concentrated in those areas. For example, in Lake Tahoe stream water inputs of CDOM are approximately 10 times higher than CDOM levels offshore where most of the long term transparency monitoring has been conducted (Swift 2004). An understanding of the mechanisms underlying UVR transparency in Lake Tahoe could enable us to better understand how regional and global environmental changes related to the factors that mediate UVR transparency could in turn affect habitat invasibility in this large, highly transparent lake.

Methods To test the hypothesis that UVR controls the suitability of nearshore habitats for bluegill invasion we measured UVR exposure at multiple nearshore locations around the perimeter of the lake using DNA dosimeters (Fig. 2.2). In two of these nearshore locations, we carried out a four day in situ incubation experiment that manipulated the incident UVR levels to which larval bluegill were exposed. DNA dosimeters were also deployed with the larval bluegill in these in

situ incubations as a means for comparing levels of DNA damage in dosimeters with larval bluegill mortality. Standardized DNA damage values obtained from dosimeters incubated alone around the lake were compared to DNA damage values from dosimeters included with larval bluegill to evaluate the potential for larvae to survive in multiple nearshore locations. To assess the relative importance of dissolved organic carbon and chlorophyll as regulators of the UVR environment in nearshore areas of Lake Tahoe we measured levels of these light attenuating components at 13 nearshore sites, including each of the sites where we deployed dosimeters.

Larval incubation experiment Larval yolk sac bluegill were collected from a single nest at approximately 1 m depth in the Tahoe Keys on 17 July 2007. Larvae (n=5) were placed in Whirl-Pak bags (Nasco, Fort Atkinson, Wisconsin, USA) filled with 100 mL of 48-μm filtered lake water to exclude most zooplankton. To isolate the effect of UVR between incubation sites, the Whirl-Pak bags were either shielded from incident UVR in Courtgard (CP Films, Martinsville, Virginia, USA) sleeves or exposed to incident UVR in Aclar (Honeywell International, Morristown, New Jersey, USA) sleeves. Courtgard is a long-wave-pass plastic that transmits photosynthetically active radiation (PAR; 95% 400–800 nm in water) but blocks most UVR (transmits no UV-B 295–319 nm, and only 9% of UV-A 320–400 nm with a sharp wavelength cutoff and 50% transmittance at 400 nm). Aclar is a long-wave-pass plastic that in water transmits both PAR (100% 400–800 nm) and most UVR (98% of UV-B 295–319 nm, 99% UV-A 320–399 nm, with a sharp wavelength cutoff and 50% transmittance at 212 nm). The two incubation sites for the larval exposure experiment included waters with low and high UVR transparencies, that is, the Tahoe Keys and Sand Harbor areas, respectively. Four replicates of each of the UVR shielded and unshielded treatments were deployed at dusk on 17 July at one meter depth in both the high (Sand Harbor) and low (Tahoe Keys) UVR sites and retrieved early on the morning of 21 July. After collection, larvae were examined under a dissecting microscope and scored as live if a heartbeat was observed. The 84 hour incubation period used here is similar to the time it takes larvae to reach swim-up stage and leave the nest (Gross and MacMillan 1981). Average daily water temperature at 1 m over the course of the incubation was 20.4°C in Sand Harbor and 22.4°C in Tahoe Keys.

These temperatures are well within threshold temperatures for bluegill spawning (Wallus and Simon 2008). All procedures involving animals were in accordance with the policies set forth by Miami University’s Institutional Animal Care and Use Committee (IACUC protocol #683). Both incident and submersible radiometers and DNA dosimeters were deployed to measure incident UVR and water transparency during the incubation. Underwater solar radiation was measured at each site with a BIC profiling UVR-PAR radiometer (BICLogger; Biospherical Instruments, San Diego, California, USA). This instrument quantifies incident solar irradiance at three different UVR wavelengths (305, 320, and 380 nm) as well as visible wavelengths of photosynthetically active radiation (PAR, 400–700 nm). Transparency data from BIC profiles

were used to calculate diffuse attenuation coefficients (Kd) for each site and were combined with cumulative surface irradiance data measured with a Biospherical Instruments BICLogger, a multichannel, internally recording radiometer of a similar design and specifications to the BIC, to estimate total exposure for the duration of the incubation experiment. Two DNA dosimeters were included with fish in each of the four UVR unshielded bags and in two of the four UVR shielded bags at each site during larval fish incubations. Logistic regression analysis of a 22 factorial design was performed using SAS 9.1 (SAS Institute 2005) to test for main effects of site, and UVR+ or UVR- microcosm on larval survival.

DNA dosimeters The DNA dosimeters were 10 mm diameter by 40 mm long quartz cuvettes filled with 0.4 mL of raw salmon testes DNA solution diluted to 100 μg/mL in double distilled water and sealed on each end with silicone stoppers and parafilm. DNA in dosimeters accumulates damage as a function of UVR exposure, and the frequency of cyclobutane pyrimidine dimers (CPDs) per mega-base (Mb) of DNA, the most common photoproduct, were estimated using radioimmunoassay (RIA) as described by Mitchell (Mitchell 1996, 1999). Dosimeters were included with fish in larval incubation experiments as described in Methods: Larval incubation experiment. Dosimeters (n=4) were also deployed alone for approximately four days during the period between 15–20 July 2007 at 1- and 2-m depths at 10 sites around Lake Tahoe, including the two larval incubation sites. These dosimeters were placed in Whirl- Pak bags filled with 100 mL of 48-μm filtered lake water and then inserted into Aclar

sleeves. Because the dosimeters deployed alone were incubated at multiple depths and for various times a relationship between 320- nm UVR exposure and DNA damage values was derived and then used to estimate DNA damage for each site around the lake based on a standard 320-nm UVR exposure. We used a standard exposure equivalent to the cumulative incident (i.e., surface) 320-nm UVR exposure for larval incubation experiments, so that estimated DNA damage values for sites around the lake are comparable to DNA damage values from dosimeters that were included in the larval fish incubations. In more detail, the estimated DNA damage values based on a standard exposure were derived as follows. First, using UVR exposure data and DNA damage values obtained during dosimeter incubations 15–20 July, the relationship of 320-nm UVR exposure to measured DNA damage was derived (Fig. 2.3). A linear relationship between measured DNA damage and UV 320-nm exposure was observed and the equation describing this relationship was derived by simple linear regression (y=27.408x, R2 = 0.9622). The cumulative 320-nm UVR exposure for the duration of the dosimeter deployment at each given site was derived as described above for the larval incubation experiment, combining transparency data from BIC profiles at a given site with cumulative surface irradiance data measured with the BICLogger. Measured DNA damage values for each site were plotted as the average of dark corrected dosimeter values in CPDs/Mb DNA from the RIA. Next, the 320-nm UVR exposure at 1 m depth for a given site was calculated for a standard surface exposure. The 1 m depth is equivalent to the depth at which dosimeters were deployed with the larval fish. The standard surface exposure was 26.3 kJ/m2, the mean value of 320-nm surface exposure for the duration of the larval fish incubations. For reference, the highest incident UVR exposure that we observed with the BIC logger at Lake Tahoe for a single, full day of 320-nm UVR exposure was 9.73 kJ/m2 on 17 July 2007. This is essentially equivalent to one exposure day (sensu Williamson et al. 1999). The estimated 320-nm 2 UVR exposure at 1 m depth for each site (E1ms) for the 26.3 kJ/m standard surface exposure was then derived from the equation that describes the attenuation of light in water:

-Kdz E1ms= E0std x( e )

2 where E0std = 26.3 kJ/m , integrated 320-nm surface irradiance for the duration of the larval fish

dosimeter deployment; Kd= diffuse attenuation coefficient, derived as the slope of the semi-log plot from vertical UVR profiles at a given site; and z = 1 m, standard exposure depth.

Finally, this E1ms value was inserted into the equation describing the relationship of 320-nm UVR exposure to DNA damage (i.e., y=27.408x). The resulting value (y) was the estimated DNA damage value based on this relationship (from Fig. 2.3).

DOC and chlorophyll a analysis Water samples were collected in pre-rinsed 1-L polyethylene bottles from within the mixed layer. Water used in dissolved organic carbon (DOC) analysis was filtered through pre- ashed 25-mm 0.7 μm Whatman GFF filters within 8 hours of sample collection using a glass filter support. The filtered sample was stored in the cold and dark in 40-mL glass bottles until

analysis. The DOC samples were analyzed with a Shimadzu TOC-VCPH analyzer (Shimadzu, Columbia, Maryland, USA) within one week post sampling. For chlorophyll a, 100 mL of the water sample was filtered through pre-ashed 25-mm 0.7-μm Whatman GFF filters within 8 hours of collection and the filter was immediately frozen until chlorophyll analysis. Chlorophyll a extraction was completed with an acetone–methanol mixture and chlorophyll a concentration was completed via fluorometry within one month of sample collection. UVR attenuation was also measured at each site with the BIC profiling radiometer, and diffuse attenuation coefficients

(Kd) were calculated for each site from the slope of the natural log relationship of UVR irradiance vs. depth. Using SAS v. 9.1, we performed a likelihood ratio test to compare models

that predicted Kd,320nm from DOC and/or chl a values.

Results For the larval incubation experiment, exposure to 320-nm radiation in unshielded treatments was nearly 40x higher in the Sand Harbor site (22.65 kJ/m2) compared to the Tahoe Keys site (0.60 kJ/m2, Fig. 2.4A). The mean DNA damage levels, measured in DNA dosimeters as the frequency of cyclobutane pyrimidine dimers (CPDs), at the Sand Harbor site were more than 30x higher than those measured at the Tahoe Keys site (729 vs. 22 CPDs/Mb DNA, Fig. 2.4B). Larval survival was inversely related to UVR exposure with 84% of larvae surviving in

unshielded microcosms in the low UVR site and only 11% survival in the high UVR site (Fig. 2.4C). For bluegill in unshielded UVR microcosms, there was a statistically significant effect of site on larval fish survival (PROC LOGISTIC; P , 0.0001). In the UVR-shielded microcosms, larval survival was high (90– 100%) at both sites. DNA damage measured in the dosimeters also increased with increasing UVR transparency across the ten sample sites (Fig. 2.5). In seven of the 10 sample sites DNA damage levels were higher than those measured at the Tahoe Keys, where bluegill survival was high. Indeed, the majority of sites showed DNA damage levels above the threshold for larval survival (Figs. 2.4B and 2.5), implying high potential UVR induced mortality in bluegill at most sample sites. The 1% attenuation depths, that is the depth where 320-nm UVR reaches 1% of surface irradiance, show the wide range of UVR transparency of nearshore sites in Lake Tahoe (Table 2.1). UVR (320 nm) transparency of the near shore sites was strongly dependent upon DOC 2.53 2 (Kd,320nm = [2.57 x DOC ]; R = 0.81). However, a model that included both DOC and 3.01 chlorophyll a (Kd,320nm = [1.95 x DOC ] + (0.02 x chl a]) was the best predictor of UVR attenuation (R2 = 0.98) for the sites sampled (likelihood ratio chi-square = 11.2, df = 1, P = 0.0008).

Discussion In this study, dosimeters of raw DNA in solution were used as tools to assess potential UVR effects on larval bluegill by relating DNA damage levels in dosimeters with larval fish mortality. The observed levels of DNA damage in the dosimeters support the hypothesis that UVR is a potent force contributing to the suitability of nearshore habitats for successful bluegill reproduction. Current UVR conditions were substantial enough to reduce reproductive success of bluegill in the majority of nearshore sites sampled. Both DOC and chlorophyll a were important regulators of variation in the UVR environment in nearshore areas of Lake Tahoe. This suggests that effective regulation of chlorophyll and DOC inputs could stem future declines in UVR transparency in Lake Tahoe and in turn help mediate habitat invasibility. Our study was motivated by a framework for predicting species invasion that highlights the importance of identifying the specific abiotic factors that will ultimately constrain

distribution in an invaded range. Current approaches for predicting habitat invasion tend to rely on correlating species’ distribution with selected habitat parameters that implicitly incorporate biotic constraints on distribution. These biotic constraints may not always be present in an uninvaded range (Kearney and Porter 2004). It has been argued that a more powerful approach is to identify specific abiotic factors with demonstrable fitness consequences for an organism, and then map the fitness consequences (e.g., survival or reproduction) at various levels of the abiotic factor onto the landscape (Kearney 2006). This kind of approach is fundamental if we wish to improve our confidence in extrapolating species’ potential distributions to novel circumstances under climate change scenarios, and it could be especially useful for predicting invasions in systems where a specific factor regulating invasion (e.g., UVR) is closely tied to a global change element (e.g., climate driven changes in DOC). In our study, we have accomplished the crucial first step in this approach by demonstrating that UVR is a key abiotic factor with the potential to constrain the reproductive success of bluegill in Lake Tahoe. By identifying some of the key mechanisms underlying UVR transparency we have also increased our understanding of how regional and global environmental changes related to the factors that mediate UVR transparency could in turn affect habitat invasibility in this lake. We suspect that this framework and our results could be directly relevant to other transparent lakes. Whereas few lakes are as highly transparent as Lake Tahoe, estimates from DOC measurements in North American lakes indicate that UVR transparency is relatively high throughout western, northwestern, and southeastern portions of the USA (Williamson et al. 1996). For example, based on modeling the relationship between DOC concentration and UVR attenuation, the depth to which 1% of 320-nm UVR surface irradiance penetrates is greater than 1 m in 75% of lakes sampled in the western United States. About 25% of these lakes exhibit 1% UVR depths greater than 4.75 m. This is noteworthy because bluegill generally nest at depths less than 4 m (Carlander 1977), and other studies have demonstrated significant UVR effects on reproductive success of temperate fish species (including bluegill) in the eastern USA in lakes with a 1% UVR depth not in excess of 4.9 m (Huff et al. 2004, Olson et al. 2006). The DOC concentration in most of the transparent lakes referenced above is quite low (i.e., <1 mg/L), suggesting that even small changes in DOC could significantly reduce current UVR levels in these lakes (Williamson et al. 1996). Although there are no specific predictions

for future DOC concentrations in western and southeastern U.S. lakes, widespread and strong trends of generally increasing DOC concentrations have been observed in lakes and rivers elsewhere (Evans et al. 2005, Monteith et al. 2007). Therefore it is reasonable to consider the potential for substantial changes in UVR transparency, and consequently habitat invasibility in these transparent lakes. Just as relevant and better documented in high elevation transparent lakes like Tahoe are trends of increased algal growth and reduced water clarity as a consequence of increased nitrogen deposition (Jassby et al. 1994, 1995, Sickman et al. 2003). These trends, documented in the western United States, are predicted to continue across that region (Lamarque et al. 2005). Chlorophyll has a proportionately greater effect on UVR attenuation in low-DOC systems (Laurion et al. 2000, Sommaruga and Augustin 2006). Consequently variations in chlorophyll levels, like changing DOC concentrations, have the potential to modify transparency in very low DOC lakes. This in turn could facilitate the establishment of exotic species in formerly unsuitable habitats. One critical question pertinent to the role of UVR in mediating habitat invasibility in transparent lakes is whether adult bluegill are able to respond to these selective pressures by reducing UVR exposure through either nesting deeper or shifting their spawning time to coincide with periods of decreased water transparency. For Lake Tahoe, this seems an unlikely possibility. First, in this study the 1 m depth and the seasonal timing of our experiments were consistent with actual nest depths and nesting times in Lake Tahoe. Moreover, later spawning times, coincident with increasing water temperatures that might allow bluegill to nest at greater depths, are unlikely to decrease UVR exposure because UVR transparency (320 nm) actually increases on the order of 20–90% from spring to summer as allochthonous inputs decrease in the nearshore (Rose et al. 2009). On the other hand, accelerated spawning phenologies that could potentially enable bluegill to benefit from decreased water transparency earlier in the growing season are likely constrained by thermal conditions required for spawning. Bluegill are reported to spawn at temperatures from 15.6°C to 32°C, with optimum spawning temperatures in the range of 21°C to 24°C (Wallus and Simon 2008). Surface water temperatures measured at an index site in May 2007 and 2008 never exceeded 11.1°C, well below the minimum spawning temperature. Even in June, the maximum surface water temperature over this two year period was 17.2°C (C. E. Williamson, unpublished data), still below the optimal spawning temperature for bluegill. Thus

the primary opportunity for invasion is likely to be in the shallow nearshore embayments where both water temperatures are high enough and transparency to UVR is low enough to permit adult spawning and larval survival. It is also important to note that we have used the most severe response metric (i.e., mortality) in evaluating the role of UVR for regulating habitat suitability for larval bluegill invasion. Consequently, our study likely underestimates the full extent of UVR induced effects on larvae when considered in terms of the interactions of sublethal effects with sources of background mortality in developing larvae and other ‘‘life history bottlenecks’’ that young fish face. For example, UVR exposure impedes larval growth in a variety of fish species (Hunter et al. 1979, Winckler and Fidhiany 1996, Vehniainen et al. 2007) and body size in young fish, including bluegill, is a critical determinant of overwinter survivorship and mortality due to predation (Belk and Hales 1993, Cargnelli and Gross 1996). Other potential sublethal UVR effects that may ultimately reduce bluegill survival include diminished immune system function and increased incidence of infectious disease resulting from ‘‘sunburn’’ (Salo et al. 1998, Nowak 1999), developmental anomalies that might increase susceptibility to predators (Vehniainen et al. 2007), indirect trophic mediated UVR effects on food availability (Williamson et al. 1994, Zagarese and Williamson 2001), or phototoxic effects (Bullock and Roberts 1979, Oris and Geisy 1987). It is unclear to what extent UVR may play a role in the invasion ecology of other invasive species or other life history stages. We contend that it could have relevance for any UVR sensitive species that is constrained to shallow water environments by, e.g., requirements for warmer spawning temperatures in clear, cold-water lakes. For older more tolerant and mobile life history stages other biotic and/or abiotic factors (e.g., food availability or habitat structure) likely play a more important role in determining habitat suitability. However, we have emphasized the earliest life history stages here for two reasons. First, early life history stages are less pigmented, less mobile, and thus likely to be less well defended against UV damage. Second, in a biological invasion context the naturalization and eventual invasion of a species in a novel environment depends critically on the ability of that species to establish self perpetuating populations through successful reproduction (Richardson et al. 2000). Whereas other among habitat characteristics may be important in regulating species invasions, we have shown that for these critical early life

history stages UVR alone is an adequate determinant of habitat suitability and thus a potential regulator of habitat invasibility. The extent to which UVR ultimately controls bluegill invasion in Lake Tahoe or any other system will depend upon the potential for these organisms to adapt to local conditions. It is possible for example, that constitutive levels of maternally derived photoprotective compounds (PPCs) could increase in larval fish spawned in high UVR environments, thereby increasing UVR tolerance and the ability to spread into new habitats. High UV environments tend to stimulate PPC synthesis by algae and bacteria. These can be transferred in food chains and accumulated at higher trophic levels by organisms that have such capability, which may in turn be enhanced by UV exposure (e.g., copepods [Tartarotti et al. 2004, Moeller et al. 2005, Tartarotti and Sommaruga 2006]; coral reef fish [Zamzow 2004]). Environmental stress is often considered a driver of adaptation during invasion and it has been demonstrated that abiotic conditions can select for adaptive genotypes in invasive species (Lee et al. 2007). Our data suggest that UVR can similarly act as a selective force in highly transparent systems, and the potential for the development of more resistant genotypes could be tested. Future research concerning the role of UVR in controlling biological invasion should consider these and other possibilities. Nevertheless, we have shown that for the current bluegill population in Lake Tahoe UVR is a potent stressor that mediates habitat suitability for larval fish in nearshore areas and therefore controls habitat invasibility. Further efforts to quantify the effect of abiotic controls on the growth, survival, and reproduction of organisms and to map those effects onto the landscape will help us to more accurately predict the full potential of species invasion in imperiled environments. Knowledge of the particular levels of important abiotic factors that reduce the fitness of non-natives could also enable us to manage abiotic conditions in habitats for the prevention of species invasion (Alpert et al. 2004). In lakes, for example, one goal might be to establish and manage UVR transparency thresholds that prevent the establishment of non-native species by inhibiting successful reproduction. We suggest that future studies in highly transparent aquatic ecosystems consider UVR and other abiotic habitat features as important factors controlling habitat invasibility and invasion risk.

Acknowledgements We thank Geoff Schladow, Monika Winder, Sudeep Chandra, and Marcy Kamerath for logistical support. We are also grateful to Anne Liston and the staff of the Tahoe Environmental Research Center for their assistance. Thanks to Neil Winn for GIS mapping assistance and Michael Hughes of the Statistical Consulting Center of Miami University for data analysis assistance. This work was supported in part by Miami University’s Field Workshop Program (field expenses), NSF IRCEB 0210972 (UV radiometers), and NIEHS Center grant ES07784 (DNA analysis).

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Figure 2.1. Nearshore sites in Lake Tahoe with contrasting (left) high and (right) low UVR transparency. Photo credits: left, Carrie Kissman; right, A.J. Tucker

Figure 2.2. Map of Lake Tahoe indicating the location of nearshore sites where DNA dosimeters were deployed. Site numbers correspond to those plotted in Fig. 2.5. Sites where larval incubations were deployed are indicated by a dagger.

Figure 2.3. Relationship between measured DNA damage and 320-nm UV exposure at 10 sites in Lake Tahoe. DNA damage values were plotted for each site as the average of dark- corrected dosimeter values (cyclobutane pyrimidine dimers [CPDs]/Mb DNA from the radioimmunoassay [RIA]) for dosimeters deployed during the one-week sampling period, 15-20 July 2007. The 320-nm UV exposure for the duration of the dosimeter deployment at a given site was derived by combining transparency data from BIC UVR-PAR profiles at a given site with cumulative surface irradiance data measured -K Z with the BICLogger, according to the equation EZ = E0 x (e d ), where EZ = 320-nm 2 2 UVR irradiance in kJ/m at depth z(m) and E0 = 320-nm UVR irradiance in kJ/m at the water surface. Dosimeters were deployed at more than one depth at some sites.

Figure 2.4. (A) Ultraviolet radiation (UVR) exposure levels, (B) DNA damage, and (C) survival of bluegill larvae in experimental microcosms. UVR exposure levels in the experimental microcosms were estimated from incident UVR measurements. DNA damage was measured in DNA dosimeters incubated in microcosms with larvae. Survival of bluegill larvae was measured after an 84- hour incubation in low (TK, Tahoe Keys) and high (SH, Sand Harbor) UVR sites when shielded (-) and unshielded (+) from incident UVR. For panels B and C, bars indicate maximum and minimum values within treatments. Boxes indicate the median and 25th and 75th percentiles.

Figure 2.5. Relationship showing the increasing DNA damage (measured as CPDs/Mb DNA) with increasing UVR transparency (indicated here by the percent of incident surface 320- nm UVR present at 1 m depth). DNA damage values are estimated from a derived exposure vs. dosimeter value relationship and are standardized for depth and deployment duration for comparison to larval fish incubation experiments. Site numbers appear above data points: site 2 is the low-UVR Tahoe Keys; site 9 is high- UVR Sand Harbor.

Chapter 3

The invasion-window for warmwater fish in a large sub-alpine lake: the role of UV radiation and temperature.

To be submitted: Biological Invasions

Abstract

The success of biological invaders depends on the suitability of habitats for invasion (habitat invasibility) and the characteristics of invaders (species invasiveness). However, very few studies have attempted to integrate habitat invasibility and species invasiveness to explain invasion success/potential. In this study we develop an invasion-window model which predicts that largemouth bass (Micropterus salmoides) establishment in a large sub-alpine lake (Lake Tahoe, CA/NV, USA) is controlled by the ability of these non-native fish to tolerate ultraviolet radiation (UV) and temperature stress (species invasiveness) along a UV-temperature stress gradient (habitat invasibility). We tested the UV tolerance of pre-swimming yolk-sac largemouth bass larvae collected from Lake Tahoe and we used data from the literature detailing the effect of temperature on hatching success of bass embryos to parameterize the invasion-window model. The model shows that UV and temperature conditions substantially influence larval bass survival. In situ incubation experiments confirmed model predictions. UV exposure and in situ incubation experiments with native Lahontan redside minnows (Richardsonius egregius) showed that the native species overlaps broadly with the range of ambient environmental conditions in the Tahoe littoral zone. To our knowledge this is the first time that the roles of UV and temperature have been considered with respect to species invasion. As such, this research could provide important insights into the potential for warmwater species invasion in lakes, especially for coldwater lakes characterized by exceptional clarity but threatened by both increasing temperature and declining transparency. We discuss the potential for the invasion-window model to be applied as a powerful tool for exploring some unanswered questions in invasion biology, including the role of alternate stable states in facilitating invasion and the potential for invasive species to promote ‘invasional meltdown’ through their impacts on water clarity and temperature.

Introduction The factors that control the potential for exotic species to invade and establish have been viewed as components of an “invasion-window” (Johnstone 1986) or “niche opportunity” (Shea and Chesson 2002) that can be altered by disturbance. It is generally assumed that high stress environments are less susceptible to invasion (Baker 1986, Alpert et al. 2000). Therefore, disturbance is necessary to facilitate invasion in high stress habitats and thus open an invasion- window. In other words, habitat invasibility (i.e. the susceptibility of a habitat to invasion) is low when stress is high. Biological invasion theory also emphasizes the importance of species invasiveness (i.e. species traits that enable invasion) when considering the probability of invasion (Williamson 1996). More recent theory suggests that the probability of invasion depends not on habitat invasibility or species invasiveness alone, but rather on the fit between a particular non- native species and a particular habitat, the so called ‘key-lock’ approach (Cornelius et al. 1990, Higgins and Richardson 1998). Consequently, some general theories of invasion have attempted to unify these two central organizing themes in invasion ecology (e.g. Sher and Hyatt 1999, Davis et al. 2000). However, habitat invasibility is often thought to depend on an abiotic factor (i.e. resource supply) whereas species invasiveness is contingent upon biotic forcing (i.e. a species’ ability to compete for resources), as in the often cited study of Davis et al. (2000) on plant invasion. Here we propose an example where abiotic forcing is paramount with respect to both habitat invasibility and species invasiveness. We test a conceptual model that predicts warmwater fish invasion potential in a large sub-alpine lake (Lake Tahoe, CA/NV, USA) is controlled by the ability of non-native fish to tolerate ultraviolet radiation (UV) and temperature stress along a UV-temperature stress gradient resulting from anthropogenic disturbance in the nearshore environment (Fig. 3.1). This model is consistent with recent findings that demonstrate the central (sometimes exclusive) role of abiotic factors as regulators of biological invasion (Holway et al. 2002, Dethier and Hacker 2005, Menke and Holway 2006, Gerhardt and Collinge 2007). The conceptual model applied to Lake Tahoe extends disturbance/stress models, originally developed in plant ecology, to explain the limited invasiveness of warmwater fish in clear, high-UV environments. In Lake Tahoe, the warmer waters required for non-native fish spawning only occur in the shallow nearshore environment where exposure to UV is potentially

high. Thus, UV and temperature interact to define a high stress environment accommodating a limited, vigorous native community but probably excluding many potential invaders in the absence of disturbance. According to this model, historical conditions in Lake Tahoe likely excluded warmwater fish from spawning due to the high levels of damaging UV radiation in the warmer, shallow, nearshore habitats. Human and natural disturbance have expanded the physical habitat conditions to include warmer temperatures and lower UV transparency, and thus an “invasion-window” has opened where warmwater fish can successfully survive (a “survival window”) and reproduce (an “establishment window”). The establishment window is especially important from a biological invasion perspective because the naturalization and eventual invasion of a novel environment depends critically on the ability of a species to reproduce (Richardson et al. 2000). Here we examine largemouth bass (Micropterus salmoides) establishment in Lake Tahoe as a case study to test our conceptual model. Our goals are to, 1) extend the ‘invasion-window’ concept and disturbance/stress models to aquatic systems, 2) evaluate the role of abiotic forcing (in particular the roles of UV and temperature) in determining invasion potential, and 3) integrate habitat invasibility (along a UV stress gradient) with species invasiveness (the ability of a species to cope with UV and temperature stress) to predict invasion potential in Lake Tahoe and potentially in other clear, cold lakes. To accomplish these goals we tested the hypothesis that UV will induce mortality in pre-feeding larval bass (relevant to the establishment window) at natural UV exposure levels that are not lethal to native Lahontan redside minnows (Richardsonius egregius). We predicted that reproductive success for non-native bass would vary as a function of UV transparency and temperature in nearshore locations. We use in situ incubations of larval fish to corroborate predictions of reproductive success.

Methods

To test the hypothesis that UV exposure induces mortality in pre-feeding larval bass at levels that are not lethal to native minnows we conducted UV exposure experiments in June and July 2009 under ambient solar conditions at Lake Tahoe. The exposure experiments proceeded as follows: embryos of either bass or redside minnows were collected from locations in Lake Tahoe and hatched in the laboratory. Within a day or two of hatching the larvae were added, across

treatments, to each of 12 treatment dishes and 3 control dishes. Neutral density stainless steel mesh screens were placed over treatment dishes to provide a range of UV exposure levels (100%, 78%, 57%, and 45% of ambient solar UV). Treatments were then placed in an outdoor temperature controlled water bath (at the hatching temperature) and larval survival was monitored every 2 hours for up to 6 days. Control dishes were placed in the dark in a temperature controlled environment. Larval survival (i.e. presence of heartbeat) was scored with the aid of a dissecting microscope. Ambient UV305 nm exposure for the duration of the experiment was recorded with a BIC-logging radiometer (Biospherical Instruments, San Diego, California, USA). A generalized linear mixed effects model was used to determine the effect of UV exposure on the cumulative proportion dead for each larval species. Mixed effects models allow for fixed and random effects (Verbeke and Molenberghs 1997). The fixed effects included species type and exposure level. Because of the random effect of replicate differences the model was fit using PROC GLIMMIX (SAS v. 9.2). The model accommodates dose/species interaction effects. The indicator variable coding treats redside minnow as a “baseline” species. The model specification was:

where:

= empirical odds of mortality

Exp = cumulative UV Exp

ILMB = 1 for largemouth bass, 0 otherwise

b0 = random intercept effect due to chamber

ε = random error

Information on the effects of temperature on the larval life history stage of bass is limited. Therefore, temperature effects on bass survival were estimated from a published study on the effects of incubation temperature on hatching success of largemouth bass eggs (Kelly 1968). A 76

local polynomial regression model (Cleveland and Devlin 1988) was fit to the data using R software (function loess, stats package; R Development Core Team 2011). Predictions of survival at 0.01 ° C increments over the range 10-29° C were also evaluated. To test the prediction that bass reproductive success would vary with UV and temperature conditions, we developed a simple model based on the UV and temperature responses of bass larvae. We used this model in two ways. First, we estimated reproductive success in 3 nearshore sites at 0.1 m intervals up to 1.2 m over the course of the spring/summer season. This approach allowed us to explore how reproductive success of bass varied both temporally and spatially in the lake. Second, we assessed the potential for larval bass survival in 11 sample sites over the range of UV exposure and temperature conditions observed from 4 years of sampling data in these nearshore locations. This allowed us to examine invasion potential on a broader spatial scale. The first approach paired the experimentally derived UV exposure and temperature response curves with field survey data of UV exposure and temperature to estimate mean survival of bass larvae at three nearshore sites. The response of larval bass to UV and temperature (i.e. species invasiveness) was predicted from the temperature controlled outdoor exposure experiments and from the literature (as described above; see Fig. 3.2a, 3.2b). Temperature and UV conditions at each site (i.e. habitat invasibility) were directly measured (for temperature) or calculated (for UV) from once monthly profiles with a BIC profiling radiometer (Biospherical Instruments, San Diego, CA). For Tahoe Keys and Sand Harbor, diffuse attenuation coefficients (kd) for 305 nm UV and water temperature were measured once per month from May-August 2009. UV and temperature conditions in Emerald Bay were measured in the same fashion but only in June and July for 2010. For May through August 2009, mean surface UV exposure over a four day window was determined from a frequency distribution of UV surface irradiance as measured with a ground based UV radiometer (GUV; Biospherical Instruments, San Diego, CA) deployed at Lake Tahoe (Fig. 3.2c). For June and July 2010, surface UV exposure at Lake Tahoe was measured with a BIC-Logger (Biospherical Instruments, San Diego, CA) for a single four day window of time under sunny conditions. The four day window is used for comparison to the four day in situ incubations described below, a typical (though conservative) duration for larval bass incubation on the nest before swim-up

(Wallas and Simon 2008). UV exposure at each depth was calculated using the Beer-Lambert -k Z law that describes the attenuation of light in water, EZ=E0e d where, E0 is surface UV exposure,

kd is the diffuse attenuation coefficient derived from BIC profiles, and Z is depth. Larval survival as a function of UV exposure was predicted for each depth, location, and month using the calculated UV exposure values and the derived UV exposure-response relationship. Temperature based larval survival was predicted for each depth, location, and month using the field measured temperature values and the derived temperature-response relationship. Finally, temperature and UV based survival estimates were multiplied to produce a single metric for larval survival as a function of depth (Fig. 3.2d).

The second approach also used the experimentally derived UV exposure and temperature response curves but these response curves were used to predict reproductive success (i.e. larval survival) over the range of UV exposure and temperature conditions observed from 4 years of sampling data for 11 nearshore locations (Fig. 3.3). The upper range for UV exposure was based on 95th percentile surface UV exposure from the frequency distribution of four day windows -2 calculated each month for May through August 2009 (i.e. 7.12 kJ*m UV305 nm). The temperature range was based on minimum and maximum temperatures measured from monthly BIC profiles in June and July (2007-2010) at 0.5 meters, except that the lower limit for temperature was constrained to 10° C, the lowest temperature used in the experiment. For comparison, the minimum observed temperature at 0.5 m during our sampling was 8.5° C. We used the ‘mean’ response of bass to UV exposure and temperature (i.e. the product of the UV and temperature response models described above) to predict reproductive success at each UV and temperature combination over the range of possible conditions. To predict invasion potential at multiple nearshore locations we evaluated bass survival at a depth of 0.5 m in June at all 11 sample sites based on the range (i.e. mean ± SEM) of UV exposure and temperature conditions measured at each site over the four year sample period. UV exposure at 0.5 m was calculated for each site from the mean of four day UV surface exposures measured in June 2009 and the mean of diffuse attenuation coefficients from BIC profiles in June each year. Temperature was taken as the mean (± SEM) at 0.5 m from June BIC profiles. The model predictions were validated with in situ incubation experiments. For each incubation experiment, embryos of largemouth bass or redsides were collected from a single

location in Lake Tahoe and hatched in the laboratory. After hatching, yolk-sac larvae (n= 5) were placed in Whirl-Pak bags (Nasco, Fort Atkinson, Wisconsin, USA) filled with 100 mL of 48-μm filtered lake water to exclude most zooplankton. To isolate the effect of UV between incubation sites, the Whirl-Pak bags were either shielded from incident UV in Courtgard (CP Films, Martinsville, Virginia, USA) sleeves or exposed to incident UV in Aclar (Honeywell International, Morristown, New Jersey, USA) sleeves. Courtgard is a long-wave-pass plastic that transmits photosynthetically active radiation (PAR; 95% 400–800 nm in water) but blocks most UV (transmits no UV-B 295–319 nm, and only 9% of UV-A 320–400 nm with a sharp wavelength cutoff and 50% transmittance at 400 nm). Aclar is a long-wave-pass plastic that in water transmits both PAR (100% 400–800 nm) and most UV (98% of UV-B 295–319 nm, 99% UV-A 320–399 nm, with a sharp wavelength cutoff and 50% transmittance at 212 nm). Four replicates of each of the UV shielded and unshielded treatments were deployed at a single depth (≤ 1.1 m) at a given site for four days. After collection, larvae were examined under a dissecting microscope and scored as live if a heartbeat was observed. Largemouth bass were deployed at Tahoe Keys and Sand Harbor in June 2009 and at all other sites in June 2010. Redside minnow larvae were deployed in Taylor Creek July 2009 and at all other sites in July 2008. All procedures involving animals were in accordance with the policies set forth by Miami University’s Institutional Animal Care and Use Committee (IACUC protocol #683).

Results An interaction test for the fixed effects (exposure * species) from the generalized linear mixed effects model indicated that the native redside minnow larvae were significantly more UV tolerant than non-native largemouth bass (Fig.3.4, t693= 9.11, P < 0.0001). Estimated LE99 values indicated that the native minnow was more than six times as UV tolerant as largemouth bass larvae. Temperature also had a substantial effect on largemouth bass reproductive success (Fig. 3.5). At the upper end of the modeled temperature range bass survival was just 5% (29° C), but over the range of 14°-23° C predicted survival exceeded 95%. Predicted larval survival (at depths less than1.2 m) in three nearshore sites varied considerably as a function of UV and temperature across sites and within sites seasonally (Fig. 3.6). Survival in Tahoe Keys, a relatively warm and turbid marina, exceeded 90% at depths > 0.4 m for spring and summer

months, whereas high UV conditions at Sand Harbor, a transparent nearshore embayment, precluded larval bass survival in any month. Reproductive success at Emerald Bay was intermediate to Tahoe Keys and Sand Harbor. Predicted larval bass survival was generally higher in June than in July, at least over the shallowest depths (Fig 3.6a, 3.6b). This finding is consistent with personal observations that the majority of largemouth bass spawning in Lake Tahoe occurs in mid to late June and tails off by mid-July. In situ incubation of largemouth bass at each of these three locations corroborated the predictions for reproductive success (i.e. larval survival). For example, mean larval bass survival after four day incubation at 1.1 m depth in June 2010 at Emerald Bay was 85%, which matched the predicted survival based on mean response from outdoor UV exposure and temperature experiments (Fig. 3.6b). Mean larval survival for four day incubations in UV blocking microcosms was greater than or equal to 95% for all three sites. Consistent with the conceptual invasion-window model, UV and temperature conditions substantially influenced predicted larval bass survival (Fig. 3.7). UV exposure levels in excess of -2 -2 2 kJ*m UV305 nm were predicted to induce complete mortality. For 0.5 kJ*m UV, predicted survival was > 90% over the temperature range 13.4° – 24.1° C. Based on the range (mean± SEM) of UV exposure and temperature conditions during June at 0.5 m depth, 6 of 11 sample sites are susceptible to bass establishment (i.e. predicted larval survival at 0.5 m >10%; Fig. 3.8). If deeper spawning habitat is available additional sites may be susceptible to bass establishment (assuming suitable spawning temperature at greater depth), and sites that are only marginally suitable at 0.5 m depth may provide more optimal spawning habitat at greater depths (e.g. Emerald Bay, Fig 3.4b). In situ incubation of largemouth bass at 3 of the 6 sites suitable for establishment and at 2 of the 5 unsuitable sites corroborated the predictions for reproductive success (Table 3.1). Survival of native redside minnow larvae at all 5 sites was ≥ 90%.

Discussion Fish in the earliest life history stages (e.g. pre-swimming larvae) are non-mobile and thus subject to the conditions in which they are spawned. Consequently, they cannot avoid UV or move out of sub-optimal temperature conditions to reduce UV damage or thermal stress. In this study we tested the UV tolerance of pre-swimming yolk-sac largemouth bass larvae collected

from Lake Tahoe and we used data from the literature detailing the effect of temperature on hatching success of bass embryos to parameterize a statistical model that predicted reproductive success for young bass over the range of UV and temperature conditions in nearshore Lake Tahoe. The model results suggested that larval bass survival (and thus establishment potential) varied seasonally, spatially, and over depth gradients and was constrained by UV and temperature conditions. In situ incubation experiments confirmed model predictions. Non- native bass were significantly more UV sensitive than native redside minnows. In the absence of calculated ‘invasion-windows’ for the native redside minnow, results from in situ experiments suggest support for the expectation that native species overlap broadly with the range of ambient environmental conditions in the Tahoe littoral zone. These findings support our conceptual model (Fig 3.1), which predicts that invasion potential (i.e. reproductive success as larval survival) for relatively UV sensitive warmwater fish larvae should be greatest when UV is low and temperatures are at least moderately warm. Historically, low-UV warm-water conditions would have been limited temporally and spatially in Lake Tahoe. However, sustained disturbance (e.g. eutrophication, stream bank erosion, climate warming) has reduced water clarity and increased water temperature both locally and at the whole-lake scale (Goldman 1988, Byron and Goldman 1989, Coats et al. 2006). In fact, 3 of the 6 samples sites where bass establishment was predicted are extensively modified marina areas, including the only location (site 5) where largemouth bass nesting is known to occur in Lake Tahoe (Kamerath et al. 2008). Other disturbance/stress models (e.g. Davis et al. 2000) are based on an assumption that success for an invading species is greatest when competition for resources is minimal. Such an assumption is rooted in theory that suggests competition intensity is inversely related to the amount of unused resources (Davis et al. 1998), and consequently many studies of invasibility are couched in terms of the availability of unexploited resources. The primary importance of our model is that it shows how abiotic conditions are sufficient to explain invasion potential even in the absence of competitive interactions. Furthermore, this approach facilitates explicit predictions of invasion potential in lakes because it emphasizes specific and quantifiable aspects of the invader and the habitat. In other words, the invasion-window model for warmwater fish successfully integrates species invasiveness (i.e. UV and temperature tolerance of larval bass) and habitat invasibility (i.e. UV and temperature conditions in nearshore

habitat). It is apparent that in Lake Tahoe changing UV and temperature conditions have indeed opened an invasion-window. Warmwater fish have been observed in a variety of habitats. For example, an extensive survey of more than 3000 Minnesota, USA lakes showed that warmwater fish (including largemouth bass) inhabit lakes of various depth, surface area, and productivity, including relatively large oligotrophic lakes (Stefan et al. 1995). Yet, this is the first time that we know of where the roles of both UV and temperature have been considered with respect to warmwater fish distribution. A single large lake like Tahoe, with its heterogeneity of nearshore habitats, offers an advantage for testing ideas about invasion resistance. Strong swimming adult fish potentially have access to all habitats, but establishment of local reproducing populations reflects local conditions. The “invasion-window” is held partly (locally) open over many years, and we can analyze regulatory factors as they operate concurrently in different habitats, on a common species pool. Thus, our research could provide important insights into the potential for warmwater species invasion in lakes, especially for coldwater lakes characterized by exceptional clarity but threatened by both increasing temperature and declining transparency.

Dissolved organic carbon (DOC), the ozone of the underwater world (Williamson and Rose 2010) is the primary regulator of underwater UV levels (Morris et al. 1995). Many lakes in Europe and North America have shown increasing trends in dissolved organic carbon (DOC) concentrations (Roulet and Moore 2006, Zhang et al. 2010). There is also some evidence that recovery of lakes from acidification can lead to a rebound in DOC concentration (Evans et al. 2006, Monteith et al. 2007). These trends of increasing DOC concentrations are likely to reduce UV transparency in such systems, and potentially create refugia for invasive species. On the other hand, decreasing DOC concentrations are still reported in some locations (Striegl et al. 2005, Monteith et al. 2007) , with likely corresponding increases in UV transparency. In very low DOC lakes in alpine regions (Laurion et al. 2000, Sommaruga and Augustin 2006), phytoplankton play a more important role in regulating underwater UV exposure than DOC. This suggests that anthropogenic eutrophication could also significantly reduce UV penetration in some lakes. For example, in some higher elevation (1800-3500m) transparent lakes like Tahoe, trends of increased algal growth and reduced water clarity stemming from increased nitrogen deposition have been observed (Jassby et al. 1994, Sickman et al. 2003).

Climate change will likely exacerbate the direct and indirect effects of UV and temperature on aquatic invasive species. For example, the magnitude of changes in DOC concentration vary with temperature and depend largely on drought, precipitation, and the timing of ice-out (Pace and Cole 2002). Thus, climate driven changes in temperature and precipitation regimes are likely to have strong direct effects on the major factor (DOC) controlling UV attenuation in aquatic systems. Climate change is also likely to alter water temperature and duration of ice cover, factors that have already been linked to changes in the likelihood of potentially invasive species establishing in novel habitats (Rahel and Olden 2008). Each of these factors could also indirectly influence UV exposure for aquatic invaders. Rising water temperatures, for instance, could increase the depth refuge for UV sensitive species by allowing them to reproduce at greater depths. Given the strong tradeoffs in reproductive success induced by temperature and UV (Huff et al. 2004) this could substantially increase the likelihood of establishment for some species. It may also be possible that earlier ice out on lakes could accelerate spawning phenologies, allowing some species to take advantage of reduced underwater UV conditions characteristic of the early spring and related to seasonal variation in both incident UV and phytoplankton production (Norsang et al. 2009, Williamson et al. 2007). The invasion-window model could even have relevance for some of the most unique and biologically significant lakes in the world, including Lake Baikal which like Tahoe is characterized by its exceptional water clarity. Baikal supports more species than any other lake in the world and more than half of its animal species are endemic (Martin 1994, Timoshkin 1995). However, the integrity of this great lake is threatened by a number of factors including changes in temperature and potential loss of water clarity. For example, summer surface water temperatures have warmed at an average rate of 0.38° C*decade-1 over the last 60 years and in the last quarter century algal biomass has tripled (Hampton et al. 2008). Although there is not yet any documented decline in transparency (at least to visible light as measured by Secchi depth) climate change will increase thawing of permafrost and could thereby accelerate cultural eutrophication by exacerbating nutrient and sediment inputs to the lake (Moore et al. 2009). Also of note is the significant increase in discharge for rivers in the Siberian Arctic and predictions of as much as 70% greater discharge than at present based on IPCC global surface air temperature projections (Peterson et al. 2002). The substantial increase in river discharge, coupled with

thawing permafrost, could dramatically reduce water clarity, especially in nearshore locations at the mouth of large tributaries. One warmwater fish, the common carp (Cyprinus carpio), is already widely distributed in Lake Baikal after introduction in the 1940s (Kozhova and Silhow 1998). Other invasive species have had significant negative impacts in the nearshore zone, including American waterweed (Elodea canadensis), which has reduced biodiversity of plant and invertebrate bottom fauna where it occurs (Kravtsova et al. 2010). The role of UV and temperature in controlling the distribution of these aquatic invaders has not been investigated. Another relevant concern is the potential for drastic shifts from clear to alternate turbid states in shallow nearshore ecosystems. The sudden loss of transparency in lakes subjected to eutrophication is a well documented phenomenon (Scheffer et al. 1993, Carpenter et al. 1999, Jeppesen 1999). Our model suggests that a sudden state shift from clear to turbid conditions could facilitate the introduction of potentially invasive species in nearshore habitat. In Lake Tahoe the nearshore habitat where warmwater fish have established also supports a dense stand of an invasive emergent macrophyte, water milfoil (Myriophyllum spicatum; Anderson et al. 2005). Although submerged macrophytes in lakes generally reduce turbidity and are most characteristic of clear-water states (Sheffer et al. 1993) large densities of decaying milfoil can act as pumps to move phosphorous into the water column (Landers 1982, Carpenter 1980) which could further decrease water clarity. Light also decays exponentially under macrophyte canopies (Westlake 1964) and thus may provide additional protection from damaging UV exposure. On the other hand, macrophyte stands also generally reduce water temperature, by as much as 10° C m-1 in one study (Dale and Gillespie 1977). Whether submerged macrophtyes will generally support or hinder warmwater fish invasion is an interesting question that needs to be tested and one that has relevance to the potential effects of alternate stable states on biological invasion in lakes. It is uncertain whether introduced species themselves could induce further declines in water transparency and thus affect invasion potential for other non-native species. According to the classic ‘trophic cascade hypothesis,’ increasing the biomass of piscivorous fish in lakes reduces the biomass of planktivorous fish, increases zooplankton biomass, and reduces phytoplankton biomass (Carpenter and Kitchell 1993). For species like largemouth bass, a major predator of planktivorous fishes in temperate lakes (Carlander 1977) the trophic cascade

hypothesis suggests a negative feedback mechanism that could potentially limit widespread bass establishment. In fact, stocking of piscivorous fish has been used as a successful means of biomanipulation to return turbid lakes to a permanent clear state (Shapiro and Wright 1984). However, the cascade effect can depend heavily on the nature of the food-web in lakes and is actually only weakly expressed or not at all evident in some systems (Hambright et al. 1991, Stein et al. 1995). In Lake Tahoe, the plankton population is relatively depauperate, especially in terms of important cladoceran grazers (Richards et al. 1975). Furthermore, the likely fish prey for largemouth bass in nearshore Tahoe are Lahontan redsides (Richardsonius egregius) and speckled dace (Rhinicthys osculus robustus), neither of which are heavily dependent on plankton for food and in fact rely largely on flying insects and benthic organisms (Miller 1951, Evans 1969). Thus the strength of the trophic cascade in Tahoe, and its subsequent effect on water clarity could be minimal. Alternatively, other warmwater fish species could have more pronounced effects on water clarity. For example, bluegill (Lepomis macrochirus) are established in a limited number of nearshore sites in Lake Tahoe. Previous research suggests that bluegill distribution in the lake may be related to UV (Tucker et al. 2010). Bluegill are important planktivores in lakes as both adults and fry (Werner 1967, Mittlebach 1981) and large bluegill populations have been linked to reduced water clarity in some systems (Hanson and Butler 1994). Thus, bluegill might actually exacerbate water clarity loss. Whether bass and/or bluegill establishment in Lake Tahoe is likely to facilitate their invasion or the establishment of other non-native UV sensitive species (‘invasional meltdown’ sensu Simberloff and Von Holle 1999) is yet another interesting question that could be explored in the context of this invasion- window model. The invasion-window model of warmwater species invasion could be a powerful tool for understanding and predicting invasion potential in lakes where transparency and temperature are changing and for exploring some unanswered questions in invasion biology. We recognize the limitations of generalizing from just a few species in a single lake, however we have carefully tested the role of UV in a system that offered a wide range of UV and temperature conditions and within which non-native species were established locally but not pervasively within a native fish community. It is evident from model predictions and in situ incubation experiments that UV and temperature impose strong constraints on the reproductive success of non-native largemouth

bass. We note that many general ecological models that opened new frontiers are developed from and stimulated by more circumscribed studies of just one or two species. For example, a literature search for articles published in the past 2 years with the word “invasive” in the title in five leading journals that publish on invasion ecology (Biological Invasions, Diversity and Distributions, Biological Conservation, Oecologia, Ecology) revealed that of the 122 experimental or observational papers found, only 12 studied more than 2 species, and most (94) studied only a single species. We hope that the invasion-window model based on UV and temperature encourages scientists and managers to more seriously consider the role of these two abiotic factors in regulating species invasion.

Acknowledgements

We thank Geoff Schladow and the staff of the Tahoe Environmental Research Center for their assistance. Robbyn Abbitt and Neil Winn created the GIS map of Lake Tahoe. Michael Hughes of the Statistical Consulting Center of Miami University provided data analysis assistance. Michael Cohen, Ian Lizzadro-McPherson, Amanda Gevertz, Annie Bowling, Jeremy Mack, Kevin Rose, and Sandra Connelly provided field assistance. Erin Overholt conducted the literature search. This work was supported in part by Miami University’s Field Workshop Program.

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Table 3.1. Percent Survival of largemouth bass and Lahontan redside larvae after in situ incubation in UV transparent microcosms for each of 5 nearshore sites. Corresponding UV exposure and temperature conditions are indicated for the duration of the 4 day incubations. Largemouth bass were deployed in June 2009 and 2010. Lahontan redside were deployed in July 2008 and 2009. 100 100 100 100 95 ± 5 90 ± 10 ± 90 (± SEM) Lahontan Redside Redside Lahontan Temp (°C) survival Mean 2 - kJ*m 305-nm UV (± SEM) (°C) Temp Mean survival 2 - kJ*m 305-nm 6- Taylor Creek 6- Taylor 0.5 13.9 85 ± 5 1.2 23.9 Largemouth Bass Bass Largemouth UV 1- Crystal Bay Harbor 2- Sand 2.2 4.0 12.9 17.2 0 0 2.1 3.6 20.7 21.2 5- Tahoe Keys 5- Tahoe 7- Emerald Bay 0.2 0.3 21.4 13.4 94 ± 6 ± 10 85 0.0 1.8 23.0 21.4

Invasion Windows

High Survival Historical Environmental Conditions Establishment 3 2 1 UV Disturbance

Survival Low Reproduction Low High Temperature

Figure 3.1. Conceptual diagram showing how natural or human disturbance can lead to progressive expansion of historical physical environmental conditions to facilitate invasion. This expansion may be (1) too slight to enable any invasion, (2) enough to permit larval or adult survival, creating a survival invasion-window, or (3) enough to permit establishment and reproduction, creating an establishment invasion-window.

1.0 a) 1.0 b) ortion p Survival Pro Survival Survival Proportion Survival

0.0 0.0 Cumulative UV Dose (kJ*m-2) Temperature ( °C ) Mean Survival Proportion c) d) Depth (m) Frequency

Cumulative 4 Day UV Exposure (kJ*m-2)

Figure 3.2. Conceptual diagram of the method used to predict invasion potential (reproductive success as larval survival) in nearshore sites. Predicted response for largemouth bass is indicated by the solid lines in panels a, b, and d. Predicted response for native redside minnow is also indicated for comparison (dashed line) although only the UV response was actually modeled for the redside. Step 1) Determine bass response to UV and temperature (species invasiveness, Panels a and b). 2) Determine temperature and UV conditions in nearshore sites (habitat invasibility). For example, panel c shows a frequency distribution of cumulative surface UV exposure levels used to estimate UV exposure conditions in each site. 3) Combine species invasiveness with habitat invasibility into a single metric for bass survival as a function of depth (Panel d, for a moderately UV transparent site).

Figure 3.3. Map of Lake Tahoe indicating the location of sample sites where invasion potential was assessed. Site numbers correspond to those plotted in Figure 3.9. Sites where fish larvae were incubated in situ are indicated with a symbol. Bathymetric contour interval is 100 ft.

0.1 Survival Proportion

RS LMB 0.01 024681012 -2 UV (kJ*m )

Figure 3.4. Effect of UV exposure (305 nm kJ*m-2) on survival of native Lahontan redside minnow (RS) and non-native largemouth bass (LMB) with 95% upper and lower confidence limits. Note: Survival proportion is presented on log scale.

0.8

0.6

0.4 Survival Proportion Survival

0.2

0 8 1012141618202224262830 Temperature (° C)

Figure 3.5. Effect of temperature on hatching success of largemouth bass eggs (data from Kelly 1968). A local polynomial regression model (gray dashed line) was fit to the data.

a) REPRODUCTIVE SUCCESS 0 0.2 0.4 0.6 0.8 1 0

0.2

0.4 (m)

0.6 DEPTH 0.8 MAY JUN JUL 1 AUG IN SITU (0 %) 1.2 b) REPRODUCTIVE SUCCESS 0 0.2 0.4 0.6 0.8 1 0

0.2

0.4 (m)

0.6 DEPTH 0.8

1 JUN JUL IN SITU (6 %) 1.2 c) REPRODUCTIVE SUCCESS 0 0.2 0.4 0.6 0.8 1 0

0.2

0.4 (m)

0.6 DEPTH 0.8 MAY JUN 1 JUL AUG IN SITU (88%) 1.2

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Figure 3.6. Predicted reproductive success (i.e. larval survival) based on UV and temperature for largemouth bass at 3 nearshore locations, a) Tahoe Keys, b) Emerald Bay, and c) Sand Harbor in spring and summer. Reproductive success predictions agree with observed larval survival following in situ incubation experiments in June at each site (indicated by red X). Percent of surface UV305 nm irradiance at in situ incubation depth is indicated in parentheses.

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) -2 UV (kJ*m

Temperature (°C)

Figure 3.7. Contour plot of predicted reproductive success (z). Predicted reproductive success is reported here as the product of UV * T effects on larval bass, based on mean responses (see Fig. 3.5 and Fig. 3.4). The contour plot is analogous to the establishment-window in our conceptual model, Fig. 3.2.

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1 4 10 3 ) -2 4 3

UV (kJ*m UV 7 2

8 6 1 11 5 9 0 8 12162024 Temperature (° C)

Figure 3.8. Plot of mean (± SEM) UV and temperature conditions in June at 0.5 m depth for 11 nearshore sites from 4 years of sampling data (except sites 2 and 9 n=3, site 6 n=2). The establishment window (i.e. where UV and temperature conditions suggest bass survival ≥ 10%, based on Fig. 3.8) is plotted in the background. Observed survival of larval bass agrees with predicted survival from the invasion-window model based on UV and temperature conditions during in situ incubation (Table 3.1).

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Chapter 4

A UV attainment threshold for the prevention of warmwater aquatic invasive species

To be submitted: Ecological Applications

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Abstract

Biologists and managers continue to search for effective tools to predict what habitats are susceptible to biological invasion and to prevent the establishment and spread of harmful invasive species. Abiotic factors are important regulators of biological invasion and a recent study in Lake Tahoe (CA/NV, USA) showed that ultraviolet radiation (UV) can mediate warmwater fish invasion. In this study we highlight field and laboratory experiments that indicate strong species related differences in UV-induced stress between native and non-native fish species inhabiting the nearshore environment in Lake Tahoe. We use this differential UV sensitivity to develop a UV Attainment Threshold (UVAT) that if realized could be used to reduce susceptibility to warmwater fish invasion in nearshore Lake Tahoe. The UVAT is a target value for water transparency based on, 1) incident solar UV exposure levels during peak spawning season, and 2) experimentally derived UV exposure levels lethal to larval warmwater fish. We suggest that this value can be easily measured, monitored, and targeted to reduce the establishment and spread of invasive fish species. We also discuss how the UVAT increases the cost effectiveness of invasive species management. This approach could be relevant for any aquatic invasive species that is UV sensitive and constrained to relatively shallow water environments by, e.g., requirements for warmer spawning temperatures. Because many aquatic systems are threatened by both invasive species and declining water transparency we suggest that this research could have important applications in many systems particularly in colder more UV transparent waters found in regions such as western North America.

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Introduction

Invasion by exotic species is a major concern in freshwater aquatic ecosystems (Richter et al. 1997, Dudgeon 2006), and poses a particularly grave threat to the persistence of native fish populations (Lassuy 1995). Yet considerable uncertainty remains concerning the most effective ways to prevent and/or manage biological invasion in inland waters (Enserink 1999, Williamson 1999). A central unresolved question is: what controls the suitability of habitats for invasion by exotic species? Traditionally the physical and biological characteristics of habitats that control the potential for exotic species to invade and establish have been viewed as components of an “invasion-window” (Johnstone 1986) or “niche opportunity” (Shea and Chesson 2002) that can be altered by disturbance. In freshwater ecosystems most natural and human disturbances that might open an invasion-window will also generate changes in water transparency. For example, transparency to visible light decreases with cultural eutrophication (Edmondson 1991, Seehausen et al. 1997) and with the introduction of planktivorous or predatory fish (Kaufman 1992, Mazumder et al. 1990). A recent report from Lake Tahoe, a large sub-alpine lake in western USA, demonstrated that water transparency to ultraviolet radiation (UV) can control the establishment of non-indigenous warmwater fish (Tucker et al. 2010). Invasive warmwater larval fish incubated in transparent high-UV sites experienced high mortality, while those incubated in turbid low-UV sites survived well. This finding is consistent with the emerging consensus that abiotic factors are important regulators of biological invasion (Holway et al. 2002, Menke and Holway 2006, Dethier and Hacker 2005, Gerhardt and Collinge 2007). Here we use Lake Tahoe as a case study to show how aquatic invasive species prevention and management strategies that focus on regulating water transparency, and thus UV exposure, could help to stem biological invasion in aquatic systems.

Experiments in eastern USA lakes have demonstrated that high UV transparency reduces the reproductive success of warmwater fish in shallow waters (Williamson et al. 1997, Huff et al. 2004, Olson et al. 2006). Lake Tahoe is much more transparent than these lakes. However, reductions in nearshore UV transparency may have provided a refuge for warmwater fish to successfully nest and to establish self-sustaining populations in Lake Tahoe. In fact, Lake Tahoe has experienced a gradual loss of water clarity in the last four decades, with a decline in average

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annual Secchi transparency from 31 m in 1968 to 21 m by 1998 (Jassby et al. 1999). During this period a number of non-indigenous warmwater fish species (including bluegill, Lepomis macrochirus, and largemouth bass, Micropterus salmoides) established populations in a limited number of nearshore locations (Reuter and Miller 2000). Data from nearhore-to-offshore UV profiling transects in Lake Tahoe demonstrate that shallow environments nearshore to some of the major inflows are far less UV transparent than offshore and that patterns of UV transparency change from month to month (Rose et al. 2009). This is important because during summer months shallow nearshore habitat is necessary to provide both the warm temperatures and low UV conditions that permit spawning by non-indigenous species such as largemouth bass (Carlander 1977). In present day Lake Tahoe, native minnow species and introduced warmwater fish both inhabit the nearshore environment. However, the only well established non-indigenous fish populations are limited to sites in the southern end of the lake that are characterized by extensive development and that are in close proximity to some of the lake’s largest tributaries (Kamerath et al. 2008). Water transparency at these sites tends to be lower than elsewhere in the lake and may explain the suitability of such sites for the establishment of non-indigenous fish populations (Tucker et al. 2010). Native minnows occur widely and in nearshore habitats with high levels of UV.

In this study we highlight field and laboratory experiments that indicate strong species related differences in UV-induced stress for fish inhabiting Lake Tahoe. We use this differential UV sensitivity to develop a “UV Attainment Threshold” that if realized could reduce susceptibility to exotic centrarchid fish species establishment in nearshore Lake Tahoe. We present the results from experiments comparing the UV tolerance of native minnow (Lahontan redside, Richardsonius egregius) larvae with that of non-native warmwater fish (bluegill and largemouth bass) larvae (Fig. 4.1). We also present data from field surveys of UV and temperature in nearshore Lake Tahoe. From these data we develop UV attainment thresholds (UVAT) for 11 nearshore locations in Lake Tahoe. The UVAT is a target value for UV transparency based on, 1) surface UV exposure during peak spawning season, and 2) experimentally derived UV exposure levels lethal to larval warmwater fish. For our purposes we present the UVAT as the % of 305- nm surface UV exposure that must penetrate to any given depth in order to prevent largemouth bass reproduction within a site. The UVAT emphasizes

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largemouth bass because they were the more UV tolerant of the two warmwater fish tested. Thus, water clarity improvements that prevent bass survival will also likely prevent bluegill reproduction. We suggest that this value can be easily measured and monitored with a profiling UV radiometer or modeled from water samples analyzed for transparency in the lab with a spectrophotometer, and used to manage nearshore waters in an effort to minimize invasion by warmwater fish species. This approach could be relevant for any aquatic invasive species that is UV sensitive and constrained to relatively shallow water environments by, e.g., requirements for warmer spawning temperatures. Because many aquatic systems are threatened by both invasive species and declining water transparency we suggest that this research could have important applications in many systems.

Methods

UV and temperature profiling

UV and temperature profiles at each of the 11 nearshore sites were taken in June and July 2007, 2008 and 2010, and monthly (except September) from May through October 2009 with a BIC profiling UV-PAR radiometer (Biospherical Instruments, San Diego, CA, USA). The BIC radiometer quantifies incident solar irradiance at three different UV wavelengths (305, 320, and 380 nm) as well as visible wavelengths of photosynthetically active radiation (PAR, 400–700 nm) and measures temperature continuously from surface to maximum depth during sampling

profiles. Diffuse attenuation coefficients (kd) were estimated from the slope of the line derived

by plotting the log of UV intensity at 305 nm versus depth from BIC profiles. Mean kd values for June from 2007-2010 were used to calculate UV exposure at 1m depth as a percent of surface irradiance, except for Sand Harbor and Meeks Bay (2008-2010) and Taylor Creek (2009-2010). UV exposure at 1m depth as a percent of surface irradiance was calculated by modifying the -kdZ Beer-Lambert equation that describes the attenuation of light in water, EZ=E0e where, E0 is surface UV exposure, kd is the diffuse attenuation coefficient derived from BIC profiles, and Z is depth. Nearshore sample sites included: Tahoe Keys East, Crystal Bay, Sand Harbor, Star Harbor, Cave Rock, Emerald Bay and Emerald Bay at Eagle Falls Creek, Sunnyside Marina, Round Hill Pines, Meeks Bay, and Taylor Creek (Fig. 2). Sites were selected to represent the

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wide range in UV transparency and temperature conditions characteristic of nearshore habitats in Lake Tahoe.

UV exposure experiments: Establishing larval UV tolerance For largemouth bass, bluegill, and Lahontan redside larvae, exposure-response relationships of UV-induced mortality were established from outdoor exposure experiments with natural solar radiation in a temperature controlled environment. Yolk sac larvae were selected for these experiments because their high transparency and lack of mobility are likely to make them the life history stage that is most vulnerable to UV. Here we present methods for largemouth bass and redside minnow experiments, from which the UVAT values were determined. Methods for larval bluegill were similar. Largemouth bass were collected as eggs from a single nest in the Tahoe Keys on June 20, 2009 and hatched in the lab on or before June 23, 2009. On the evening of June 26, 5 yolk-sac larvae were added across treatments to each of 12 treatment dishes (1750ml Pyrex crystallizing dishes filled with 1.5 L of 48 μm filtered lake water). Larvae were also added to each of 3 control dishes (1200 ml plastic bowls filled with 48 μm filtered water). Control dishes were maintained in a temperature controlled environment at 18° C under low artificial light conditions. Neutral density stainless steel mesh screens (McMaster Carr, Robbinsville, NJ, USA) were cut to size and placed over the dishes on the morning of June 27 to provide a range of UV exposure levels (100%, 78%, 57%, and 45% of ambient solar UV). Treatment dishes were then placed in an outdoor water bath and larval survival was monitored every 2 hours for 34 hours. Dishes were removed from the bath for no longer than 2 minutes and with the aid of a dissecting microscope, larvae were deemed dead or alive by checking for the heartbeat. Ambient UV exposure for the duration of the experiment was recorded with a BIC-logging radiometer (Biospherical Instruments, San Diego, California, USA). Lahontan redside were collected as eggs from Sunnyside Marina on July 15, 2009 and hatched in the lab on July 19 and 20. On the morning of July 22, 10 yolk-sac larvae were added across treatments to each of 12 treatment dishes (250 ml Pyrex crystallizing dishes filled with 250 mL of 48 μm filtered lake water). Larvae were also added to each of 3 control dishes. Control dishes were placed in the water bath with treatment dishes but completely covered in

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courtgard (CP Films, Martinsville, VA USA). Courtgard is a long-wave pass plastic that transmits no UV-B (295-319 nm; transmits 95% PAR, 400–800 nm and only 9% of UV-A 320– 400 nm in water with a sharp wavelength cutoff and 50% transmittance at 400 nm). Neutral density mesh screens were placed over the treatment dishes to provide a range of UV exposure levels (as above for largemouth bass). Treatment dishes were then placed in an outdoor water bath and mortality was recorded every 2 hours for approximately 150 hours. Mortality was recorded as the cessation of a heartbeat observed under a dissecting microscope. Ambient UV exposure was recorded with the BIC-logger. A generalized linear mixed effects model was used to determine the effect of UV exposure on the cumulative proportion dead for each larval species. Mixed effects models allow for fixed and random effects (Verbeke and Molenberghs 1997). The fixed effects included species type and exposure level. Because of the random effect of replicate differences the model was fit using PROC GLIMMIX (SAS v. 9.2). The model accommodates dose/species interaction effects. The indicator variable coding treats redside minnow as a “baseline” species. The model specification was:

where:

= empirical odds of mortality

Exp = cumulative UV Exp

IBG = 1 for bluegill, 0 otherwise

ILMB = 1 for largemouth bass, 0 otherwise

b0 = random intercept effect due to chamber

ε = random error

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Determining the UV attainment threshold The UVAT was determined as follows: (1) Exposure-response relationships of UV-induced mortality were established from outdoor exposure experiments and used to determine the effective exposure of 305-nm UV to achieve the target amount of bass mortality. Based on a typical approach for determining efficacy of a pesticide (Ritz et al. 2006) we selected a UV-exposure level that caused a high amount of mortality (99%) in bass, but a low amount of mortality in native species (<1%). (2) Cumulative surface 305-nm UV exposure was determined for 4 day periods from a frequency distribution of surface exposure for the month of June 2009, as measured with a ground based GUV- radiometer (Biospherical Instruments, San Diego California, USA). Surface UV exposure was calculated for four day windows because four days represents a typical, though conservative (i.e., short), incubation period for yolk-sac largemouth bass larvae in the nest before they reach the swim-up stage. The month of June was selected because June represents the peak spawning season for largemouth bass in Lake Tahoe. (3) The UVAT was then calculated simply as the percent of surface 305-nm UV exposure that must penetrate to any given depth to prevent larval bass survival (i.e., result in mortality of 99% of the population):

UVAT = 100*(LE99 / E0) -2 where: LE99 = effective UV exposure selected to target bass mortality (kJ*m ) -2 E0 = cumulative 4 day 305-nm UV exposure at surface (kJ*m )

In situ incubation experiments: Corroborating the UVAT

In situ incubations were conducted to test the validity of the UVAT. Larval yolk sac largemouth bass used in incubations were collected from a single nest at approximately 1 m depth in the Tahoe Keys. Larvae (n= 5) were placed in Whirl-Pak bags (Nasco, Fort Atkinson,

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Wisconsin, USA) filled with 100 mL of 48-μm filtered lake water to exclude zooplankton. To isolate the effect of UV between incubation sites, the Whirl-Pak bags were either shielded from incident UV in courtgard sleeves or exposed to incident UV in Aclar (Honeywell International, Morristown, New Jersey, USA) sleeves. Aclar is a long-wave-pass plastic that in water transmits both PAR (100% 400–800 nm) and most UV (98% of UV-B 295–319 nm, 99% UV-A 320–399 nm, with a sharp wavelength cutoff and 50% transmittance at 212 nm). Four replicates of each of the UV shielded and unshielded treatments were deployed at one meter depth at a given site for four days. After collection, larvae were examined under a dissecting microscope and scored as live if a heartbeat was observed. All procedures involving animals were in accordance with the policies set forth by Miami University’s Institutional Animal Care and Use Committee (IACUC protocol #683).

Results

UV transparency of nearshore Lake Tahoe varies considerably (Table 4.1). For the most transparent sites, an average of more than 70% of 305-nm UV present at the surface penetrated to a depth of 1 m. In the least transparent sites all of the 305-nm UV was attenuated by 1 m. There is also some variability in incident surface exposure due to changes in cloud cover and atmospheric aerosols throughout the month-long study period (Fig 4.3). For the month of June 2009, incident surface 305-nm UV exposure for all possible consecutive four day periods ranged from 2.7 to 6.2 kJ* m-2. The majority of these 4 day windows of time indicated surface exposure somewhere between 4 and 6 kJ*m-2. The median surface exposure value of 4.99 kJ*m-2 was used to calculate the UVAT.

The disparity in UV tolerance of native (Lahontan redside) versus non-native (bluegill and largemouth bass) larvae was striking (Fig. 4.4). The native minnow was more than six times as UV tolerant as largemouth bass and almost 10 times more UV tolerant than bluegill. Approximately 2 kJ*m-2 305-nm UV exposure was lethal to at least 99% of non-native fish larvae with little or no affect on the native minnows. Thus, the UVAT value for the prevention of largemouth bass was estimated to be 40% (UVAT= 100* [1.99 kJ*m-2/ 4.99 kJ*m-2]). 5 of 11 sample sites attained the UVAT (i.e., > 40% of surface 305-nm UV exposure was still present at

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1 m depth in June, a typical spawning depth and time for largemouth bass in Lake Tahoe), suggesting low probability of successful warmwater fish reproduction. UV transparency did not meet the UVAT standard at the remaining 6 sample sites (Table 4.2). In situ incubation experiments in UV transparent microcosms support predictions of attainment/non-attainment (i.e. larval mortality/survival) based on the UVAT (Table 4.2). Mean survival of larval bass in UV blocking microcosms was ≥ 95% for all sites.

Discussion Our results demonstrate that the larval stage of a native fish is significantly more UV tolerant than representative non- native warmwater fish larvae. We leveraged this strong species related difference to develop a UV transparency threshold (the UVAT) that could prevent larval warmwater fish survival in Lake Tahoe. In situ incubation experiments across a gradient of UV transparency corroborate the validity of the UVAT (Table 4.2). Our approach shows that UV transparency of nearshore sites significantly impacts the survival of warmwater fish larvae and may influence whether or not these potentially invasive fish species are able to establish throughout nearshore Lake Tahoe. By quantifying the effect of UV exposure on the earliest life history stage of warmwater fish and measuring levels of this important abiotic factor in nearshore Lake Tahoe our research provides critical new insights that will allow managers to more cost effectively control and/or prevent warmwater fish invasion. In particular this approach allows lake managers to identify ‘at risk’ sites where systematic monitoring and early detection efforts for non-indigenous species can be targeted. In this example ‘at risk’ sites represent locations in the nearshore environment where warmer water temperatures and reduced UV transparency provide a refuge for reproduction of warmwater fish. From our analysis, ‘at risk’ sites are those sites where, for the month of June, water temperature is greater than the lower limit for bass spawning and less than 40% of surface 305-nm UV is present at a depth of 1 m. Our analysis was based on observations that suggested June spawning at 1 m depth is typical for largemouth bass in Lake Tahoe. We used a very conservative lower thermal limit of 12.7 ° C for bass spawning (Wallus and Simon 2008). Water temperature at 1 m depth in June exceeded 12.7° C for 9 of 11 sample sites (Table 4.3). In calculating the UVAT of

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40% we assumed normal UV conditions (i.e. median June 305-nm surface exposure, see Fig 4.3). Based on these criteria six of our eleven sample sites are ‘at risk’ of bass establishment, assuming the presence of reproductive adults. Emerald Bay at Eagle Falls Creek was in non- attainment based on UV exposure but low water temperature in June likely precludes successful bass spawning at this site. Regular monitoring of these ‘at risk’ sites increases the likelihood of early detection and the potential for successful implementation of control measures. This approach also increases cost effectiveness of management efforts because it 1) focuses efforts in space (i.e. on ‘at risk’ sites) and 2) focuses management efforts in time, by assessing the possibility of larval survival based on temperature and UV during the warmwater fish spawning season. Effective monitoring programs are still likely to require substantial effort to succeed in stemming the spread of invasive species. For example, a successful monitoring program for the presence of warmwater fish nests in Lake Tahoe would at the very least include plans to sample during the months of possible warmwater fish spawning and at depths where nest establishment is most likely. Our nearshore survey data suggest that in most years water temperature in Lake Tahoe is suitable for spawning in ‘at risk’ sites beginning in June. In some sites (e.g. Tahoe Keys) water temperature exceeds 12.7° C beginning in May (Table 4.3). Therefore, a comprehensive monitoring program for Lake Tahoe would need to consider site specific temperature data and initiate at least bi-weekly surveys to ensure nest detection. Furthermore, even though we have not observed bass spawning after early July in Lake Tahoe, water temperatures are still suitable for spawning well into summer, and even into early fall in some locations (Table 4.3), which might warrant continued monitoring later into the growing season. On the other hand, largemouth bass are not known to have especially protracted spawning seasons (Wallus and Simon 2008) and age-0 bass survival is known to depend largely on size at overwintering while recruitment to age 1 favors fish spawned earlier in the season (Miranda and Hubbard 1994, Ludsin and DeVries 1997, Garvey et al. 1998). Also, despite observations suggesting that the majority of largemouth bass spawning occurs at less than 1 meter in Lake Tahoe (an observation consistent with descriptions from the literature that report typical nest depths of 15 cm to 2 m, Carlander 1977), the most thorough monitoring program may require surveys up to 10 meters in depth. Nevertheless, the identification of ‘at risk’ sites based on the

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UVAT significantly reduces the workload for managing invasions by identifying target sites and constraining monitoring times. Coupling the UVAT approach with other more traditional metrics could even further constrain monitoring efforts and costs. For example, in lakes where managers already have data on other metrics for bass success (e.g. substrate availability, food resources, and adult presence or absence), ‘priority at risk’ sites could be identified that, based on multiple metrics (including UV), are most likely to support reproduction of invasive warmwater species. At its best this approach might eliminate monitoring costs altogether, as sustained high levels of UV throughout a lake could effectively prevent invasion. Thus managing water clarity to maximize underwater UV levels may be an important end goal in itself. In Lake Tahoe transparency to visible light has declined rather dramatically (Jassby et al. 1999). The decline has traditionally been attributed to increases in both organic (i.e., phytoplankton and detritus) and inorganic (i.e., terrestrial sediment) particulate matter (Swift et al. 2006) resulting largely from human impacts in and around the basin related to eutrophication (Goldman 1988) and stream bank erosion (Byron and Goldman 1989). However, dissolved organic carbon (DOC) is often the most important regulator of UV transparency in aquatic systems (Morris et al. 1995, Rae et al. 2001) and has been called the ‘ozone of the underwater world’ (Williamson and Rose 2010). DOC may be especially important in nearshore habitats where fish spawning occurs, since DOC inputs are likely to be concentrated in those areas. For example, in Lake Tahoe, chromophoric dissolved organic matter absorption coefficients for stream water were 10 times higher than offshore values, which is indicative of potentially much higher concentrations of DOC in the nearshore environment than in offshore sites where long term transparency monitoring has been conducted (Swift 2004). In fact, water quality analysis of nearshore sites in Tahoe shows that UV transparency in the nearshore is strongly dependent on DOC concentration though the best model for predicting UV transparency included both DOC and chlorophyll (Tucker et al. 2010). The influence of inorganic particulates on UV transparency has not been assessed in Lake Tahoe. Efforts to quantify the relative importance of each of these factors in controlling UV transparency (especially in ‘at risk’ sites) and adapting management approaches accordingly could prove to be an effective approach for preventing the establishment of aquatic invaders. For example, whether or not best management practices (BMPs) and other water clarity improvement measures move sites toward UV attainment thresholds could be an additional

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criterion for evaluating the effectiveness of BMPs. In Lake Tahoe, BMPs and water clarity improvements have focused largely on reducing levels of organic and inorganic particulate matter (Schuster and Grismer 2004). Where these efforts are effective in improving water clarity they are also likely increasing underwater UV exposure and therefore moving sites toward a UVAT standard that could prevent warmwater fish invasion. However, given the important role that DOC plays in regulating UV transparency in nearshore Lake Tahoe (Tucker et al. 2010) management practices that are tailored to increase underwater UV levels in particular (e.g. by reducing [DOC]) may be a more cost-effective means of both increasing water clarity and preventing warmwater fish spread.

The major advantage of the approach that we have outlined here is that it quantifies the effect of a specific abiotic factor on the life history stage that determines whether or not potential invaders will establish, a key phase in the invasion process (Williamson and Fitter 1996; Moyle and Light 1996). This kind of approach has been advocated as an especially effective means for predicting invasions because, unlike correlative approaches that use species distribution data to infer species range limits, this approach explicitly considers physiological constraints on organisms and effectively maps those fundamental niche limits onto the landscape (Kearney and Porter 2009). Furthermore this methodology fits nicely into a framework for assessing vulnerability of inland waters based on explicit assessment of site suitability, and is likely to be especially useful for forecasting and/or preventing secondary spread of invaders in aquatic systems where isolated populations of invaders have not yet spread to all sites with suitable habitat (Vander Zanden and Olden 2008). Our method is not unlike the threshold approach that others have used to develop maps of invasion risk for zebra mussels (Neary and Leach 1992, Whittier et al. 2008), except that we have identified a novel factor (UV exposure as a function of water transparency) that is relevant for at least warmwater fish establishment in more transparent lakes. We contend that this novel indicator has relevance for any UV sensitive species that is constrained to shallow water environments in more transparent lakes. It may, for example, be helpful for preventing the spread of two of the most problematic aquatic invaders in North America. The Asian clam (Corbicula fluminea) and zebra mussel (Dreissena polymorpha) have inflicted a substantial economic and ecological toll on the numerous aquatic habitats that they

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have invaded (Pimentel 2005, Karatayev 2007). C. fluminea is already a problematic invader in Lake Tahoe (Tahoe State of the Lake Report 2009, UC Davis) and an aggressive monitoring program is in place to prevent the introduction of D. polymorpha in the Tahoe basin. Both species require relatively warm water temperatures for reproduction, 16° C for C. fluminea (Aguirne and Poss 1999) and 12-15°C for D. polymorpha (Lvova 1994). D. polymorpha is also sensitive to even moderate hypoxia (Shkorbatov 1994) and is usually restricted to littoral and sublittoral zones (Karatayev 1998). Presumably then C. fluminea and D. polymorpha reproduction will be constrained to shallow nearshore waters. If these species are also sensitive to UV exposure then (as with warmwater fish) variable nearshore transparency could affect reproductive success, colonization, and ultimately the spatial extent of invasion. Preliminary research suggests that at least the larval stage (veligers) of D. polymorpha is sensitive to UV exposure (Gilroy 2003). The UV tolerance of C. fluminea is unknown, but exposed tissues during feeding may make this and other bivalve species sensitive to UV exposure. Whereas invasion risk for bivalves and other potentially invasive aquatic species has previously been modeled based on abiotic factors like water chemistry (Bossenbroek et al. 2007, Whittier et al. 2008) no attention has been given to UV. We suggest that a concerted research effort be made to identify species and habitats where UV may play an important role in mediating biological invasion. It is our contention that UV could have broad relevance in aquatic invasions in transparent, coldwater lakes in particular. Although few lakes are as highly transparent as Lake Tahoe, estimates from DOC measurements in North American lakes indicate that UVR transparency is relatively high throughout western, northwestern, and southeastern portions of the USA (Williamson et al. 1996). For example, based on modeling the relationship between DOC concentration and UVR attenuation, the depth to which 1% of 320-nm UV surface irradiance penetrates is greater than 1 m in 75% of lakes sampled in the western United States. About 25% of these lakes exhibit 1% 320-nm UV depths greater than 4.75 m. In the Alps and Pyrenees of Europe a survey of 26 lakes found 1% UVA depth ranges of 1.1 to 46.1 meters (Laurion et al. 2000). The DOC concentration in most of the transparent lakes sampled in the North American study is quite low (i.e. <1 mg/L) and mean DOC concentrations in the European study were 0.97 mg/L, suggesting that even small

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changes in DOC could significantly reduce current UVR levels in these lakes (Williamson et al. 1996). We have suggested that UV potentially has broad relevance (both taxonomically and geographically) as a factor controlling aquatic invasive species establishment in more transparent lakes. The development of UV attainment thresholds could therefore be an important tool to manage and prevent biological invasion in aquatic systems. UVATs increase the cost effectiveness of aquatic invasive species control by focusing management efforts in space (on ‘at risk’ sites) and time (during spawning season), and they allow managers to evaluate the effectiveness of BMPs in terms of their ability to prevent biological invasion by selecting water clarity improvement methods that most effectively increase underwater UV transparency. In this ‘era of globalization’ the magnitude of the biological invasion threat only continues to grow (Hulme 2009 J App. Ecol 46:10-18). The UVAT may be one more tool to enhance our ability to predict and prevent species invasion.

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Acknowledgements

We thank Geoff Schladow and the staff of the Tahoe Environmental Research Center for their assistance. Sudeep Chandra and Christine Ngai (University of Nevada-Reno) provided helpful input on this manuscript. Neil Winn created the GIS map of Lake Tahoe. Michael Hughes of the Statistical Consulting Center of Miami University provided data analysis assistance. Michael Cohen, Ian Lizzadro-McPherson, Amanda Gevertz, Annie Bowling, Jeremy Mack, Kevin Rose, and Sandra Connelly provided field assistance. The experimental design for in situ incubations was borrowed from Mark Olson This work was supported in part by Miami University’s Field Workshop Program.

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Table 4.1. UV transparency at each site as percent of 305-nm UV surface irradiance present at 1 meter depth.

Site (#) % Surface Irradiance Lat Long 1- Crystal Bay 61 39.24833 -119.98458 2- Sand Harbor 78 39.20680 -119.93213 3- Cave Rock 73 39.04573 -119.94933 4- Round Hill Pines 57 38.98982 -119.95358 5- Tahoe Keys 0 38.89867 -120.00343 6- Taylor Creek 5 38.94043 -120.05770 7- Emerald Bay (EB) 16 38.91505 -120.16820 8- EB- Eagle Falls Creek 2 38.92398 -120.18412 9- Meeks Bay 9 39.03655 -120.12222 10- Sunnyside 61 39.13900 -120.15305 11- Star Harbor 0 39.18260 -120.11892

Note: Percent surface irradiance is derived from mean kd305 values from once monthly June sampling 2007-2010, except sites 2 and 9 (2008-2010 only) and site 6 (2009,2010 only)

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Table 4.2. Attainment/non-attainment status for 11 nearshore sites based on a UVAT for the prevention of largemouth bass. Sites with greater than 40% (the UVAT) of surface 305-nm UV exposure still present at 1 m depth are considered in attainment and susceptibility to largemouth bass establishment is reduced. In situ experiments show mean survival (± SEM) of largemouth bass larvae in a subset of the sample sites for 4-day incubations at 1 m depth.

% surface Site UV @ 1m Attainment In situ ± SE Crystal Bay 61.0 Y 0.0 Sand Harbor 78.0 Y 0.0 Cave Rock 73.0 Y Round Hill Pines 57.0 Y Tahoe Keys 0.0 N 93.8 (6.3) Taylor Creek 5.0 N 85.0 (5.0) Emerald Bay 16.0 N 85.0 (9.6) Emerald @ Eagle Falls Crk† 2.0 N Meeks Bay 9.0 N Sunnyside 61.0 Y Star Harbor 0.0 N †in non-attainment based on UVAT but not likely susceptible to warmwater fish establishment because of low June temps (see Table 3)

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Table 4.3. Temperature (°C) at 1 m depth (or maximum depth if less than 1 m) for all sample sites from BIC profiles. Sampling dates were from 2009 as follows: May 12-13, June 18-20, July 16-18, Aug 27-29, Oct 1-2. Site May June July Aug Oct Crystal Bay 9.5 15.3 17.3 19.8 16.6 Sand Harbor 9.6 13.7 16.6 19.2 17.2 Cave Rock 13.9 16.2 18.1 16.3 Round Hill Pines 14.5 18.1 18.0 15.7 Tahoe Keys 15.8 18.3 21.2 19.4 14.9 Taylor Creek 19.1 22.2 16.4 11.2 Emerald Bay (EB) 10.0 14.9 18.4 18.6 16.2 EB- Eagle Falls 11.5 16.8 14.9 Creek Meeks Bay 14.5 17.2 16.7 10.3 Sunnyside 7.9 12.5 15.6 18.5 15.8 Star Harbor 12.8 15.5 17.4 6.7

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(A)

(B)

(C)

Figure 4.1. Yolk-sac larvae, A) bluegill (Lepomis macrochirus), B) Largemouth bass (Micropterus salmoides), and C) Lahontan redside shiner (Richardsonius egregius). Note the greater density of photoprotective pigment in the native redside larva.

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Figure 4.2. Map of Lake Tahoe indicating the location of sample sites. Site numbers correspond with the numbers listed in Table 4.1.

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16 14 12 10 8 6 Frequency 4 2 0 0‐11‐22‐33‐44‐55‐66‐77‐8 305 nm kJ/m2

Figure 4.3. Frequency plot of 305-nm UV surface exposure for 4 day windows of time in June 2009 (e.g. of all consecutive 4 day periods in the month of June, 305-nm UV surface exposure was between 3 and 4 kJ*m-2 four times). The 4 day window represents a typical (though conservative) incubation period for yolk-sac largemouth bass larvae on the nest before swim-up stage. The median value from this frequency distribution was used in calculating the UVAT -2 (i.e., Eo = 4.99 kJ*m ).

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1 BG LMB RS

0.8

0.6

0.4 Mortality Proportion Species LE99 0.2 BG 1.32 LMB 1.99 RS 13.00 0 024681012 305-nm UV exposure (kJ*m-2)

Figure 4.4. ‘Exposure-response’ curves from rooftop exposure experiments for bluegill (BG), largemouth bass (LMB) and Lahontan redside minnow (RS) larvae. Dashed lines indicate upper and lower 95% confidence limits. Calculated LE99 values are displayed (SAS v 9.2 proc -2 GLIMMIX). The LE99 value (305-nm UV kJ*m ) for largemouth bass was selected as the effective UV exposure level used to achieve the target amount of bass mortality. This UV- exposure level (i.e. 1.99 kJ*m-2) induced high mortality (≥99%) in bass and bluegill larvae, but low mortality in the native Lahontan redside larvae (<1%).

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Concluding Remarks

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In the last 15 years ecologists have developed a more nuanced understanding of the role that UV plays in aquatic systems and in lakes especially. This new perspective recognizes not only the direct negative effects of UV exposure for individual organisms but also the complex and potentially beneficial effects of UV on both individuals and whole ecosystems. By synthesizing many of the key insights regarding the ecology of UV in recent years the first chapter of this dissertation aimed to draw attention to the implications of UV for disease dynamics, contaminant toxicity, carbon cycling, and biodiversity in lakes. Many questions remain in each of these subject areas, but the final three chapters in this dissertation provide some novel insights into at least one area of uncertainty, namely “Does UV play a part in regulating biological invasions?” The answer, I think, is a resounding yes!

A primary goal of this dissertation was to test the role of UV as it relates to warmwater fish invasion in transparent lakes using Lake Tahoe as a case study. An initial study (described in chapter 3) established the ‘proof of concept’ indicating that UV is a potent stressor that mediates the suitability of nearshore habitats for larval bluegill in Lake Tahoe. The study used dosimeters of raw DNA in solution to assess potential UV effects on larval bluegill. By relating DNA damage levels in dosimeters with larval bluegill mortality this study showed that UV is a potent force contributing to the suitability of nearshore habitats for successful bluegill reproduction in Lake Tahoe. UV exposure was high enough to reduce reproductive success of bluegill in the majority of nearshore sites sampled. This initial work also identified dissolved organic carbon and chlorophyll as two key mechanisms underlying UVR transparency in Lake Tahoe, suggesting an important link between habitat invasibility and the regional and global environmental changes that alter UV attenuating substances in lakes. For example, climate change and increased atmospheric deposition of nutrients will likely increase DOC concentrations and/or induce eutrophication in some transparent lakes, which in turn could facilitate the establishment of exotic species in formerly unsuitable habitats. A subsequent study (described in chapter 4) built on the foundation laid in chapter 3 an invasion-window model which predicted that invasion potential (i.e. reproductive success as larval survival) for warmwater fish larvae should be greatest when UV is low and temperatures are at least moderately warm. The model focused on larval fish because their immobility prevents them from avoiding UV or moving out of sub-optimal temperature conditions where

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they may be spawned and also because the ability for a species to establish self-perpetuating populations is a critical stage in the invasion process. The invasion-window model was parameterized using data from UV exposure experiments with largemouth bass larvae collected from Lake Tahoe and data from the literature detailing the effect of temperature on hatching success of bass embryos. The model results suggested that larval bass survival (and thus establishment potential) varied seasonally, spatially, and over depth gradients and was constrained by UV and temperature conditions. In situ incubation experiments confirmed model predictions. Non-native bass were significantly more UV sensitive than native redside minnows. In the absence of calculated ‘invasion-windows’ for the native redside minnow, results from in situ experiments suggested support for the expectation that native species overlap broadly with the range of ambient environmental conditions in the Tahoe littoral zone.

A second major goal for this research was to apply the findings in a relevant management context to help lake managers answer the questions, “How do we predict where invasive species will occur and how do we prevent their establishment?” The final chapter leveraged the strong differences in UV tolerance between native and non-native species to develop a UV transparency threshold (the UVAT) that if realized could prevent larval warmwater fish survival in Lake Tahoe. The UVAT was presented as a target value for water transparency based on 1) incident solar UV exposure levels during peak spawning season, and 2) the experimentally derived UV exposure levels lethal to larval warmwater fish. In situ incubation experiments across a gradient of UV transparency corroborated the validity of the UVAT. The UVAT approach improves cost effectiveness of management efforts in at least 2 ways. First, it allows lake managers to focus their efforts in space. The UVAT effectively identifies ‘at risk’ sites where warmwater fish establishment is possible. Rather than operating monitoring programs for the entire lakeshore, the UVAT suggests regular monitoring is necessary only at sites where UV and temperature conditions will permit non-native nesting. Second, because the UVAT focuses on a single life history stage (the critical larval stage) the approach focuses efforts in time. Rather than trying to mitigate invasion based on prevention and removal of adult fish the UVAT approach emphasizes the importance of monitoring during the potential warmwater fish spawning season (May- July). From a management perspective this approach is also unique because it suggests that invasive species mitigation (i.e. complete elimination of the invasive species), which has often been an

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unrealistic management outcome, is possible if UV transparency can be maintained at levels capable of prevent non-native fish establishment.

In addition to the useful management tool, a key contribution of this research is the invasion-window model for understanding invasion based on specific abiotic factors (UV and temperature) that might ultimately constrain distribution in an invaded range. It was important to develop this model in a way that took into account the response of potential invaders to these abiotic factors (species invasiveness) and levels of these factors in the environment (habitat invasibility). Other approaches for predicting habitat invasion have tended to rely on the role of biotic constraints that may or may not always be present in an uninvaded range or they focus on either the susceptibility of habitats to invasion or the traits of invaders that promote invasiveness. Few studies have integrated both habitat invasibility and species invasiveness and fewer still have focused primarily on abiotic forcing. None that I am aware of have investigated the role of UV in a biological invasion context. In this dissertation I have integrated all of these elements into a single model. This model will move the field of invasion ecology forward in a few key ways:

1) It provides a template that could be used to assess invasion potential for UV sensitive species in clear lakes threatened by declining water transparency and/or rising water temperature.

2) The model could be applied as a powerful tool for exploring some unanswered questions in invasion biology related to invasive species impacts including, “What is the potential for invasive species to promote ‘invasional meltdown?’ and “Are ‘stress specialists’ generally competitively inferior in more benign habitats?”

3) It opens the door for new questions to be asked like, “Are UV defense mechanisms important characteristics for invasive species?”

In the closing paragraphs I discuss these important contributions of the invasion-window model for future research in invasion ecology.

The invasion-window model based on UV and temperature is an ideal template for helping to understand the potential for invasion in the many lakes around the world that face the 136

triple threat (exemplified by Lake Tahoe) of increasing temperature, declining water transparency, and invasive species. This includes Lake Baikal, one of the most biologically and culturally significant lakes in the world. Future research should be directed towards applying this model, developed for warmwater fish in Lake Tahoe, to other systems and for other species. Some relevant examples include: common carp establishment in Lake Baikal (discussed in chapter 3), bivalve establishment in Lake Tahoe and other transparent alpine and sub-alpine lakes (discussed in chapters 3 and 4), and invasive macrophyte establishment including Eurasian milfoil and Canadian waterweed (discussed in chapters 3 and 4).

I have emphasized how this model successfully integrates two of the key themes in invasion ecology, species invasiveness and habitat invasibility. The model could also help to explore questions regarding invasive species’ impacts (the third major theme in invasions). In particular this model begs the question, do invasive species promote ‘invasional meltdown?’ Invasional meltdown, the notion that synergistic interactions among invaders lead to accelerated impacts on native ecosystems, still has only anecdotal support. The invasion-window model could help to explicitly test the invasional meltdown hypothesis because it establishes a potential mechanism whereby one species could facilitate the invasion of another (discussed in chapter 3). For example, bluegill are important planktivores in lakes as both adults and fry. Large bluegill populations have been linked to reduced water clarity in some systems. Thus, bluegill could potentially exacerbate water clarity loss and facilitate the establishment of other non-native UV sensitive species.

The invasion-window model showed that high stress (i.e. high UV, low temperature) environments were less susceptible to invasion. This is often the case and it has generally been assumed that species native to these high stress environments (i.e. ‘stress specialists’) would be competitively inferior in more benign habitats, implying the potential for significant negative impacts of non-natives on natives in disturbed habitats. The invasion-window model offers an opportunity to test this assumption. For example, absence of non-native species in habitats where the invasion-window model predicts establishment could be indicative of strong biotic control by native species. On the other hand if ‘stress specialists’ are indeed competitively inferior then the incidence of non-native species in sites where the invasion-window model predicts establishment

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may be high. Incidentally, in the nearshore habitats of Lake Tahoe where UV and temperature have enabled successful establishment of warmwater fish densities of the native ‘stress specialists’ have been significantly reduced.

Invasion ecologists have largely abandoned the search for a single species trait that controls invasion success in all habitats. However, the search continues for species traits that may confer invasiveness in particular habitats. The example presented herein suggests that UV tolerance may not be an especially important trait among successful invasive species. Bass and bluegill have established in some locations despite being UV intolerant. In fact, if we assume that there is some ‘cost’ in tolerating stress (as suggested in the preceding paragraph) perhaps these species are actually more effective invaders in some habitats because they are UV intolerant. Nevertheless, it is generally assumed that high physiological stress tolerance is a trait shared by successful invaders. Future research should aim to find whether high UV tolerance is common among invasive species. If it is, then the mechanisms underlying UV tolerance (e.g. high levels of photoprotective pigments or an ability to detect and avoid UV) might be useful predictors for invasion success in some high UV systems. If UV tolerance is quite variable among invaders then perhaps largemouth bass and bluegill (as physiologically intolerant species) are not so much of an exception as a new rule.

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