The Role of Forest Composition on Pool-breeding Amphibians: Colonization, Larval Communities, and Connectivity

Dissertation

Presented in Partial Fulfillment of the Requirements for the Degree Doctor of Philosophy in the Graduate School of The Ohio State University

By Michael Paul Graziano, M.S. Graduate Program in Environment and Natural Resources

The Ohio State University 2017

Dissertation Committee: Stephen N. Matthews, Advisor H. Lisle Gibbs Patrick Charles Goebel Roger A. Williams Copyright by

Michael Paul Graziano

2017

! Abstract

Few studies investigate the intricate effects that the plant community has on amphibian populations. Plants shape ecosystems, affecting both physical and chemical attributes of the landscape. Conspicuous artifacts of the plant community include canopy cover, physical structure, and modified temperature and moisture profiles. However, less conspicuous artifacts of the tree community, namely the physiochemical characteristics of their resulting leaf litter, have the ability to shape their ecosystem just as greatly as their more conspicuous traits. Leaf litter represents the primary energy source in vernal pools and other aquatic systems that are critical amphibian breeding habitat. As plant communities shift across the landscape due to ecosystem degradation, invasion by nonnative species, climate change, and shifting disturbance regimes, there is a critical need to investigate how these potential changes can influence the amphibian community and the mechanisms by which they occur.

My research investigates how the vernal pool breeding amphibian community responds to differing plant communities across a heterogeneous forested landscape and throughout their life cycle. As such, my overall objectives are multi-faceted: (1) to determine if the tree community impacts colonization and use of vernal pool-breeding amphibians (2) to scale up mesocosm studies that document the strong regulatory response tree litter inputs can have on growth and development of amphibian larvae to a

! ii! field setting (3) to determine if small, constructed ridge-top pools are a viable option for enhancing amphibian populations in the landscape, particularly with regards to increasing functional connectivity and maintaining diverse amphibian communities, and (4) to establish a landscape-level study design for conducting future, field-based experiments that can serve as a baseline to document changes in forest ecosystems. These objectives are addressed within each of my primary research pursuits below.

The first component of my research investigated the underlying drivers of colonization of novel breeding sites by amphibians, with a focus on the surrounding forest community. Fourteen ponds were created of similar size and depth along a gradient of different tree communities ranging from oak-dominance to maple-dominance Vinton

Furnace State Experimental Forest in southeast Ohio in June 2014. In 2015 and 2016 we documented colonizing amphibians, and within 10 months of construction, pool-breeding amphibians had colonized all pools. 1,114 captures were recorded comprising 12 species and 817 unique captures. Generalized linear model-based analyses performed at the community level consistently found that the tree community is significant predictor of amphibians that colonize isolated woodland pools. Additional predictors of the community included year (age), initial pH, percent coverage of leaf litter, and canopy openness of pools. As plant communities shift within the landscape due to invasion by non-native species, climate change, and an alteration of disturbance regimes, amphibian communities may also shift in response.

The second component of my research focused on the larval community occupying these constructed pools. The vast majority of studies investigating litter effects utilize mesocosms, and I sought to determine if these effects carry over into the field. We

! iii! modeled the growth rate, community structure, and size/mass at metamorphosis of four amphibian species to naturally colonize the pools: American toads (Anaxyrus americanus), chorus frogs (Pseudacris crucifer + P. brachyphona), wood frog

(Lithobates sylvaticus), and Jefferson salamanders (Ambystoma jeffersonianum). The percentage of allochthonous input derived from oak and hickory trees was a significant predictor in all models for all species, though the effect varied by species. There were also significant effects from year, canopy openness, density, and dissolved oxygen. Our study supports findings from mesocosm experiments suggesting an effect of leaf litter input on amphibian larvae and lends further evidence that shifting plant communities may carry over into the amphibian community.

The third component of my research addressed aspects of connectivity and dispersal of pool-breeding amphibians, by examining the degree of breeding site connectivity in a managed forest and the degree to which that effected subsequent colonization by amphibians in constructed ridge-top pools. Most studies on dispersal and connectivity of pool-breeding amphibians focus on fragmented landscapes, and while this research remains important, the ability of species to dispersal through relatively unfragmented landscapes should not be overlooked, as it is these landscapes (i.e. national forests, parks) that harbor some of the most robust amphibian populations. We utilized abundance data of species colonizing constructed ridge-top pools to model landscape connectivity based on three metrics: distance to natural wetlands, slope to natural wetlands, and area of natural wetlands. We found that Jefferson and spotted salamander

(Ambystoma jeffersonianum and A. maculatum, respectively) and eastern newt

(Notophthalmus viridescens) connectivity was largely dictated by the distance and mean

iv! ! slope to wetlands. Anurans were difficult to accurately assess and were more diverse in their responses than salamanders. My results show connectivity is species specific, and is largely an artifact of distance and slope to existing wetlands in relatively unfragmented landscapes. Further, this shows that connectivity of pool breeding amphibian populations may be enhanced by strategically placing breeding sites on the landscape that capitalize on dispersing individuals.

Collectively my results show that the tree community impacts pool-breeding amphibians from the larval stage through maturity, and that wide scale shifts in the plant community may alter the amphibian community as well. Perhaps more importantly, we used a field setting to confirm that results from mesocosms can carry over into the field, highlighting important underlying mechanisms, but also demonstrating additional abiotic determinates that modulate the magnitude of these effects. Further, this research provides insight to patterns and factors that structure initial colonization of novel breeding habitats by pool-breeding amphibians. From an application and management perspective, these results support the construction of small, ridge-top pools as a means of enhancing or buffering local amphibian populations. Additionally, these pools can be quickly and easily constructed in managed forests, potentially enhancing breeding and non-breeding habitat heterogeneity for the amphibian community.

v! !

For my family, especially my mom and dad, who have always supported me

! vi! Acknowledgments

“It takes a village to raise a child”: a quote that no graduate student should ever forget, for it is equally applicable to them. I have been fortunate to have the support of a committee during this journey that exceeded my expectations in every aspect. Steve

Matthews is owed more appreciation and gratitude than I am afraid I’m ever capable of repaying. Steve accepted me as a PhD student, quickly learning how to tackle the nearly insurmountable task of keeping me both on task and in good spirits. He allowed me develop a project that I was passionate about, supported my non-academic interests and buffered me from unnecessary stressors. He pushed me when I needed to be pushed, encouraged me I needed encouragement, and knew when to step back and give me the space to grow. I am indebted to Lisle Gibbs for providing me with a new lens through which to frame and handle research questions. Charles Goebel spurred my interest in ecological restoration and provided me with alternative ways for viewing and handling issues in restoration. Roger Williams provided me with the invaluable perspective of real-world practicality and applications of research, helping me to appreciate the science of forest ecosystem management. I would be remiss if I did not acknowledge Bill

Peterman, my “unofficial” committee member who spent hours guiding me through the statistical process, discussing potential projects, and served as both friend and mentor.

! vii! The research that I envisioned would never have come to pass without the many supporting agencies and people that were involved. Funding support was received from

The Ohio State University, the ODNR-Division of Wildlife administered through the

Ohio Biodiversity Conservation Partnership, and the U.S. Fish and Wildlife Service.

Additional funding was provided by the Nebraska Herpetological Society. The Ohio

Division of Forestry and United States Department of Agriculture: Northern Research

Station provided field resources and support. A special thanks is extended to my field technicians, who braved the cold and wet amphibian breeding season. My field season would not have been nearly as enjoyable without the assistance of these field technicians:

David Alsbach, David Munoz, and Steven Maichak. Special thanks also goes to Bill

Borovicka and David Runkle, who were not only a friendly voice during an otherwise solitary period, but also were incredibly supportive and never failed to assist me when needed (especially when your truck becomes stuck in the mud…).

Many other people assisted with this project, fostered my development as a graduate student, and provided me the camaraderie necessary during this time, including

Jason Tucker, Sarah Rose, Bryant Dossman, Chris Johnson, Olivia Smith, Erin Andrew,

Jim Palus, Bryce Adams, Kaley Donovan, Katie Robertson, Linnea Rowse, and Adam

Kautza. I could not have gotten to this point without the support of fellow graduate students Jason Tucker and Sarah Rose, who became some of my most loyal and devoted friends. Finally, I would like to thank my family, Grandma and Grandpa, and Zach, who have never doubted my abilities even when I doubted them myself. They always for and understood my departure and absence with the onset of the first spring rains. Thank you all for your lifetime of support.

! viii! Vita

June 2006……...……….B.S. Animal Behavior and Evolution, The Ohio State University March-July 2007…………………..Forest Salamander Research Assistant, Virginia Tech January 2008-May 2010……………………....M.S. Biology, The University of Nebraska December 2010-present……..………...……Adjunct Instructor, Biology, Columbus State Community College August 2012-present……………………….Doctoral Student, School of Environment and Natural Resources, The Ohio State University

Publications Graziano, M.P. 2017. Plethodon wehrlei (Wehlre’s salamander): Arboreal Behavior. Herpetological Review. In press.

Graziano, M.P. Ravine Amphibians. Ravinia. Spring/Summer and Fall/Winter 2014.

Gerald, G.W., M.P. Graziano, and K.E. Thiesen. 2013. Ophisaurus attenuatus attenuatus: Death feigning. Herpetological Review 44(1): 93-94.

Gerald, G.W. and M.P. Graziano. 2011. Serpentes: Storeria dekayi texana: (Texas Brownsnake): Death Feigning. Herpetological Review 42(3): 444-445.

Graziano, M.P. and J.A. Fawcett. The natural history of small-mouthed salamanders in Nebraska. In preparation.

Wilson, J.A. and M.P. Graziano. Metabolic rates of Great Plains anurans in relation to overwintering success in a changing climate. In preparation.

Fields of Study Major Field: Environment and Natural Resources Specialization: Ecological Restoration

ix! ! Table of Contents

Abstract……………………………………………………………………………………ii Dedication…………………………………………………………………..…………….vi Acknowledgments…………………………………………………………………….…vii Vita….…………………………………………………………………………………….ix List of Tables………………………………………………………………………..…...xii List of Figures……………………………………………………………………………xv Chapter 1: Plant Communities and Associated Litter: A Review of the Impacts on Amphibian Growth, Development, and Behavior………………………………...………1 Abstract……………………………………………………………………………1 Introduction……………………………………………………………..…………2 Methods……………………………………………………………………………5 Results…………………………………………………………………………..…5 Discussion………………………………………………………………………..17 References………………………………………………………………………..22 Chapter 2: Does forest composition influence colonization of vernal pool-breeding amphibians?……………………..………………………………..……………………...29 Abstract…………………………………………………………………………..29 Introduction………………………………………………………………….…...30 Methods…………………………………………………………………………..32 Results……………………………………………………………………………39 Discussion………………………………………………………………………..41 References………………………………………………………………………..48 Chapter 3: Responses of Larval Amphibian Communities to Variations in Litter Input Across a Naturally Existing Oak-Maple Gradient……………………………………….62 Abstract…………………………………………………………………………..62 x! ! Introduction………………………………………………………………………63 Methods…………………………………………………………………………..65 Results……………………………………………………………………...…….72 Discussion………………………………………………………………………..74 References………………………………………………………………………..81 Chapter 4: Amphibian Connectivity in a Heterogeneous Landscape: The Role of Constructed Ridge-top Pools…………………………………..………………………...94 Abstract…………………………………………………………………………..94 Introduction………………………………………………………………………96 Methods…………………………………………………………………………..99 Results…………………………………………………………………………..104 Discussion………………………………………………………………………105 References…………………………………………………………...………….112 Bibliography……………………………………………………………………………124

xi! !

List of Tables

Table 2.1. Total amphibian captures in constructed pools for 2015-2016. *Denotes species omitted from statistical analyses. Numbers in ( ) indicate unique captures for each species……………………………………………………………………………………57

Table 2.2. Candidate models ranked by relative support for amphibian community in constructed pools with Akaike Information Criteria (AIC) scores, Δ AIC, and Akaike parameter weights (ωρ)* Indicates significant variables (p≤0.05)……………………….58

Table 2.3. Coefficients of species comprising amphibian community in fourteen constructed pools to associated variables in top-ranked model in Vinton Furnace State Experimental Forest for 2015-2016……………………………………………………...59

Table 2.4. Alternate models for colonizing amphibian community in constructed pools with “year” omitted from explanatory variables with Akaike Information Criteria (AIC) scores, Δ AIC, and Akaike parameter weights (ωρ) * Indicates significant variables (p≤0.05)………………………………………………………………………………..…60

Table 2.5. Coefficients of species comprising amphibian community in fourteen constructed pools to associated variables in top-ranked model in Vinton Furnace State Experimental Forest with the omission of “year” as a variable for 2015-2016………….61

Table 3.1. Total numbers of larvae sampled from species utilizing consructed pools in Vinton Furnace State Experimental Research forest from 2015-2016 used for growth analyses. * denotes species for which we conducted further analyses………………….88

Table 3.2. Models for growth rate of larval Jefferson salamanders (Ambystoma jeffersonianum) in constructed pools in Vinton Furnace State Experimental Forest in 2015 and 2016. Marginal r2 (r2 m) describes variance explained by fixed effects. Conditional r2 (r2 c) describes variance explained by fixed and random effects. Akaike parameter weights (ωρ). * denotes significant predictors (p<0.05)……...………………89

Table 3.3. Models for growth rate of larval American toads (Anaxyrus americanus) in constructed pools in Vinton Furnace State Experimental Forest in 2015 and 2016. Marginal r2 (r2 m) describes variance explained by fixed effects. Conditional r2 (r2 c) describes variance explained by fixed and random effects. Akaike parameter weights (ωρ). * denotes significant predictors (p<0.05)…………………………………………..89

xii! ! Table 3.4. Models for growth rate of larval chorus frogs (Pseudacris spp.) in constructed pools in Vinton Furnace State Experimental Forest in 2015 and 2016. Marginal r2 (r2 m) describes variance explained by fixed effects. Conditional r2 (r2 c) describes variance explained by fixed and random effects. Akaike parameter weights (ωρ). * denotes significant predictors (p<0.05) ……………………………………………………….. ...90

Table 3.5. Models for growth rate of larval wood frogs (Lithobates sylvaticus) in constructed pools in Vinton Furnace State Experimental Forest in 2015 and 2016. Marginal r2 (r2 m) describes variance explained by fixed effects. Conditional r2 (r2 c) describes variance explained by fixed and random effects. Akaike parameter weights (ωρ). * denotes significant predictors (p<0.05)………………………………………….90

Table 3.6. Total numbers of larvae sampled from species utilizing constructed pools in Vinton Furnace State Experimental Research forest from 2015-2016. * denotes species underrepresented in sampling due to incongruous breeding season…………………….91

Table 3.7: Five most supported models for larval amphibian community in constructed pools in Vinton Furnace State Experimental Research forest from 2015-2016 with Akaike Information Criteria (AIC) scores, Δ AIC, and Akaike parameter weights (ωρ) * Indicates significant variables (p≤0.05)……………………………………………………………92

Table 3.8: Most supported models for size and mass at metamorphosis for four amphibian species in constructed pools in Vinton Furnace State Experimental Forest with model parameters with associated coefficients and Akaike parameter weights (ωρ). * Indicates significant variables (p≤0.05)…………………………..………………………………..93

Table 4.1. Factors influencing site connectivity of Jefferson salamanders (Ambystoma jeffersonianum) for 2015-2016. * indicates significant model (p<0.05) using Poisson distribution. α = 1500m…………………………………………………………………117

Table 4.2. Factors influencing site connectivity of spotted salamanders (Ambystoma maculatum) for 2015-2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 1500m………………………………………………………..117

Table 4.3. Factors influencing site connectivity of wood frogs (Lithobates sylvaticus) for 2015-2016. * indicates statistically significant model (p<0.10) using Poisson distribution. α = 1500m………………………………………………………………………………118

Table 4.4. Factors influencing site connectivity of green frogs (Lithobates clamitans) for 2015-2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 2500m………………………………………………………………………………118 Table 4.5. Factors influencing site connectivity of chorus frogs (Pseudacris crucifer + P. brachyphona) for 2015-2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 500m…………………………………………………………119

xiii! ! Table 4.6. Factors influencing site connectivity of American toads (Anaxyrus americanus) for 2015-2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 2500m………………………………………………………..119

Table 4.7. Factors influencing site connectivity of four-toed salamanders (Hemidactylium scutatum) for 2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 1000m…………………………………………………120

Table 4.8. Factors influencing site connectivity of eastern newts (Notophthalmus viridescens) for 2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 2500m…………………………………………………………………120

Table 4.9. Total adult amphibian captures in constructed pools for 2015-2016. *Denotes species omitted from statistical analyses. Numbers in ( ) indicate unique captures for each species. **Denotes species that chorused and oviposited in pools, but were not represented in captures………………………………………………………………….121

Table 4.10. Matrix with values for distance (m) and mean slope (%) from constructed pools (Cons.) to respective naturally occurring pools at Vinton Furnace State Experimental Forest…………..……………………………………………………122-123

xiv! ! List of Figures

Figure 1.1 Conceptual diagram illustrating the potential cascading effects of the plant community on the amphibian larval community………………………………………...28

Figure 2.1 Topographic map for a section of Vinton Furnace State Experimental Forest in Vinton County, Ohio with locations of pool pairings constructed in 2014. Alphanumeric codes correspond to treatment (O=oak, M=maple) and pair number (ex. O1 = oak 1).…54

Figure 2.2. Distribution of constructed pools at Vinton Furnace State Experimental Forest across gradient of tree communities based on importance value index (IVI)……………55

Figure 2.3. Non-metric multidimensional scaling (NMDS) ordination (stress = 0.133) using contour lines to depict the gradient of the oak importance value index on the landscape with respect to the amphibian community. Circles represent pools as they occur on the landscape, and four-letter abbreviations are code using the first two letters of the genus+species where multiple genera are present, or the first four letters of the genus where only one species is present. *Spring peeper, American toad, four-toed salamander, and eastern newt beyond margins of plot……………………………………….……….56

Figure 3.1. Four of fourteen constructed pools used in larval growth study, located at Vinton Furnace State Experimental Forest, Ohio. Pools measure approximately 5 meters diameter, with one of each pairing in a predominantly oak community and the other in a predominantly maple community………………………………………………………..85

Figure 3.2. Floating leaf litter basket used to sample litterfall in constructed pools. Baskets were tethered to the margins of the pool and were deployed from September through December. One basket was used for each pool and accurately represented the oak-hickory component of the surrounding tree community…………………………….86

Figure 3.3: Efficacy of floating litterfall traps to represent the importance value of the surrounding oak community as measured by percent of total litter sample captured 2 (r adj. = 0.85)………………………….……………………………………….…………87

Figure 4.1. Shaded relief map of Vinton Furnace State Experimental Forest with locations of natural and constructed pools. Blue triangles represent constructed pool pairs and black stars denote locations of naturally occurring pools.…………………………116

xv! ! Chapter 1: Plant Communities and Associated Litter: A Review of the Impacts on Amphibian Growth, Development, and Behavior

Abstract Plants shape ecosystems, affecting both physical and chemical attributes of the landscape. Conspicuous artifacts of the plant community include canopy cover, physical structure, and modified temperature and moisture profiles. However, less conspicuous artifacts of the tree community, namely the physiochemical characteristics of their resulting leaf litter, have the ability to shape their ecosystem just as greatly as their more conspicuous traits. Leaf litter represents the primary energy source in vernal pools and other aquatic systems that are critical amphibian breeding habitat. As plant communities shift across the landscape due to ecosystem degradation, invasion by non-native species, climate change, and shifting disturbance regimes, there is a critical need to investigate how these potential changes can influence the amphibian community and the mechanisms by which they occur. Variations in bulk elements (carbon, nitrogen, phosphorus), structural compounds (lignin and cellulose), and secondary metabolites (i.e. tannins) represented in leaf litter can impact the larval amphibian community through multiple pathways. This review highlights those ecologically relevant physical and chemical characteristics of the plant community, how these impact the physiology, behavior, and subsequent fitness of amphibians in order to ascertain what effect these changes may have on amphibian populations. Further, this synthesis of the relevant litter provides direction for future research.

! 1! Introduction

Amphibian Populations: Declines and Vulnerability

Amphibian populations have experienced noted worldwide declines for over half a century with habitat loss and degradation, disease, invasive species, climate change, and their synergistic effects frequently cited as the causes behind observed declines

(Blaustein et al. 1994; Houlahan et al., 2000; Collins and Storfer, 2003; Beebee and

Griffiths, 2005). Habitat loss, fragmentation, and degradation represent the greatest threats (Beebee and Griffiths, 2005; Cushman, 2006), impacting nearly 90% of threatened amphibians (Baillie et al., 2004). Amphibians represent particularly vulnerable taxa for numerous reasons, including low vagility, narrow physiological (and thus, environmental) tolerances (Houlahan et al., 2000), and complex life histories, which expose them to multiple stressors throughout their annual cycle (deMaynadier and

Hunter, 1995; Cushman, 2006). For example, amphibians with biphasic life history are exposed not only to aquatic stressors as larvae, but also to terrestrial stressors as juveniles and adults. Moreover, as the integument of amphibians is permeable and must remain moist to perform its role as a respiratory organ, the risk of dehydration increases their vulnerability to drought and desiccant-prone microhabitats (deMaynadier and Hunter,

1995; Stebbins and Cohen, 1995; Peterman and Semlitsch, 2014).

Habitat Degradation: The importance of plant communities

While habitat loss and degradation are widely regarded as the most significant cause of amphibian decline (Beebee and Griffiths, 2005; Cushman, 2006), it is frequently studied at the landscape scale of fragmentation, connectivity, and isolation (see Cushman,

2006). However, as many populations of amphibians are in decline in seemingly intact

! 2! habitat (Beebee and Griffiths, 2005; Cushman, 2006), a different lens may be necessary to unveil the direct effects behind species decline. Amphibians lack a high degree of vagility (Bowne and Bowers, 2004) and this lack of vagility may cause individuals to be captured in patches of degraded/degrading habitat quality should they be unable to disperse to new patches (Bowne and Bowers, 2004).

Plant communities, although generally noted, are often quantified from a structural perspective, focusing on canopy cover, coarse woody debris, herbaceous forbs, and litter depth. While these represent valuable data, the overall effect of the plant community may consistently be underestimated with respect to the corresponding amphibian community-both from a terrestrial and an aquatic perspective. Plant communities have foundational impacts on ecosystems, effecting microclimate, substrate, and nutrient availability (Bazzaz, 1979).

Plant litter represents a resource in both terrestrial and aquatic environments that varies in both quantity and quality (Stoler and Relyea, 2013). Moreover, differential litter inputs may directly impact the prey-base for subsequent trophic levels (Rubbo and

Kiesecker, 2004; Stoler and Relyea, 2011). Amphibians occupy various mid-upper trophic levels, and a recent pulse in research investigating performance with respect to plant litter – both non-native and native –has revealed a wide-range of species-specific effects: from retarding to accelerating growth and development, to impacting size at metamorphosis, survival, and fitness (Scott, 1994; Maerz et al., 2005a,b; Maerz et al.,

2010; Brown et al., 2006; Williams et al,, 2008; Rittenhouse, 2011; Watling et al, 2011;

Adams and Saenz, 2012; Cohen et al., 2012a; Rogalski and Skelly, 2012; Martin and

Blossey, 2013).

3! ! The exposure of amphibians to different or novel plant species, each varying in quality and quantity of litter coupled with the different chemical compounds inherent with that litter, may represent a widespread, yet cryptic, threat to amphibian populations should those species exert negative effects (Maerz et al., 2005; Brown et al., 2006;

Watling et al., 2011; Martin and Blossey, 2013). As such small scale, patchy transformations of plant communities could manifest into habitat suitability shifts. With disease, climate change, and invasive species shifting and redistributing plant communities worldwide, it stands to reason that amphibian communities may also experience shifts. Well-known, pervasive situations exist in eastern North America with the mortality of widespread and important species, including ash (Fraxinus spp.) and eastern hemlock (Tsuga canadensis), which will undoubtedly lead to landscape-level changes in habitat (Stephens et al., 2013). Although research is limited (albeit rapidly growing), the effects of plant communities on amphibians are important in understanding current changes in the distribution and abundance of amphibians, and may be important in predicting potential effects of future changes (Rubbo and Kiesecker, 2004).

Purpose and Goals

The objective of this review is to synthesize current knowledge regarding the effects of plant communities and resulting litter on amphibians, identify information gaps for understanding amphibian responses to changes in plant communities, and suggest approaches for future data collection and in some cases the potential for reanalysis of past data. This paper is divided into four major sections to best accomplish this goal. The first section is a review of ecologically relevant chemical and physical properties of leaf litter. The second and third sections review what is known about the effects leaf litter on

4! ! terrestrial and aquatic life stages, with particular focus on known and potential physiological and behavioral impacts. A conceptual diagram is provided to illustrate the foundational impacts of the plant community on amphibians to synthesize sections two and three. Finally, in context of these reviews, important information gaps and research needs will be suggested.

Methods

A qualitative review was conducted of primary literature published up until 2016 using the Web of Knowledge database with the search terms: amphibian, larvae, leaf litter, and wetlands. This review was assembled from keywords: (amphibian*) and

(leaf*) and (litter*) or (amphibian*) and (larv*) and (litter*). The results of this literature search were further reduced by focusing on research that incorporated growth, development, survival, and/or behavior of larvae, particularly in isolated pool systems.

Results

Chemical properties of leaf litter

Bulk Elements: Carbon, Nitrogen, and Phosphorous

Nitrogen is a key element for aquatic organisms as it is essential to the production of plant and animal tissue and protein synthesis, and represents a focal component of many studies involving litter quality. Nitrogen content generally indicates how quickly litter will decompose: high nitrogen content litter will decompose faster than litter with lower nitrogen content (Ostrofsky, 1997). Moreover, nitrogen is often highly correlated with phosphorous (a limiting nutrient in many ecosystems) content (Stoler and Relyea,

5! ! 2013). High nitrogen content is associated with greater periphyton growth (Webster and

Benfield, 1986), which forms an important food base for aquatic systems (Schiesari,

2006). The ratio of carbon:nitrogen (C:N) and carbon:nitrogen:phosphorus (C:N:P) are generally regarded as a measure of litter quality. For example, leaves with a high C:N ratio break down more slowly and are likely to be of lower quality than those with a low

C:N ratio (Palik et al., 2006). Palik et al. (2006) found that sugar maple (Acer saccharum) leaves provide low quality food for larval invertebrate species, likely due to high C:N ratio. This low quality nutrient base resulted in a depauperate food base with limited microbial colonization. In contrast, litter with low C:N ratios will decompose quickly, presenting more of a “pulsed” food source whereas litter that decomposes slower may provide a more sustained resource (Cohen, 2013). Taylor et al. (1989) also used nitrogen (and lignin) as a predictor of litter decay rates, again finding that litter high in C was strongly correlated with low N and P and high in lignin.

Structural Polymers: Lignin and Cellulose

Cellulose and lignins represent the most abundant plant-derived polymers on

Earth. Although cellulose is a large polysaccharide, it exhibits a relatively simplistic structure and represents an intermediate quality to decomposers, as only a few enzymes are necessary to degrade it (Loranger et al., 2002). However, cellulose and lignin work in concert in plants similar to epoxy and fibers, with cellulose acting as the fibers and lignin acting as the epoxy, functioning to provide structural support (Lynch, 1992). However as lignins increase with respect to cellulose, the availability and quality of litter to decomposers decreases (Lynch, 1992; Loranger et al., 2002).

6! ! Lignins represent a complex group of hydrophobic, plant-derived polymers that are second in abundance only to cellulose (Campbell, 1996). They represent one of the most recalcitrant components of vegetation due to their complex structure and exhibit resistance to enzymatic degradation (Taylor et al., 1989). As lignin is interwoven throughout plant tissue and is resistant to decomposition, it physically interferes with decay of other chemical factions of plant tissue and is thus important to humus formation

(Melillo et al., 1982; Taylor et al., 1989; Haug, 1993). In fact, experimental evidence spanning a decade shows that the amount of humus formed is highly associated with the initial lignin contents of the material (De Haan, 1977). Further, the relationship between humus formation and initial lignin content of litter has implications for rates of nitrogen cycling in forest ecosystems, as the amount of nitrogen bound in decomposition-resistant humus can be large (Melillo et al., 1982). From an ecological perspective, humus expands the capacity of moisture retention between flood and drought conditions, providing important buffering to environmental extremes (De Haan, 1997). Lignin also influences flammability and packing ratios (a measure of fuelbed compactness and an important factor in predicting fire behavior) (Nowacki and Abrams, 2008). For example, the high lignin content of oak (Quercus spp.) leaves results in rigid and irregular structure, which decreases the packing ratio and allows oak leaves to remain dry over a longer period of time than other hardwoods, thus limiting decomposition and promoting flammability (Lorimer, 1985). In contrast, the thinner leaves and low lignin content of many mesophytic species (e.g. Acer rubrum) exhibit a high packing ratio as they adhere to the forest floor and trap moisture, which enhances decomposition and decreases flammability (Lorimer, 1985). In the aquatic environment, increased amounts of lignin

7! ! also slow or inhibit microbial growth, thus reducing the decomposition process and promoting a detritus layer (Melillo et al., 1982; Hunter et al., 2003). Perhaps most important, high lignin content can even negate over overcompensate for nitrogen, reducing what would otherwise be “high quality” litter (Melillo et al., 1982). However, the more recalcitrant nature of litter exhibiting higher lignin levels creates numerous ecological niches for invertebrate feeding guilds and provides a more sustained resource than litter with lower lignin content (Rubbo and Kiesecker, 2004).

Secondary Metabolites

Whereas elevated nutrient content (N and P) can promote microbial growth, increased concentrations of secondary metabolites (and structural compounds) can inhibit it (Stoler and Relyea, 2013). This is a product of the primary function of secondary metabolites: serving as a defensive mechanism. As such, the resistance of plants to natural threats (e.g. pathogens, herbivory) is largely determined by the chemical quality and quantity of their secondary metabolites (Witzell and Martin, 2008). Phenolics, which include tannins, are a commonly studied class of secondary metabolites abundant in the tissues of conifers and some hardwoods (see Witzell and Martin, 2008). While tannins act as a defense mechanism, especially with regard to pathogenic fungi, upon death or senescence these metabolites are released into the environment where they have variety of ecological effects (Giner-Chavez, 1996; Witzell and Martin, 2008 Watling et al.,

2011). Phenolics may also exhibit seasonal variation (often correlated with phenological stage), with resources available for synthesis of these and other secondary metabolites being limited during periods of active growth, heat, or drought; as such, plant material deposited in the environment may exhibit seasonal variation (Frutos et al., 2004).

8! ! Physical Conditions Created By Plant Community

As the previous section reviews, the plant community exerts powerful bottom-up effects on ecosystems from a chemical perspective. Perhaps a more perceptible process by which plants alter ecosystems however, are through the physical conditions those plants create. One of the most well documented conditions with respect to amphibians is the degree of canopy cover the tree community displays (Skelly et al., 2002; Halverson et al., 2003). Canopy cover determines exposure to light (and thus, primary productivity) as well as a suite of other associated characteristics including temperature, moisture (in terrestrial environments), and dissolved oxygen (in aquatic environments). The effects of canopy cover in aquatic systems may be negated or compounded by high leaf litter solubility, which may darken the water column and thus limit primary productivity

(Karlsson et al., 2009).

The depth and persistence of the humus layer on the forest floor is also dependent upon the plant community. For example, stands with species exhibiting a high amount of lignin tend to form a thicker humus layer than stands with species that exhibit low levels of lignin (De Haan, 19777; Melillo et al., 1982). In the absence of fire a variety of mesophytic, highly competitive, later-successional plant species (i.e. red maple (Acer rubrum), American beech (Fagus grandifolia), and tulip poplar (Liriodendron tulipifera) are replacing shade-intolerant oaks and hickories (Quercus spp. and Carya spp., respectively) (Nowacki and Abrams 2008). The result of this replacement is an altered understory microclimate characterized by heavy shade (decreased solar radiation) and limited air movement (increased humidity) - effectively a cooler and wetter understory and forest floor (Nauertz et al., 2004). As leaf litter thickness, specific mass, and

9! ! structure (i.e. lignin content) greatly affect humidity levels in litter beds, this altered understory produces a seedbed more conducive to mesophytic plant species and creates a positive feedback loop for this cycle (Nowacki and Abrams, 2008).

A largely overlooked and inconspicuous aspect of the plant community involves the ability of transpiration to alter ecosystem hydrology. Boyce et al. (2012) found that transpiration from the invasive shrub (Amur honeysuckle (Lonicera maackii)) can affect hydrology in forested ecosystems, hypothesizing that invasions by certain species may shorten the hydroperiod of ephemeral ponds and streams.

Litter Effects on Amphibian Life Stages

Aquatic life stages

Allochthonous litter provides the majority of energy and nutrients in isolated wetlands and influences not only larval survival, but also two plastic responses frequently correlated with adult fitness: size at and timing of metamorphosis (Maerz et al., 2005,

2010; Brown et al., 2006; Williams et al., 2008; Watling et al., 2011; Cohen et al., 2012a;

Earl et al., 2012; Earl and Semlitsch, 2012; Martin and Blossey, 2013). Rubbo and

Kiesecker (2004) provided one of the initial papers documenting the impacts of leaf litter on larval amphibian communities. In their study, amphibian biomass and survival were significantly lower in mesocosms containing only maple litter in comparison to those containing oak litter. Their findings lead them to propose that compositional shifts in the forest have the potential influence litter-based communities and destabilize species interactions. Since their paper, a pulse in research has investigated litter quality (C:N and

C:N:P ratios), phenolic compounds, structural compounds (cellulose and lignin) and the effect of these components on larval growth and survival. Maerz et al. (2010)

10! ! investigated detritus quality, finding that leaf mass loss, total metamorph production, and the number of species that metamorphosed declined as a function of increasing C:N ratio of plant leaves. Further, they found that mean time to metamorphosis and mean mass at metamorphosis declined as a function of increasing plant C:N ratio. Despite being associated with more recalcitrant litter, percent lignin was not a significant predictor of tadpole performance. Stephens et al. (2013) conducted a comprehensive, rigorous study investigating litter effects on amphibians. Using 10 species of litter differing in regional abundance and chemical composition and wood frogs as their model amphibian, found that litter from red maple, cattail (Typha latifolia), and sedge (Carex stricta) – abundant and widespread species – produced anuran metamorphs that were small and had poor survival. Ironically, litter from a once-omnipresent species (green ash (Fraxinus pennsylvanica)) that is now in sharp decline from the introduced emerald ash borer

(Agrilus planipennis) produced metamorphs that grew larger, developed faster, and survived better than any other treatments. Further, they found that the mass at metamorphosis (a proxy for adult fitness) was best explained by the C:N:P content.

Relyea et al. (2015) also investigated the effects of litter resource quality on wood frog

(Lithobates [Rana] sylvaticus) metamorphs, finding both external and internal morphological changes associated with litter nutrient content, decomposition rate, concentrations of phenolic acids, and dissolved organic carbon.

Stoler and Relyea (2013) demonstrated that tadpoles could respond to changes in leaf litter quality and quantity by altering their internal and external morphology. Their study found that feeding off of litter with greater N content was associated with shorter intestines, larger mouths, longer and shallower bodies, longer and deeper tails, and deeper

11! ! tail muscles and indicated that wood frogs were capable of grazing of microbial communities on litter. Interestingly, and counter to many other studies, Stoler and Relyea

(2013) found few correlations of total lignin or total phenolics with tadpole responses.

Further, not all species respond to increased N resources equally, with energy investment primarily in either growth or development. Schiesari (2006) compared the performance of two sympatric anuran species (leopard frogs (Lithobates [Rana] pipiens) and wood frog (L. sylvaticus)) finding a significant positive interaction between growth responses and litter quality, although the two species differed in their reaction. Leopard frogs gained more mass when provided with high N resources and wood frogs developed faster under the same conditions. However, despite a faster growth rate realized by leopard frogs, mortality was also increased - findings consistent with tradeoffs between intrinsic growth and mortality. This differential allocation of resources may influence patterns of species distributions across resource gradients (Schiesari et al., 2006).

Soluble phenolic compounds have been observed to exert a wide-range of disproportionate effects on amphibian larvae when compared to other taxa (Watling et al.,

2011). Tannins and other phenolics can negatively affect sensitive tissues, particularly gills of aquatic organisms, and thus impact respiration (Maerz et al., 2005). For example,

American toad (Anaxyrus [Bufo] americanus) survival and metamorphosis has consistently been found to be severely limited in the presence of high concentrations of soluble phenolics, such as those derived from nonnative-invasive purple loosestrife

(Lythrum salicaria) and the native swamp loosestrife (Decodon verticillatus) (Maerz et al., 2005b; Brown et al., 2006; Watling et al., 2011; Cohen et al., 2012a). High phenolic

12! ! concentrations have been linked with high mortality consistently enough that Cohen et al.

(2012a) suggest that this may be an important driver of American toad ecology.

With respect to phenolic compounds, while some species may exhibit the behavioral plasticity to ameliorate the negative effects, others are unable to do so.

Watling et al. (2011) compared the behavior of several amphibian larvae, finding that one species (plains leopard frog (Lithobates [Rana] blairi)) was able to compensate for comprised respiratory behavior by increasing surface breathing, in contrast to American toads, which exhibited higher mortality. As phenolics can influence species-specific behaviors, long-term consequences from the organismal level to population level may be incurred. Increased surfacing in response to lowered respiratory ability in the presence of phenolics may make sensitive tadpoles more susceptible to predation (Moore and

Townsend, 1998) and can interfere with their ability to acquire food resources through foraging, leading to negative indirect effects on larvae (Werner and Peacor, 2003).

An unexpected and recently described benefit of secondary compounds on amphibian communities involves potential resistance to fungal diseases. Davidson et al.

(2012) found that tiger salamanders (Ambystoma tigrinum) exposed to extracts of

Eucalyptus camadulensis and Quercus spp. were not only cured from infection by the lethal and widespread fungus, Batrachochytrium dendrobatidis (Bd), but also developed some resistance to reinfection. However, contrasting results were found with eastern newts (Notophthalmus viridescens), where leaf litter and vegetation were related to higher Bd infection (Raffel et al., 2010). It is important to note however, that the authors did not identify the tree community, nor did it appear in their results. Raffel et al. (2010) elaborate on their findings and suggest the leaf litter may be indicative of shading, and

13! ! thus cooler temperatures, which creates an environment that favors higher Bd growth

(Rohr and Raffel, 2010). While quantitative evidence does not yet exist, it could be hypothesized that the shift from shade-intolerant oaks may indirectly contribute to Bd infection through the cooler mesophytic conditions that are created by shade-tolerant red maple and American beech.

Interestingly, Cohen et al. (2012b) suggest that litter quantity may be a more significant determinant of larval performance, and that the increased litter quantity often associated with nonnative, invasive species invasions may have impacts that override species-specific effects. Further, as forested wetlands and streams are systems where amphibian larvae convert aquatic resources into amphibian biomass, which is later be distributed throughout the forest in the form of metamorphosed individuals, the shifting of plant communities combined with amphibian decline has the potential to alter nutrient cycling in forest ecosystems through bottom-up effects (Stephens et al., 2013).

Terrestrial life stages

Terrestrial salamanders (family Plethodontidae) and metamorphosed mole salamanders (family Ambystomatidae) are highly dependent upon leaf litter for both refugia and foraging, yet comparatively few studies investigate the terrestrial effects of different litter types on amphibians. While litter depth is a frequently recorded metric of habitat with reference to amphibians, few accounts further investigate the composition of litter and the potential impacts that it may impart. The differential thermal and physical microhabitats created by litter are well noted (Nowacki and Abrams, 2008), and those differences should translate to the amphibian community. Maerz et al. (2009) incorporated the potential impacts of nonnative plants on woodland salamander

14! ! abundance, and while no relationship between nonnative plant cover and abundance was detected, woodland salamander abundance declined exponentially with decreasing litter volume. Further, while leaf litter was not the focus of the study, Grover (1998) investigated the influence of cover and moisture on two plethodontids, finding that areas of high cover density and higher moisture levels were significant factors in regulating abundance of terrestrial salamanders. As such, it could be hypothesized that leaf litter exhibiting varying levels of recalcitrance (i.e. amount of lignin) or moisture may impact terrestrial salamanders habitat use and abundance.

Dominant plant species have been observed to influence terrestrial invertebrates

(McEwan et al., 2009), which could impact the prey-base for foraging salamanders.

McEvoy and Durtsche (2004) compared an area invaded by Amur honeysuckle with a nearby area that had not been invaded, finding that slimy salamanders (Plethodon glutinosis) and green frogs (Lithobates [Rana] clamitans) had greater body mass in non- invaded sites. Their findings suggest that a loss or alteration in native food resources in areas invaded by Amur honeysuckle may be responsible for the differences observed in mass. While this may be accurate, it is likely not portraying the entirety of the interaction: leaf litter of Amur honeysuckle is low in lignin and decomposes rapidly (Blair and

Stowasser, 2009), so it is possible that the microhabitats salamanders use for both refugia and foraging are limited in these areas.

Plant Communities: Of Native Versus Nonnative Origin

Knowledge of litter effects on amphibian communities has spurred considerable research investigating interactions between amphibians and invasive plant species. These studies elucidate the direct influence plants can have on amphibians, underscoring the

15! ! effects of leachates from litter on growth and survival. Consistently, these studies show that chemical traits are more important than nonnative origin with respect to growth and survival (Watling et al., 2011; Cohen et al., 2012a; Martin and Blossey, 2013). For example, Watling et al. (2011) found that litter from a nonnative Amur honeysuckle resulted in a shorter development period than litter derived from native species (though interestingly the litter from Amur honeysuckle exhibited a low concentration of total phenolics, which American toads are sensitive to (Cohen et al., 2012a)). Martin and

Blossey (2013) further support these findings with their study of native and introduced common reed (Phragmites australis) on amphibian larvae, finding that geographic origin is a poor predictor of effects on some amphibians. Martin et al. (2015) continued this line of research, investing the effects of litter diversity, species, origin, and traits on American toad tadpoles, further corroborating pH and litter C: N as being significant predictors of tadpole performance and physiochemical traits outweighing geographic origin of litter.

They also found, interestingly, that increased plant diversity can not be assumed to increase performance of developing larvae, as the strong negative effects of some species may outweigh diversity benefits (Martin et al., 2015). Cohen et al. (2012a) found that plant traits improved models that included well-known drivers of tadpole performance, including hydroperiod, temperature, and dissolved oxygen concentrations. Cotten et al.

(2012) incorporated an additional artifact of invasive species effects, not only underscoring their impact on amphibian growth and development, but also finding that the season in which those amphibians breed as being significant in their exposure to secondary metabolites produced by those plants. Ultimately, the impact of nonnative plant species on amphibians will be dependent upon the specific chemical traits, quantity,

16! ! and quality of their resulting litter (Watling et al., 2011; Cohen et al., 2012a,b; Martin and

Blossey, 2013; Martin et al., 2015).

Discussion

It is strikingly evident that plant communities and the litter resulting from them have profound impacts on the life history of amphibians, particularly the aquatic stages

(Figure 1.1). Schiesari et al. (2006) showed that environmental conditions promoting increased larval growth can also lead to lower larval mortality. If the plant community yields litter that decreases growth or development, population-level changes to the amphibian community may arise over a greater temporal scale as a result of lower recruitment (Schiesari et al., 2006). Moreover, lower recruitment may affect nutrient cycling in forested ecosystems, where amphibians serve to transfer aquatic nutrients to the terrestrial environment.

Considerable effort has also been invested in determining the impacts of non- native, invasive plant species on the aquatic stages of amphibian development. Results have indicated that plant origin (native versus nonnative) has no effect on amphibian performance, with litter traits (quality) ultimately determining success of individuals

(Cohen et al., 2012a). It should be noted however, that nonnative, invasive plant species are often concomitant with site-to-landscape scale homogenization (Hong and Quifeng,

2010), which may ultimately favor some species and lead to an overall reduction in diversity and hinder restoration efforts. A greater understanding of the foundational influence that novel plant communities can have on amphibians will be increasingly important as land managers, volunteer groups, and other organizations expend

17! ! considerable resources in efforts to reduce or eradicate nonnative plant species under the umbrella of wildlife conservation. Cohen et al. (2014) recommend utilizing community- weighted mean trait values for four ecologically relevant plant traits (litter N, P, ligin, and soluble phenolics) as an approach for predicting potential changes in the aquatic community from shifting forest communities. The annual phenology of tree species, including when they develop and shed leaves, must also be considered when investigating impacts to the aquatic community. For nonnative species that have begun to dominate some landscapes and whose litter will continue to be more represented in aquatic systems, this could drastically impact survival of amphibian larvae, especially when coupled with more erratic/warmer winters associated with climate change (Saenz et al., 2013).

With regards to amphibian conservation efforts, restoration of the structural and physical parameters of a system may be insufficient to successfully restore fragmented or degraded habitat. Considering what is currently known about the importance of litter quality, further exploration of associated plant communities and the physiochemical properties of subsequent litter should accompany restoration or repatriation efforts, particularly in the reference to threatened or endangered species. Should they be available, a reanalysis of data from prior studies investigating factors effecting amphibian communities may benefit from incorporating more information on the plant community, particularly litter quality. If such data are unavailable, efforts should be made to conduct new studies that further investigate historic or reference-site plant communities in areas where traditional habitat restoration efforts are ineffective.

18! ! The majority of research investigating litter quality and amphibians are focused on woody plants or forbs, with few having investigated litter quality of graminoids.

Numerous amphibian species are native to areas where the landscape is graminoid- dominated, and the pools in which they breed likely receive allochthonous input from the surrounding grasslands. Many grassland species are in decline as this habitat has been largely converted to agricultural practices with remaining tracts being invaded by nonnative vegetation or woody plants (in Pilliod et al., 2003). The eastern tallgrass prairie and longleaf pine-wiregrass ecosystems would be particularly conducive to investigating litter quality associated with different forbs and graminoids, with the latter harboring threatened-endangered amphibian species that continue to decline, namely the flatwoods salamander (Ambystoma cingulatum and A. bishopi), striped newt (Notophthalmus perstriatus), and gopher frog (Lithobates [Rana] capito). Earl et al. (2012) provides some evidence that a shift from graminoid-dominated ecosystems to those dominated by woody vegetation may negatively impact some prairie species by impacting larval survival. Further studies should investigate the litter quality of various forbs and graminoids, particularly those in reference, restored, and degraded sites with respect to amphibian growth and development.

To reverberate research suggestions from studies mentioned in this review, efforts should be made to investigate not only single-species litter reactions, but also mixed-litter effects on amphibians. Mixtures of litter may produce synergistic or antagonistic effects in aquatic ecosystems, thus altering the effects of single species-dominated systems

(Stoler and Relyea 2011). This effort would likely require repetition across regions to fully understand the interactions between the litter quality and the amphibians that reside

19! ! there. Further, it could prove advantageous to contrast litter mixes from reference sites to those which have been altered or degraded, or may be altered through species invasions, in attempt to better predict how regional amphibian populations may react to an altered plant community. This would be applicable to both aquatic and terrestrial systems, as both the quality and recalcitrance of litter can vary, ultimately impacting the behavior and/or physiology of the resident amphibian community.

From an experimental perspective, while controlled mesocosm experiments are insightful and valuable, this simplified system may overestimate the importance of variables that are being tested (Skelly, 2002). This venue does not allow leachates and elemental nutrients to interact with the soil microbial community or other plant species.

In an effort to ameliorate the weaknesses of these experiments and make them more translatable to natural systems, future research should expand to using natural or constructed pools located in the field, as effects may vary depending on experimental venue and context. This expansion will allow for the inclusion of edaphic features that may affect leaf litter decomposition (Prescott, 2010), but also natural concentrations and ratios of litter inputs to the system.

While considerable research is focused on aquatic systems, future research should strive to incorporate plant effects on terrestrial life stages. Efforts should be made to collect more precise data with regards to the plant community and resultant litter where amphibian research is being conducted. While litter depth is a frequently recorded, significant metric (deMaynadier and Hunter, 1995, 1998), litter samples can be easily collected and analyzed and may permit a better understanding for fine-scale analyses of the amphibian community in forested ecosystems. However, litter depth and composition

20! ! can be tedious and challenging to collect, as within-site variability and measurement error may inhibit attempts to relate to species. As such, methods that attempt to record litter measurements that are reproducible and of the appropriate scale are needed.

While not directly linked to litter types, an analysis of energy budgets may prove useful in understanding species responses to novel or shifting plant communities. Energy budgets include basic physiological processes (digestion, growth, metabolism, reproduction), of which the standard metabolic rate (SMR)(metabolic rate of a resting, post-absorptive individual at a certain temperature) engrosses a large portion. With regards to the eastern red-backed salamander (Plethodon cinereus), Homyack et al.

(2010) found that SMR increased with body mass and temperature, presenting a physiological mechanism for differences that may be observed in body condition across silvicultural treatments or ecological gradients. As stated earlier in our review, leaf litter packing ratios (an artifact of lignin content, in part) affect both temperature and moisture of the leaf bed. As such, altered SMR and foraging of amphibians residing in areas of shifting plant communities experiencing differing temperature and moisture profiles resulting from different litter types may manifest into a reduced energy budget. Further, it could be scaled up to the population-level with regards to long-term impacts incurred by an increased SMR, as higher SMRs may translate to reduced energy invested into reproductive effort and subsequent fitness. The development of SMR response curves and how litter resulting from the surrounding plant community affects them could prove useful in understanding differences in body condition or fecundity across not only species, but sites, and may assist in elucidating regional declines of select species.

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Stoler, A.B., Relyea, R.A., 2013. Leaf litter quality induces morphological and developmental changes in larval amphibians. Ecology 94:1594-1603.

Stoler, A.B., Stephens, J.P., Relyea, R.A., Berven, K.A., Tiegs, S.D. 2015. Leaf litter resource quality induces morphological changes in wood frog (Lithobates sylvaticus) metamorphs. Oecologica 179:667-677.

Watling, J.I., Hickman, C.R., Lee, E., Wang, K., Orrock, J.L., 2011. Extracts of the invasive shrub Lonicera maackii increase mortality and alter behavior of amphibian larvae. Oecologica 165:153-159.

26! ! Werner, E.E., Peacor, S.D., 2003. A review of trait-mediated indirect interactions in ecological communities. Ecology 84:1083-1100.

Williams, B.K., Rittenhouse, T.A.G, Semlitsch R.D., 2008. Leaf litter input mediates tadpole performance across forest canopy treatments. Oecologia 155:377-384.

Witzell, J., Martin J.A., 2008. Phenolic metabolites in the resistance of northern forest trees to pathogens – past experiences and future prospects. Can. J. For. Res. 38:2711-2727.

27! !

Figure 1.1. Conceptual diagram illustrating the potential cascading effects of the plant community on the amphibian larval community.

28! !

Chapter 2: Does forest composition influence colonization of vernal pool-breeding amphibians?

Abstract Plants shape ecosystems, affecting both physical and chemical attributes of the landscape, and their communities are shifting worldwide as a result of multiple anthropogenic influences. Amphibians are also experiencing population shifts and declines, yet relatively few studies investigate the implications of how changing plant community can impact amphibian populations. Studies that have investigated the effect of the plant community on amphibians focus on the capacity of leaf litter to alter the larval stage, with few investigating impacts to the adult breeding community. Our study investigates the underlying drivers of colonization of novel breeding sites by amphibians with a focus on the surrounding forest community. Fourteen ponds were created along a gradient of tree communities ranging from oak to maple-dominance in 2014. We documented colonizing amphibians in 2015 and 2016, comprising 1114 captures of 12 species. Generalized linear model-based analyses performed at the community level found that the tree community was a significant predictor of amphibian colonization of isolated woodland pools. These data demonstrate that changing plant communities, as a result of altered disturbance regimes within the landscape, have the potential to influence amphibian communities. Further, these changes were documented within contemporary forests and suggest that with increased pressure due to invasion by non-native species and climate change, amphibian communities may also respond. From a conservation perspective, we suggest activities that maintain or restore landscape-level heterogeneity to facilitate conservation efforts will benefit a robust assemble of amphibians.! ! 29!

Introduction

Amphibian populations are experiencing worldwide declines with habitat loss and degradation, disease, invasive species, climate change, and the synergistic effects of these as the causative agents behind observed declines (Blaustein et al. 1994; Houlahan et al.,

2000; Collins and Storfer, 2003; Beebee and Griffiths, 2005). Amphibians represent particularly vulnerable taxa for numerous reasons: low vagility, narrow physiological

(and thus, habitat) tolerances (Houlahan et al., 2000), and complex life histories, which expose them to multiple stressors both within and across generations (deMaynadier and

Hunter, 1995; Cushman, 2006). However, while habitat loss/degradation and fragmentation represent some of the greatest threats (impacting nearly 90% of threatened amphibians), many populations of amphibians are in decline in seemingly intact habitat

(Beebee and Griffiths, 2005; Cushman, 2006; Baille et al., 2004). As such, a different lens may be necessary to elucidate direct effects behind the declines in these populations.

Further, while the continued research focusing on declines cannot be understated, an often-overlooked aspect and probable goal of amphibian research that warrants further investigation is their ability to recolonize habitat across the landscape

Habitat quality metrics are frequently assessed from a structural perspective: canopy cover, coarse woody debris, herbaceous vegetation, and leaf litter. While these represent valuable data, the overall effect of the plant community may consistently be underestimated with regards to the corresponding amphibian community-both from a terrestrial and an aquatic perspective. Plant communities have foundational impacts on ecosystems, affecting microclimate and nutrient availability, and it is likely that shifts in

30! ! the dominant plant community will affect higher trophic levels organisms (Bazzaz,

1979). Plant litter represents a resource in both terrestrial and aquatic environments that varies in both quantity and quality (Stoler and Relyea, 2013). Moreover, the prey-base for subsequent trophic levels may be directly impacted by differential litter inputs (Rubbo and Kiesecker, 2004; Stoler and Relyea, 2011). For example plant litter can exert a wide-range of effects on amphibians; from retarding to accelerating growth and development to impacting size at metamorphosis, survival, and subsequent fitness (Maerz et al., 2005a,b; Brown et al., 2006; Williams et al,, 2007; Maerz et al., 2010; Watling et al., 2011; Adams and Saenz, 2012; Cohen et al., 2012; Rogalski and Skelly, 2012; Martin and Blossey, 2013). Despite the recent abundance of research investigating the effects of litter on amphibian larvae, research investigating whether the tree community affects adult colonization/use of breeding sites has been scant.

Site colonization is an important factor in regulating population dynamics (Hanski,

1998) . The ability or willingness of amphibians to colonize new breeding sites is of particular importance as pools typically represent limited, discrete patches on the landscape, particularly in a fragmented system. Without these data, ecologists are potentially missing a valuable portion of data to help elucidate species colonization, occupancy, and population trends. Immediate colonization of novel or restored sites is limited, with many studies visiting sites several years after they were created (Lehtinen and Galatowitsch, 2001; Porej and Hetherington 2005). It is important for ecologists to know how quickly these sites are colonized, collect baseline data on the species to initially colonize them, and understand if or how this community changes over time.

31! ! To address this gap in research, we investigated if tree communities impact the colonization of novel breeding sites by the vernal pool-breeding amphibian community.

We employed a paired experiment utilizing newly created upland pools within areas of different tree communities, focused primarily on two dominant genera within the

Appalachian Plateau – oak (Quercus spp.) and maple (Acer spp.) – that are experiencing differential recruitment, which may ultimately lead to a shift in the dominant canopy community (Nowacki and Abrams, 2008). This shift may prove detrimental to pool- breeding amphibians, particularly during the larval stage, as larvae raised in an oak-litter environment may exhibit faster growth, higher survival rates, and exhibit larger sizes at metamorphosis than those raised in a maple-litter environment (Rubbo and Kiesecker,

2004). The primary hypothesis that we investigated was that the tree community would be a determinant of amphibian community colonization. If the amphibian community, or species comprising it, is preferentially utilizing pools whose surrounding area is dominated by oak, the documented shift from oak-dominated to maple-dominated canopies may impact amphibian breeding sites.

Methods

Study Site

This study was conducted at Vinton Furnace State Experimental Forest (VFSEF), a research forest of approximately 4,900-hectares located in Vinton County of southeastern

Ohio. Vinton county is located within the Central Appalachian Plateau and receives a mean precipitation of 1024 mm and has an annual mean temperature of 11.3°, with the coldest month (January) experiencing a mean temperature of -1.4° and the warmest

32! ! month (July) experiencing a mean temperature of 22.7°. Soils of VFSEF are unglaciated silt loams derived primarily from sandstones, siltstones, and shales. Dominant oak species at VFSEF are white oak (Quercus alba L.)(21% basal area), black oak (Quercus velutina Lam.) (17% basal area), chestnut oak (Quercus montana Willd.)(15% basal area), and scarlet oak (Quercus coccinea Muenchh.) (12% basal area) (Hutchinson et al.,

2008).

Experimental Design

Seven pairs (fourteen total) of 5-meter diameter pools approximately 0.75-meter in depth were constructed along a series of ridges in June 2014 using a small bulldozer provided by the Ohio Division of Forestry. Pools were not lined with any material as the sites exhibited a high enough percentage of clay in the native soil to retain water once the sites were sufficiently compacted by the bulldozer. Each pool was made to exhibit a gradual slope to the point of maximum depth. With the exception of utilizing a native grass and forb mix to initially stabilize soil and prevent excessive erosion at the point of entry for the bulldozer, no additional site manipulation occurred. Within each of the paired pools, one was constructed within a managed/primarily oak-hickory overstory (“oak pools” hereafter) and the other constructed within an unmanaged/primarily red maple/mesic species overstory and midstory (“maple pools” hereafter) (Figure 2.1). Due to the high tree diversity in the forest, no pools were within sites that entirely lacked other species

(i.e. a gradient was exhibited across the range of sites). To establish a design where individuals in the population had the option of selecting an “oak” or a “maple” pool, paired pools were constructed well within the dispersal distance of species in the community, between 50-75 meters apart. In order to achieve spatial independence with

33! ! respect to dispersal distances of the amphibian community, each pairing was at least 500 meters from other constructed pools.

Amphibian Sampling

We characterized colonization as the capture of breeding condition individuals and/or evidence of oviposition; using this characterization, all fourteen pools were colonized by

April 2015. Breeding amphibians were sampled from early March through mid-April of

2015 and 2016. We chose to survey during this period to maximize detectability for the pool-breeding amphibian species within the region as they migrated and emigrated from newly constructed pools. We sampled amphibians with aquatic funnel traps at a density of 1 trap/m2 of surface area (5-6 traps/pool) and drift fence-pitfall arrays partially enclosing each pool. Traps were deployed for 18 nights during the 2015 sampling period

(90 aquatic trap nights per pool) and 9 nights during the 2016 sampling period (45 aquatic trap nights per pool) and captured a total of 1114 amphibians, comprising 817 unique captures of twelve species (Table 2.1). Aquatic traps were set equidistant along the perimeter of each pool at a depth that did not exceed one-half the height of the trap and secured in place with a metal stake to ensure that traps remained secure and captured animals did not drown. Drift fence arrays only partially encircled each pool to ensure that colonizing amphibians were not dissuaded from utilizing pools. Each pool had three,

5-meter drift fences constructed of silt fabric arranged in a triangular pattern around each pool and were approximately 30 centimeters in height and were buried approximately 3-5 centimeters below the soil surface. Drift fences had a 5-quart bucket buried at either end that was equipped with a clear plastic funnel to prevent amphibians from escaping.

34! ! Captured amphibians were measured (snout-vent length + tail length), weighed (g), sexed, and permanently marked with visible implant elastomer (Northwest Marine

Technology, Inc., Shaw Island, WA) to identify the initial location of capture and prevent re-sampling of previously captured individuals. All sampling and handling protocols were approved by The Ohio State University Institute for Animal Care and Use

Committee (protocol # 2011A00000082-R1) and under Ohio Division of Wildlife Permit

16-104.

Habitat Sampling

Aquatic

Maximum pool depth, temperature, pH, and dissolved oxygen were recorded weekly during 2015 and 2016 field season. Conductivity was added as a metric in 2016 using a handheld DigitalAid Total Dissolved Solids/Electrical Conductivity water quality meter

(Loretta Green LLC, Clearwater, FL). Dissolved oxygen and pH were measured using the

Extech DO600 and Exstik pH meter, respectively (Extech Instruments Corporation,

Nashua, NH). Further, invertebrate biomass and richness were recorded via area- constrained dip netting with a net equipped with a 27 cm2 bag during amphibian sampling occasions. Five sweeps approximately 1.5 meters in length were implemented from near the center of the pool to the shoreline, while repeatedly jabbing the net into the substrate.

Invertebrates were identified to order or class and categorized as either “predator” or

“prey” with respect to amphibian larvae (Merritt and Cummins, 1996). Invertebrate biomass was recorded by taking the wet biomass to the nearest tenth of a gram of the collected samples.

35! ! Terrestrial

Habitat surrounding each pool was sampled within a 30-m radius plot. Four 30-meter transects were established in the four cardinal directions from the center of each pool, with each transect having two 10x10-meter quadrats established on alternate sides, one from the 0-10 meter portion of the transect (beginning at the edge of each pool, immediately beyond the area disturbed by the bulldozer) and the second from the 21-30 meter portion of the transect. Within the 10x10-m quadrat, all trees greater than 2.5 cm diameter-at-breast height (DBH) were recorded and the percentage of bare ground, forb, and litter coverage recorded. Within each 10x10-m quadrat, a secondary 5x5 meter quadrat was established in the northeastern corner where saplings less than 2.5 cm and shrubs were recorded. The Importance Value Index (IVI) is the summation of relative density + relative frequency (number of plots a species occurred in) + relative dominance for each species with individuals exhibiting a DBH ≥2.5 cm (Curtis and McIntosh, 1951).

For statistical analyses, the IVI for species from all eight quadrats surrounding each pond were summed and species IVI were pooled into functional groups (i.e. oak, hickory, maple, other, oak+hickory, maple+mesic species). Coarse woody debris (CWD) was quantified according to the line-intersect method (Waddell, 2002), in which the 30-meter transects that were positioned and oriented in each cardinal direction perpendicular to the pond border (Warren and Olsen, 1964). For this method, CWD with a diameter ≥7.5 cm at the point of intersection was measured for total length and had diameter recorded at each end (DiMauro and Hunter, 2002). We used these measurements to calculate cubic volume of CWD within the immediate (30-m radius) vicinity of each pool. Canopy openness was recorded from a stationary point immediately on the shore of each pool

36! ! over three time periods (March (pre-leaf out), May (post-leaf out), and June (full-leaf out)) using a Pentax K30 DSLR equipped with a 180 hemispherical (fisheye) lens.

Photographs were analyzed using Gap Light Analyzer (GLA), Version 2.0 (Fraser et al.,

1999).

Statistical Methods

Data were summarized prior to statistical analyses to identify any deviations from normality that may impact analyses. Based upon our primary hypotheses involving the composition of the tree community, we determined that our pools represent a gradient of tree communities rather than two discrete treatments based upon the IVI of our functional groups, and that our analyses should reflect this situation (Figure 2.2). This was further reinforced by exploratory analyses, which did not detect a treatment effect for our data.

To examine variation in species composition along this gradient, we used multivariate generalized linear models with environmental parameters as predictor variables.

Multivariate generalized linear models were chosen over more traditional distance-based methods (i.e. correspondence analyses, multidimensional scaling) as they are more robust and do not confound location with dispersion effects, potentially inflating Type I and

Type II errors (Warton et al., 2012). We conducted all statistical analyses with RStudio

Version 0.99.467 (RStudio Team, 2015) using the packages mvabund and vegan (Wang et al., 2012). We tested for multi-collinearity using a correlation matrix between all explanatory variables, eliminating those variables from models with an r > 0.70 (Berry and Feldman, 1985). “Year” and “treatment” were specified as objects in our models.

Explanatory variables were modified in RStudio using the scale function as they occurred across different scales of varying magnitudes. From the remaining variables, we created

37! ! one global model, three landscape-scale models, three pool-scale models, and five candidate model sets, specifying negative binomial distribution to account for overdispersion of data. Negative binomial distribution was chosen over Poisson distribution, as it is more flexible and appropriate for our count data (O’Hara and Kotze,

2010). Residuals vs. fitted plots were checked to ensure that model assumptions were met. Significance was assessed using 1000 permutations of PIT-trap resampling, a new method which bootstraps probability integral transform residuals and gives the most reliable Type I error rates (Warton and Wang, in review). Models were then evaluated using Akaike Information Criteria (AIC) methods and the most appropriate models selected for further assessment (Burnham and Anderson, 2002). As the best models are those with the lowest AIC values, we utilized Akaike weights to assess the likelihood of our other high-ranking models. To assess the importance of individual variables within our models, we used the function drop1 to evaluate how AIC values were altered when one of the predictor variables was eliminated from the model. To complement the utilization of mvabund, we conducted an indirect gradient analysis using non-metric multidimensional scaling (NMDS) on our site-by-species matrix using the vegan package. Ordinations based on Bray-Curtis dissimilarity were conducted using the metaMDS function in vegan. Since our sites were better visualized as a gradient, we overlaid our original community matrix with our “oak importance value index” scores onto the NMDS plot using ordisurf.

38! ! Results

All constructed pools filled with snowmelt and rainwater by December of 2014 and exhibited a permanent hydroperiod with a mean maximum depth of 66±16 centimeters.

Twelve species of amphibians were identified utilizing our constructed pools, with a mean of 10.1±1.5 species per pool (range 7-12 species). Spotted and Jefferson salamanders (Ambystoma maculatum and A. jeffersonianum, respectively) were the most commonly encountered species, comprising over 42% of our total captures and 39.5% of our unique captures. Four species of amphibians (marbled salamander (A. opacum), bullfrog (Lithobates catesbieanus), pickerel frogs (L. palustris), and gray treefrogs (Hyla versicolor)) were poorly represented in our samples, either due to incongruous breeding seasons (marbled salamander, bullfrog, gray treefrog) or habitat (pickerel frog) and were omitted from statistical analyses.

Of our twelve models, no landscape- or pool-scale models were competitive against the mixed-scale candidate models. We did not detect a “treatment” effect, further supporting the decision to analyze our tree community as a gradient as opposed to rank orders. The model with the most support for explaining our amphibian community was a mixed-scale candidate model “year+oakIV+pH+canopy+leaf” with an AIC weight of

0.75 (Table 2.2). This model consisted of year (2015+2016), the oak importance value index (oakIV), the initial pH of the pools upon spring thaw, the percent canopy openness, and the percent of area covered by leaf litter surrounding pools. All predictor variables

(year, oakIV, intpH, canopy openness, and leaf litter coverage) were significant in this model, with P<0.05. In order to further assess the parsimony of the model we utilized the drop1 command, which reports the change in AIC value for each of the variables that are

39! ! dropped. These results showed that our AIC score increased with the omission of any variable in our top model. The second highest-ranking models had ΔAIC value >2 points with the omission of canopy openness, and the final three models reported all have ΔAIC values >10 than our top ranking model and are also reported. In order to better visualize our data, a NMDS plot utilizing contour lines to depict an environmental gradient, the oak importance value index, against our species matrix gave similar results to our mvabund (model-based analysis) and corroborated that the oak importance value index is indeed influential in shaping our amphibian community (Figure 2.3).

Not all species responded positively to our top ranking model; while most species were positively associated with increasing year, mountain chorus frogs and spring peepers frogs were negatively associated with increasing year, a result of fewer individuals being captured in 2016 (Table 2.3). Further, there was a negative association with increasing oak importance value with all species with the exception of spring peepers and American toads, which were positively associated with an increasing oak importance value. Canopy openness was positively was associated with all species, with the surprising exceptions of green frogs and American toads. Increased percentage coverage of leaf litter surrounding pools was positively associated with all species except for American toads and spring peepers. Higher initial pH values were positively associated with most of the species in the community with the exception of four species: four-toed salamanders, mountain chorus frogs, spring peepers, and wood frogs.

As “year” was a significant variable in every predictive model, data was reanalyzed by pooling amphibian abundances from both years to determine the most supported model and predictor variables in the without a temporal component. Our top

40! ! model (year+ oakIV+ intpH+ canopy openness+ leaf litter coverage) was further strengthened with the omission of year, with an AIC weight of 1.00. The subsequent models all exhibited a ΔAIC of >10 and were the original candidate models with the exception of Model2, which was replaced by our Global model. The oak importance value decreased the AIC value for each of the top ranking models (Table 2.4). The top- ranked model consisting of “oakIV + intpH + canopy openness + leaf litter coverage” had similar results to our models that incorporated year, with most species showing a negative association with an increasing oak importance value, with the exceptions of the mountain chorus frog, spring peepers, and American toads (Table 2.5). The initial pH coefficients differed slightly, with the spotted salamander, four-toed salamander, eastern newt, spring peeper, and wood frog showing a negative association with the initial pH. The percent of area surrounding the pools that exhibited leaf litter coverage was positively associated with all species except for American toads and spring peepers.

Discussion

The results of this study provide evidence that tree community composition can influence the utilization and/or initial colonization of sites by amphibians. Though there have been several studies documenting the colonization of created wetlands by amphibians, this is the first to be designed to investigate colonization as it relates to the tree community

(Lehtinen and Galatowitsch, 2001; Rannap et al. 2009, Lesbarrères et al. 2010). It also provides field-based support for several other environmental variables that are consistently found to be significant in modeling local amphibian communities: pH of

41! ! breeding site, leaf litter coverage, canopy openness, and presence of predatory invertebrates (Dunson et al. 1992; Relyea 2001; Skelly 2005; Semlitsch et al., 2009).

Although the primary objective of this study was to evaluate the effect of the tree community on the initial colonization of novel breeding sites, it is necessary to relate these findings to those focused on larvae. The negative association of the adult amphibian community to an increasing oak IVI seems in conflict with the many mesocosm-based leaf litter trials in which larvae exhibited a significant positive effect from increased oak litter (Rubbo and Kiesecker 2004). With the omission of “year” as a variable, wood frogs (-1.03) and four-toed salamanders (-1.90) exhibited the strongest negative association with an increasing oak IVI, whereas the spring peeper (0.18) and mountain chorus frog (0.34) and American toads (0.23) showed a positive coefficient to an increasing oak IVI. The guilds formed by the responses to an increasing oak importance value – whether negative or positive – correspond well with the successional stage is typically described as being optimal habitat for these species. Four-toed salamanders, wood frogs, and ambystomatids are frequently identified as being associated with later-successional beech-maple forests, whereas spring peepers and

American toads are frequently described as using more early successional habitats, such as those associated with oak-hickory and old-field habitat (Petranka 1998; Skelly et al.,

2005)

It is likely that there are differences in adult colonization and larval performance in these systems, as factors that may hinder larval growth may inhibit adult and metamorph performance. For example, adult amphibians frequently select for sites that minimize desiccation, which may include closed canopy sites that limit sunlight

42! ! penetration to the forest floor and plant communities that help maintain cool, mesic conditions (i.e. red maple) when larval development in closed canopy vernal pools typically exhibit retarded growth and development from the decreased temperature and primary productivity (Rubbo and Kiesecker, 2004; Skelly et al., 2005; Semlitsch et al.,

2009). In concert with this study, we monitoring the growth, development, and size at metamorphosis of the larval communities in these pools. It is imperative to stress that our pools were constructed in forested setting, and as such the tree community, and thus the resulting litter, represents a gradient of species abundances whereas most larvae-litter studies occur in mesocosms (Cohen et al. 2012, Martin and Blossey 2013, Martin et al.

2015, Stoler et al. 2015). From an experimental perspective, while the controlled mesocosm experiments from which much litter research is based are insightful and valuable to mechanisms that can drive species performance, the controlled system may limited predictive responses under natural conditions (Skelly, 2002). Regardless, this research provides further support that amphibians do react to the plant community, albeit from a different life stage.

The initial pH (mean: 6.49 ± 0.39, range: 5.32 – 7.26) of our pools was consistently among our most predictive environmental variables, with several species being positively associated with a higher pH. This finding is consistent Dunson et al.

(1992) where many pool-breeding amphibians (spotted salamanders, Jefferson salamanders, wood frogs) exhibited lower hatching success and survival in lower pH environments. However, the four-toed salamander, a species that frequently inhabits acidic environments (i.e. sphagnum bogs) (Petranka, 1998; Wahl et al., 2008),exhibited the only negative coefficient (-0.53) of our salamander species. As four-toed

43! ! salamanders often occur sympatrically with ambystomatids, it is possible that the tolerance, or preference, of this species to lower pH sites may be in part a result of predatory exclusion by the larger, predatory ambystomatids or selection for sites less favorable for the larvae of these species, allowing four-toed salamanders to occupy niches not fully exploited by ambystomatids.

As an alternative to leaf litter depth, percent leaf litter coverage, was used in this study. Again, with “year” removed a predictor, our top models all included percent leaf little coverage as a significant predictor, with all species except for three (American toad spring peeper, and green frog) exhibiting a positive coefficient. Four-toed salamanders

(1.53), eastern newts (1.24) and wood frogs (1.01) exhibited the strongest positive coefficient to increasing leaf litter coverage, findings consistent with numerous other studies regarding the behavior and natural history of these and other amphibian species that select for forested environments where leaf litter serves as important refugia

(Petranka 1998; Semlitsch, 2002; Baldwin et al., 2006; Rittenhouse and Semlitsch, 2006).

American toads (-0.61) exhibited the strongest negative association with increasing leaf litter, instead being more common in pools with higher degrees of bare ground.

Interestingly, numerous species within Anaxyrus show not only little aversion to the bare ground typical of highly disturbed sites (i.e. areas of timber harvest, fire, scouring floods), but seem to select for, and in some cases require, these sites (Graeter, 2005; Todd and Rothermel, 2006; Guscio et al., 2007). The proclivity of toads to rapidly colonize newly realized sites as a result of their high vagility and physiological ability to move through disturbed areas where they have to traverse tracts of bare soil is well-

44! ! documented, and in some cases it is suggested that the presence of these sites is necessary for population stability (Todd and Rothermel, 2006; Guscio et al., 2008).

Canopy openness (or canopy cover) ranks consistently as being one of the most important determining factors in shaping the amphibian community (Skelly et al., 2005;

Williams et al., 2007). Canopy cover determines exposure to light (and thus, primary productivity) as well as a suite of other associated characteristics including temperature, moisture (in terrestrial environments), and dissolved oxygen (in aquatic environments).

From the short-duration of our study, canopy openness was present in our top models, further suggesting its importance in structuring amphibian communities. In support of many studies, canopy openness was positively associated with adult colonization of these sites for most species, although it was negatively associated with two species that we expected to see positive coefficients (American toads and mountain chorus frogs)

(Guscio et al., 2007; Semlitsch et al. 2009). It is likely that our created pools do not span a wide enough gradient of canopy openness to permit comprehensive analysis of this variable to permit further understanding of the response by these two species, or that unaccounted for covariates are exerting this irregular reaction. Successive years and continued sampling may find the coefficients of this environmental variable shifting in our models. Many of the environmental variables that we utilized, particularly the oak

IVI, leaf litter coverage, and canopy openness, are associated with early successional stages. While we cannot be certain that other traits are not further driving the amphibian community to colonize certain pools, the environmental conditions associated with certain tree communities, if not the tree community itself, is a significant variable in modeling colonization by amphibians on the landscape Other variables that are

45! ! consistently ranked as being significant with regards to the amphibian community include wetland size/depth, presence of fish, hydrology, and structure; however, these variables were controlled for in that all pools were of the same size and depth, lacked fish, had similar structure with regards to slope and vegetation, and exhibited a permanent hydroperiod during our study (Porej and Hetherington 2005; Denton and Richter 2013).

Our most supported models were neither at the landscape- nor pool-scale, but instead at the mixed-scale. This supports what is well known about amphibians: a complex life history that is dependent upon both the larval (breeding) habitat and adult

(nonbreeding) habitat (Semlitsch 2002). Neglecting to restore or maintain both of these habitats for amphibians is insufficient for conservation purposes as species are interacting with both of these environments and at different scales (Semlitsch 2002; Skelly et al.

2005; Roznik and Johnson 2009). Moreover, this underscores the necessity to continue research that investigating the habitat use and selection for pool-breeding amphibians; if we lack a clear understanding of what dictates optimal habitat for amphibians (which may vary across the range of a species), we will be limited in our ability to conserve them. This study underscores the necessity to maintain landscape-level heterogeneity in order to maintain biodiversity, as the amphibian response to the plant community is species-dependent. With the widespread shift in tree communities through much of the

Appalachian region as a result of differential regeneration resulting from lack of disturbance, the shift from mixed oak to maple forests may have landscape-level consequences (Abrams, 1998). Specifically, in the absence of fire a variety of mesophytic, highly competitive, late-successional plant species (i.e. red maple (Acer rubrum), American beech (Fagus grandifolia), and blackgum (Nyssa sylvatica)) are

46! ! replacing shade-intolerant oaks (Quercus spp.) and hickories (Carya spp.) (Nowacki and

Abrams 2008). The result of this replacement is an altered understory microclimate characterized by heavy shade and increased humidity - effectively a cooler and moister understory and forest floor (Nauertz et al., 2004). Moreover, this altered understory produces a seedbed more conducive to mesophytic plant species and creates a positive feedback loop for this cycle, and thus the homogenization of the functional plant community (Nowacki and Abrams, 2008). Should these tree community shifts result in differential use of breeding habitat by amphibians, not only could fine-scale habitat degradation be observed with respect to adults, but also subsequent reproductive output and fitness of larvae. Temporary or isolated wetlands may be a limiting factor for amphibian populations in this landscape, and the lack of pools within certain contexts

(e.g. tree community, canopy openness, pH) may even preclude population viability for a number species, particularly early successional species, namely American toads, mountain chorus frogs, and spring peepers, as these species may depend on disturbance to generate habitat heterogeneity, such as that found in actively managed forests (Roznik and Johnson 2009).

The utilization of constructed pools has become common method in amphibian conservation, but care must be taken to ensure that constructed pools are capable of replacing or replicating the characteristics of vernal pools, namely the hydroperiod

(Lesbarreres et al., 2010; Calhoun et al. 2014). While our pools are currently exhibiting a permanent hydroperiod, they remain shallow (66±16 centimeters) and may eventually develop a semi-permanent hydrology. Although pools exhibiting a permanent hydroperiod should be discouraged (as they encourage a homogenization of the aquatic

47! ! community atypical of vernal pools) it is likely less detrimental to have a pool with a hydroperiod that extends too long than one that consistently fails to hold water for a sufficient enough period for amphibians to complete metamorphosis, thus creating an ecological trap (Calhoun et al., 2014). Regardless of the hydroperiod, the pools created in this study provided encouraging results in the short-term, as every species of pool- breeding amphibian in the region had colonized them within two years. While long-term monitoring may be necessary to determine their functionality, Petranka et al. (2003) suggests that 2-3 years of monitoring may be sufficient to characterize the community that will utilize the pools for the first decade or more (Petranka et al. 2007; Rannap et al.,

2009). Further, the size of these created pools is substantially smaller than most other wetlands created for amphibian use, with our pools having a mean diameter of 5-meters, yet still exhibiting a mean species richness of 10 species (Snodgrass et al., 2000; Lehtinen and Galatowitsch, 2001; Pechmann et al., 2001; Petranka et al., 2006; Denton and

Richter, 2013). If pools of this small size are capable of maintaining or promoting viable populations of amphibians, the result would be a much more feasible tactic for enhancing local amphibian populations.

References

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53! !

Figure 2.1. Topographic map for a section of Vinton Furnace State Experimental Forest in Vinton County, Ohio with locations of pool pairings constructed in 2014. Alphanumeric codes correspond to treatment (O=oak, M=maple) and pair number (ex. O1 = oak 1).

54! !

2.2. Distribution of constructed pools at Vinton Furnace State Experimental Forest across gradient of tree communities based on importance value index (IVI).

55! !

Figure 2.3. Non-metric multidimensional scaling (NMDS) ordination (stress = 0.133) using contour lines to depict the gradient of the oak importance value index on the landscape with respect to the amphibian community. Circles represent pools as they occur on the landscape, and four-letter abbreviations are code using the first two letters of the genus+species where multiple genera are present, or the first four letters of the genus where only one species is present. *Spring peeper, American toad, four-toed salamander, and eastern newt beyond margins of plot.

56! ! Table 2.1. Total amphibian captures in constructed pools for 2015-2016. *Denotes species omitted from statistical analyses. Numbers in () indicate unique captures for each species.

Family Species Common Name Total Bufonidae Anaxyrus americanus American toad 70 (62) Hylidae Pseudacris brachyphona mountain chorus frog 66 (57) P. crucifer spring peeper 66 (65) Ranidae Lithobates catesbeianus bullfrog *10 (10) L. clamitans green frog 162 (118) L. palustris pickerel frog *4 (4) L. sylvaticus wood frog 96 (83) Ambystomatidae Ambystoma jeffersonianum Jefferson salamander 206 (140) A. maculatum spotted salamander 267 (183) A. opacum marbled salamander *3 (3) Plethodontidae Hemidactylium scutatum four-toed salamander 40 (40) Salamandridae Notophthalmus viridescens eastern newt 124 (52)

57! ! Table 2.2. Candidate models ranked by relative support for amphibian community in constructed pools with Akaike Information Criteria (AIC) scores, Δ AIC, and Akaike parameter weights (ωρ)* Indicates significant variables (p≤0.05).

Model Parameters AIC ΔAIC ωρ drop1 AIC Can4 year*+oakIV*+intpH*+canopy*+leaf* 1019.05 0 0.75 year: 1078.8 oakIV: 1029.5 intpH: 1035.8 canopy: 1021.2 leaf: 1039.3

Can1 year*+oakIV+intpH*+leaf* 1021.23 2.2 0.25 year: 1097.1 oakIV: 1025.0 intpH: 1051.8 leaf: 1032.8

Can3 year*+oakIV*+canopy*+leaf* 1035.82 16.8 0.00 Year: 1103.1 oakIV: 1069.0 canopy: 1051.8 leaf: 1051.4

Can2 year*+intpH* +canopy*+invert 1038.81 19.8 0.00 year: 1114.0 intpH: 1079.0 canopy: 1044.6 invert: 1037.7

Can5 year*+oakIV+intpH*+canopy 1039.27 20.2 0.00 year: 1098.7 oakIV: 1037.7 intpH: 1051.4 canopy: 1032.8

58! ! Table 2.3. Coefficients of species comprising amphibian community in fourteen constructed pools to associated variables in top-ranked model in Vinton Furnace State Experimental Forest for 2015-2016.

Year OakIV Canopy Leaf Initial pH Open Cover

Ambystoma 0.24 -0.65 0.38 0.313 0.19 jeffersonianum Jefferson salamander

Ambystoma maculatum 2.09 -0.53 0.64 0.52 0.29 spotted salamander

Hemidactylium scutatum 15.20 -0.67 0.30 0.42 -0.53 four-toed salamander

Notophthalmus 5.03 -0.20 0.22 1.15 1.25 viridescens eastern newt

Anaxyrus americanus 0.70 0.34 -1.23 -1.25 1.02 American toad

Pseudacris brachyphona -2.78 -0.35 0.89 0.96 -1.31 mountain chorus frog

Pseudacris crucifer -0.24 0.14 0.54 -0.01 -0.42 spring peeper

Lithobates sylvaticus 1.15 -0.80 0.86 0.88 -0.37 wood frog

Lithobates clamitans 0.06 -0.05 -0.31 -0.18 0.71 green frog

! ! !

59! ! Table 2.4. Alternate models for colonizing amphibian community in constructed pools with “year” omitted from explanatory variables with Akaike Information Criteria (AIC) scores, Δ AIC, and Akaike parameter weights (ωρ) * Indicates significant variables (p≤0.05).

Model Parameters AIC ΔAIC ωρ drop1 ΔAIC Can4 oakIV*+intpH*+canopy*+leaf* 1078.77 0 1.0 oakIV: 1107.9 intpH: 1103.1 leaf: 1098.7 canopy:1097.1

Can1 oakIV*+intpH*+leaf 1097.10 18.33 0.0 oakIV: 1103.4 intpH: 1105.7 leaf: 1094.0

Can5 oakIV+intpH*+canopy 1098.73 19.96 0.0 oakIV: 1108.1 intpH: 1106.6 canopy:1094.0

Global oakIV+cwd+intpH*+invert+mxslope+canopy 1110.23 21.46 0.0 oakIV: 1090.6 cwd: 1099.5 intpH: 1120.3 invert: 1108.6 mxslope:1107.8 canopy: 1098.9

Can3 oakIV*+canopy+leaf* 1103.08 24.24 0.0 oakiv: 1124.2 leaf: 1106.6 canopy:1105.7

60! ! Table 2.5. Coefficients of species comprising amphibian community in fourteen constructed pools to associated variables in top-ranked model in Vinton Furnace State Experimental Forest with the omission of “year” as a variable for 2015-2016.

OakIV Canopy Leaf Initial Open Cover pH

Ambystoma jeffersonianum -0.69 0.46 0.36 0.09 Jefferson salamander

Ambystoma maculatum -0.86 1.17 0.91 -0.55 spotted salamander

Hemidactylium scutatum -1.90 1.62 1.53 -0.56 four-toed salamander

Notophthalmus viridescens -0.74 1.18 1.24 -0.02 eastern newt

Anaxyrus americanus 0.23 -1.00 -1.12 0.69 American toad

Pseudacris brachyphona 0.34 -0.18 0.49 0.12 mountain chorus frog

Pseudacris crucifer 0.18 0.47 -0.05 -0.33 spring peeper

Lithobates sylvaticus -1.03 1.17 1.01 -0.84 wood frog

Lithobates clamitans -0.06 -029 -0.17 0.68 green frog

61! ! Chapter 3: Responses of Larval Amphibian Communities to Variations in Litter Input Across a Naturally Existing Oak-Maple Gradient

Abstract Plant litter has a profound impact on amphibian larvae, from morphology, growth and development, to survival and size at metamorphosis. Most studies investigating the effects of litter utilize mesocosms, which may exaggerate ecological relationships due to a simplified food web. It is imperative to determine if these effects carry over into the field and natural systems. In order to scale up the research whose foundation was pioneered in mesocosm, we constructed fourteen woodland pools across a naturally occurring gradient of oak-dominance to maple-dominance, and collected data on the larval community. We modeled growth rate, community structure, and size/mass at metamorphosis of four amphibian species to naturally colonize the pools: American toads

(Anaxyrus americanus), chorus frogs (Pseudacris crucifer + P. brachyphona), wood frogs (Lithobates sylvaticus), and Jefferson salamanders (Ambystoma jeffersonianum).

Allochthonous input derived from oak and hickory exerted a significant positive effect on size and mass at metamorphosis for all species, although the effects were not as distinct as those observed in mesocosm studies. Increasing input of oak litter also was a significant predictor in modeling the larval community. Our study supports findings from mesocosm experiments and suggests that shifts in plant communities as a result of climate change, lack of disturbance, and invasive species may carry over and impact the amphibian community.

! 62! Introduction

Understanding the relationship of species to ecosystems comprises the core of ecological study (Lawton 1994). Some species exert such strong effects in the ecosystem that they occur as to designate them “foundational” species. Many tree species are considered to be foundational species because of the effect that they exert on their environment, such as hemlocks (Tsuga spp.) (Mills et al. 1993). As such, the loss or shift in dominance of some species that exert these disproportionate effects on the ecosystem may impact numerous species through altered habitat, resource availability, nutrient cycling, and synergistic effects of these and other stressors (Mills et al. 1993). One such situation is occurring in eastern North America via mesophication, whereby fire-tolerant, shade-intolerant species such as oak (Quercus spp.) and hickory (Carya spp.) are being replaced by fire-intolerant, shade-tolerant species, such as red maple (Acer rubrum), blackgum (Nyssa sylvatica) and

American beech (Fagus grandifolia), an occurrence referred to as the “red maple paradox” (Abrams 1998). With the recognition of this shift, numerous studies have been conducted investigating the potential impact of a landscape-level shift from oak- dominated canopies to red maple-dominated canopies (Nowacki and Abrams 2008).

The effects of a community shift from oak to red maple can alter numerous characteristics of the landscape; from temperature and moisture, to physiochemical characteristics of soil and water, to nutrient and subsequent prey availability to other organisms, such as larval amphibians (Russell et al. 1999; Nowacki and Abrams 2008).

Allochthonous litter provides much of the energy in isolated bodies of water in forest ecosystems - habitats crucial to sustaining many amphibian species (Martin and Blossey

2013). The leaf litter originating from oak and maple varies in the ratio of key

! 63! components: carbon:nitrogen:phosphorus (C:N:P) ratio, secondary metabolites, and lignin - characteristics that can drastically alter the aquatic environment (Rubbo and

Kiesecker 2004; Maerz et al. 2005; Stoler and Relyea 2011; Martin and Blossey 2013;

Saenz et al. 2013). Amphibian larvae in these ponds respond to the quality and quantity of litter through the alteration of traits that can impact subsequent fitness, including: growth, development, size at, and timing of metamorphosis (Rubbo and Kiesecker 2004;

Maerz et al. 2005; Brown et al. 2006; Williams et al. 2008; Cohen et al. 2012; Earl and

Semlitsch 2012). For example, Rubbo and Kiesecker (2004) found that wood frogs

(Lithobates [Rana] sylvaticus) raised in mesocosms containing oak litter exhibited faster growth, higher survival, and larger size at metamorphosis than those raised in red maple litter. Additional studies also underscore the importance of leaf litter in the aquatic environment, though results seem to indicate strong species-specific effects (Maerz et al.

2005, 2010; Brown et al. 2006; Williams et al. 2008; Watling et al. 2011; Cohen et al.,

2012; Cotten et al. 2012; Earl et al. 2012). Oak litter exhibits a high C:N:P ratio, high amounts of lignin, and high amounts of tannins. While these traits may initially suggest a poorer nutrient source resulting from lower quality, this lower quality coupled with high amounts of lignin (which both physically and chemically inhibits decomposition) produce a sustained food source, capable of hosting various organisms upon which others feed

(Melillo et al. 1982; Leroy and Marks 2006). In contrast, red maple litter exhibits a lower

C:N:P ratio and low quantities of lignin, resulting in a rapidly decaying litter, as it is quickly colonized and broken down by fungi and bacteria, resulting in a pulsed resource

(Rubbo and Kiesecker 2004).

! 64! Most of the studies investigating the importance of the litter on amphibian larvae have utilized mesocosms for their research (Rubbo and Kiesecker, 2004; Brown et al.

2006; Maerz et al. 2010; Martin and Blossey, 2013; Martin et al. 2015). While mesocosms are valuable tools for elucidating underlying drivers in ecosystems, the inherent controlling of exogenous factors may overestimate ecological strengths and relationships (Skelly 2002; Melvin and Houlahan, 2012). It is important that the information gained from these studies be applied in field settings to ascertain how these relationships transfer across space. In this study, we created fourteen isolated pools across a gradient of canopy communities, from oak-dominance to maple-dominance and documented larval growth and development from naturally colonizing amphibians. This design allowed us to test multiple hypotheses attributed to larval growth, including litter type, density of larvae, and canopy openness, in a naturally occurring system. Based on results from mesocosm studies, we hypothesized that the tree community, density, and canopy openness would be significant predictors of larval growth, community structure, and size at metamorphosis, with higher amounts of oak litter exerting positive effects, higher densities exerting negative effects, and greater canopy openness exerting positive effects.

Methods

Study Site

This study was conducted at Vinton Furnace State Experimental Forest (VFSEF), a research forest of approximately 4,900-hectares located in Vinton County of southeastern

Ohio. Vinton county is located within the Central Appalachian Plateau and receives a

! 65! mean precipitation of 1024 mm and has an annual mean temperature of 11.3°, with the coldest month (January) experiencing a mean temperature of -1.4° and the warmest month (July) experiencing a mean temperature of 22.7°. Soils of VFSEF are unglaciated silt loams derived primarily from sandstones, siltstones, and shales. Dominant oak species at VFSEF are white oak (Quercus alba L.)(21% basal area), black oak (Quercus velutina Lam.) (17% basal area), chestnut oak (Quercus montana Willd.)(15% basal area), and scarlet oak (Quercus coccinea Muenchh.) (12% basal area) (Hutchinson et al.

2008).

Experimental Design

Seven pairs (fourteen total) of 5-meter diameter pools approximately 0.75-meter in depth were constructed along a series of ridges in June 2014 using a small bulldozer provided by the Ohio Division of Forestry (Figure 3.1). Pools were not lined with any material as the sites exhibited a high enough percentage of clay in the native soil to retain water once the sites were sufficiently compacted by the bulldozer. Each pool was made to exhibit a gradual slope to the point of maximum depth. With the exception of utilizing a native grass and forb mix to initially stabilize soil and prevent excessive erosion at the point of entry for the bulldozer, no additional site manipulation occurred. Within each of the paired pools, one was constructed within a managed/primarily oak-hickory overstory

(“oak pools” hereafter) and the other constructed within an unmanaged/primarily red maple/mesic species overstory (“maple pools” hereafter). Due to the high tree diversity in the forest, no pools were within sites that entirely lacked other species (i.e. a gradient was exhibited across the range of sites). An importance value index (IVI) was determined for our focal tree species to assign an oak, oak + hickory, red maple, and red maple +

! 66! mesic species (i.e. blackgum, American beech) importance value index to each pool

(Brower et al. 1998). Paired pools were constructed between 50-75 meters apart, and each pairing more than 500 meters from the other constructed pools to achieve spatial independence relative to each pair. It was important to construct paired pools close enough such that individuals of the same populations had the option of selecting an “oak” or a “maple” pool.

Larval Sampling

The constructed pools were naturally colonized by the local pool-breeding amphibian community (comprising 10 species) within the first year and their larvae provided the specimens for this study. Larvae were sampled during four periods from late April through late June (2015) and early May through mid-June (2016), with each sampling period being between 10-16 days apart. We chose to survey during this period to maximize detectability for vernal pool obligates, namely the mole salamanders

(Ambystoma spp.) and wood frogs. Amphibian larvae were sampled by dip netting using a net equipped with a 27 cm2 bag via area-constrained dip netting, with one sweep per square meter of surface area, totaling 5 to 6 sweeps per pool. Each sweep was approximately 1.5 meters in length and initiated from near the center of the pool to the shoreline, while repeatedly jabbing the net into the substrate. Larvae were transferred to sorting pans where they were identified to species, counted, measured (mm) (snout-vent length + tail length), weighed (nearest tenth of a gram), and staged according to Gosner

(Gosner 1960). Due to the difficulty in confidently and consistently differentiating between spring peepers (Pseudacris crucifer) and mountain chorus frogs (P. brachyphona), all Pseudacris spp. larvae (and subsequently metamorphosed individuals)

! 67! were grouped together for analyses. All larvae were released after processing to their pool or origin. Metamorphosed individuals were collected from drift fence arrays and dip netting pool edges. Each pool had three, 5-meter drift fences constructed of silt fabric arranged in a triangular pattern around each pool and were approximately 30 centimeters in height and were buried approximately 3-5 centimeters below the soil surface. Drift fences had a 5-quart bucket buried at either end that was equipped with a clear plastic funnel to prevent metamorphosed amphibians from escaping. Captured metamorphosed amphibians were measured (mm) (snout-vent length + tail length), weighed (nearest tenth of a gram) and released on the opposite side of the drift fence array. All sampling and handling protocols were approved by The Ohio State University Institute for Animal Care and Use Committee (protocol # 2011A00000082-R1) and under Ohio Division of

Wildlife Permit 16-104.

Habitat Sampling

Maximum pool depth, temperature, pH, and dissolved oxygen were recorded weekly during 2015 and 2016 field season. Conductivity was added as a metric in 2016 using a handheld DigitalAid Total Dissolved Solids/Electrical Conductivity water quality meter

(Loretta Green LLC, Clearwater, FL). Dissolved oxygen and pH were measured using the

Extech DO600 and Exstik pH meter, respectively (Extech Instruments Corporation,

Nashua, NH). Canopy openness was recorded from a stationary point immediately on the shore of each pool over three time periods (March (pre-leaf out), May (post-leaf out), and

June (full-leaf out)) using a Pentax K30 DSLR equipped with a 180 hemispherical

(fisheye) lens. Photographs were analyzed using Gap Light Analyzer (GLA), Version 2.0

(Fraser et al. 1999). Allochthonous input was quantified by utilizing one litter basket per

! 68! pool (28 cm high x 44 cm wide x 61 cm long) with a section of Styrofoam (4 cm high x

50 cm wide x 70 cm long) attached beneath to allow the litter basket to float in the water.

Floating litter baskets were then tethered to the shore such that they stayed within the central portion of the pool and were deployed before senescence (early September) and collected after senescence (December) (Figure 3.2). Litter baskets were effective at representing the litter of the surrounding oak community based on the correlation of oak

2 IVI to the oak litter captured in the baskets (r adj. = 0.85)(Figure 3.3) and as such were treated as viable options for modeling litterfall. Litter samples were sorted according to species and a percentage of total litter determined for each guild :(1) oak, (2) oak + hickory, (3) red maple, (4) red maple + mesic species. Invertebrate biomass and richness were recorded via area-constrained dip netting with a net equipped with a 27cm2 bag during every amphibian sampling event. One sweep approximately 1.5 meters in length was initiated from near the center of the pool to the shoreline, while repeatedly jabbing the net into the substrate for every square meter of surface area. Invertebrates were identified to Order or Class and categorized as either “predator” or “prey” with respect to amphibian larvae (Merritt and Cummins 1996). Invertebrate biomass (to nearest tenth of a gram) was recorded by taking the wet biomass of the collected samples after having been patted dry to remove excess water. All collected invertebrates were subsequently preserved in 70% ethanol.

Statistical Methods

We conducted all statistical analyses with RStudio Version 1.0.136 (RStudio Team 2015) using the packages lme4 and MuMIn for mixed effect modeling and the package mvabund and vegan for modeling the larval community. Based upon our primary

! 69! hypotheses involving the composition of the tree community, we determined that our pools represent a gradient of tree communities rather than two discrete treatments (i.e.

“oak” versus “maple” treatment) based upon community analyses of our functional canopy tree groups, and that subsequent analyses should reflect this situation (Figure 4).

We initially tested for collinearity using a correlation matrix between all explanatory variables, eliminating those variables from models with an r > 0.70 (Berry and Feldman,

1985). We utilized repeated measures, mixed effect analyses to model growth rate as it relates to the environmental conditions of the constructed pools. We included year, density, pH, temperature, canopy openness, predatory invertebrate biomass, and percent oak + hickory litterfall as fixed effects. As random effects, we included year, location, and sampling period. We included year in both our fixed and random effects as several species were differentially represented across the two years. Visible inspection of residual plots did not reveal any obvious deviations from homoscedasticity or normality. As our explanatory variables vary in orders of magnitude, we used the scale function in RStudio for ease of cross-variable interpretation. The scale function takes the logarithm of each element of a vector, calculates the mean and standard deviation of the entire vector, and then “scales” these values by subtracting the mean and dividing by the standard deviation. This does not alter results and allows for cross-variable interpretation. The third sampling period of each year was selected for further analyses as this period resulted in maximum captures and was prior to metamorphosis for most species. Of the larvae that were sampled, only species found in greater than 10 pools, comprising at least

500 specimens were included in analyses. From our non-correlated variables, a global model was constructed along with five subsequent, nested models that were applied to

! 70! each species. An additional four candidate models were individually constructed for each species, for a total of 10 models (Tables 3.2-3.5). To examine variation in larval community composition along the environmental gradient, we used multivariate generalized linear models with environmental parameters as predictor variables.

Multivariate generalized linear models were chosen over more traditional distance-based methods (i.e. correspondence analyses, multidimensional scaling) as they more robust and do not confound location with dispersion effects, potentially inflating Type I and

Type II errors (Warton et al., 2012). We constructed ten candidate model sets to analyze the larval community, specifying negative binomial distribution to account for overdispersion of data (Table 3.7). Negative binomial distribution was chosen over

Poisson distribution, as it is more flexible and appropriate for our count data (O’Hara and

Kotze, 2010). Diagnostic plots were checked to ensure that model assumptions were met.

Significance was assessed using 1000 permutations of PIT-trap resampling, a new method which bootstraps probability integral transform residuals and gives the most reliable Type I error rates (Warton and Wang, in review). We used generalized linear models to identify environmental variables significant in both the size (SVL) and mass at metamorphosis for select amphibian species for which modeled growth rate. We applied the same10 models to each species for modeling both mass (g) and SVL (g) at metamorphosis. Models were then evaluated using Akaike Information Criteria (AIC) methods and the most appropriate models selected for further assessment (Burnham and

Anderson, 2002). As the best models are those with the lowest AIC values, we utilized

Akaike weights to assess the likelihood of our other high-ranking models. To assess the importance of individual variables with our larval community models, we used the

! 71! function drop1 to evaluate how AIC values were affected when one of the predictor variables was eliminated from the highest-ranking models.

Results

Larval Growth

A total of 3321 larvae had data recorded on them for growth rate analyses during our 8 sampling periods during 2015 and 2016. Four species were sampled in sufficient numbers and sites to be considered for further analysis (Table 3.1). We found that the percentage allochthonous input comprised of oak-hickory litter and “year” were consistently the top models for three of the four species for which we had greater than

500 individuals.

The most supported model for Jefferson salamander growth rate was “year + % oak-hickory litter” (r2m=0.49 r2c=0.74)(Table 3.2), with it being decreased by year (-

0.175!± SE 0.017) and increased by the percent oak-hickory litter (0.020 ± 0.007).

American toad growth rate resulted in three competing models, with “year + % oak- hickory litter” exhibiting the lowest AIC score (r2m=0.43, r2c=0.69)(Table 3.3). In this model, growth was positively associated with year (0.102 ± SE 0.037), with percent oak- hickory litter not having a significant effect. Chorus frog growth rate resulted in three competing models, with “year + % oak-hickory litter” exhibiting the lowest AIC score

(r2m=0.17, r2c=0.45) (Table 3.4). In this model, growth rate increased with year (0.074!±

SE 0.028) with percent oak-hickory not having a significant effect. The most supported model for wood frog growth rate was “density + % oak-hickory litter” (r2m=0.17,

! 72! r2c=0.41) (Table 3.5). In this model, growth decreased with density (-0.016!± SE 0.007) and increased with percent oak-hickory litter (0.017 ± SE 0.007).

Larval Community Analysis

A total of 17806 larvae were captured for community analyses during sampling periods in 2015 and 2016 (Table 3.6). We did not detect a “treatment” effect, further supporting our decision to analyze our tree community as a gradient as opposed to rank orders. The model with the most support for explaining our larval community was “year + percent oak litter+ mean pH” (p = 0.019) with an AIC weight of 0.92 (Table 3.7). All predictor variables were significant in this model (year, p=0.014; percent oak litter, p=0.028; mean pH, p=0.024). In order to further assess the parsimony of the model we utilized the drop1 command, which reports the change in AIC value for each of the variables that are dropped. These results showed that our AIC score increased with the omission of any variable in our top model. The next highest-ranking models had ΔAIC values >5 points than our top ranking model and are also reported.

Size at Metamorphosis

A total of 605 specimens were utilized for modeling snout-vent length (SVL) and mass

(g) at metamorphosis during the 2015 and 2016 seasons. These 605 specimens comprised four species that were in sufficient numbers to have analyses performed on them: wood frog (195), chorus frog (102), American toad (204), and Jefferson salamander (104). The highest-ranking models, coefficients, and Akaike weights (ωρ) for each species are presented in Table 3.8. Mass at metamorphosis was better modeled than

SVL at metamorphosis for all species. Percent oak litter was a positive, significant variable in all four species predicting mass at metamorphosis, and for three out of four

! 73! species for SVL at metamorphosis, with the exception of American toads, where the top model did not include percent oak litter. Dissolved oxygen was a significant predictor in all models, exerting a positive response in SVL and mass in Jefferson salamanders and wood frogs, and exerting a negative response in chorus frogs and American toads.

Density was a significant variable predicting SVL and mass at metamorphosis in four of our eight models. Density exerted a negative response on mass at metamorphosis for

Jefferson salamanders (-0.45 ± 0.09) and chorus frogs (-0.04 ± 0.008) and a positive response on American toads (0.021 ± 0.003). Density exerted a negative response on

SVL for chorus frogs (-0.65 ± 0.10). pH was a significant predictor in four of our models, exerting a negative effect on mass at metamorphosis for Jefferson salamanders (-0.85 ±

0.13 ) and SVL of wood frogs (-8.90 ± 0.89), and exerting a positive response in

American toad SVL (0.20 ± 0.05) and mass (0.031 ± 0.003). Predatory invertebrate biomass was a significant predictor in four species: Jefferson salamander SVL (-2.45 ±

0.85), wood frog mass (0.08 ± 0.01 ), chorus frog mass (0.13 ± 0.006), and American toad SVL (-0.20 ± 0.07 ).

Discussion

The findings reported here support the numerous mesocosm-based studies suggesting the effects of leaf litter on larval amphibians and the potential impact of landscape-level shifts in the plant community. Ours is the first study to be designed and implemented entirely in the field and without the use of enclosures to investigate the effects of the tree community and subsequent litter on the amphibian community. Though not all species responded similarly, the percentage of oak-hickory litter input was a significant predictor

! 74! in our top models for all species that we investigated as well community structure, and size/mass at metamorphosis. Growth rate and size/mass at metamorphosis are often used as predictors of survival and adult fitness, with larger sizes at metamorphosis being associated with greater survival and greater future reproductive output (Smith 1987;

Semlitsch et al. 1988). If the tree community can alter these metrics, then survival and fitness of the associated amphibian community may also be altered. The positive effects of oak litter have been established in numerous studies and supported by this field study, and the landscape-scale shift from oak dominance to red maple and other species may illicit landscape-level shifts in the amphibian community.

Larval Growth

The percentage of litterfall contributed to oak + hickory was present in the top supporting models for all of our species. Oak + hickory litter was positively related with Jefferson salamander growth rate (0.02 ± 0.008) and with wood frog growth rate (0.02 ± 0.007).

This is consistent with the findings of Rubbo and Kiesecker (2004), where both wood frogs and Jefferson salamanders responded positively to litter derived from oaks. While

Jefferson salamander larvae do not feed directly on leaf litter, oak litter likely supports a grazer-based food web, and thus provides greater invertebrate food sources for this species and wood frogs (Rubbo and Kiesecker, 2004). Interestingly, oak + hickory was not significant in predicting growth rate for American toads (0.005 ± 0.007, p = 0.391) or chorus frogs (0.01 ± 0.007, p= 0.20). It is possible that these generalist species, which breed in a wide range of water bodies, are not as affected by variations in litter mixes as species that typically occupy forested wetlands (Martin et al. 2015).

! 75! Canopy openness is consistently quantified as being a significant predictor of amphibian utilization and growth in isolated woodland pools (Skelly et al. 1999; Werner et al. 2002). Somewhat in contrast to a study by Skelly et al. (2002), which also utilized spring peepers and wood frogs, canopy openness was not present in any of our top models investigating growth rate. However, chorus frogs (spring peeper + mountain chorus frog) did have three competing models with a ΔAIC <2. Canopy openness was present in two of the three competing models and significant in one, where a positive correlation between canopy openness and growth rate (0.02 ± 0.008, p=0.03), seemingly in support of previous studies regarding anurans and canopy openness (Skelly et al. 2002;

Williams et al. 2008). However, the findings of this study should be taken conservatively, as our experiment consisted of only fourteen pools, and the gradient in canopy openness was limited (29.96 ± 12.03), much less than those investigated by Werner et al (2007).

Greater range in canopy openness may have revealed responses more congruent with those in other studies (Skelly et al., 1999, 2002; Werner et al. 2007).

Though predatory invertebrate biomass was not present in the top models for any species, a competing model for American toads (ΔAIC of 1.85) included a significant effect of invertebrate biomass on growth rate (0.02 ± 0.005, p = 0.001). Interestingly,

Werner et al. (2007) found that there was a significant, positive correlation between amphibian species richness and invertebrate predator biomass for both anurans and caudates. It is not clear whether the increased growth rate is a result of the higher invertebrate predator biomass, or if the American toads and their invertebrate predators select for similar habitat. These results are in contrast with those of Watling et al. (2011),

! 76! in which American toads experienced a decline in survival in the presence of predatory invertebrates.

There was an annual effect in the top growth rate model for each species, with the exception of wood frogs. While this may be criticized as affecting our data, it serves to underscore the annual variation exhibited by the larval amphibian community, which is why we chose to include “year” as both a fixed effect and random effect. Moreover, the age of our pools cannot be discounted, as at the end of this study they were approximately two years old, and dynamics related to colonization were still likely being observed. This annual variation is important to acknowledge in this and other studies, as it may result in altered responses to environmental variables experienced by the larval amphibian community. For example, Saenz et al. (2013) focused on the temporal aspect of the effects of litter on amphibian larvae using the invasive Chinese tallow (Triadica sebifera) with regards to senescence and climate change. Their results indicate that larval survival was greatly affected by the length of time Chinese tallow litter had been decomposing, with larvae in treatments simulating warm winters (and thus, early amphibian breeding) exhibiting significantly lower survival than larvae in treatments with cold winters (late amphibian breeding). The winter of 2014/2015 was colder than the winter of 2015/2016, with most species (namely, Jefferson salamander and American toad) breeding in January/February and early March of 2016, respectively. With regards to American toads, “year” was significant in the top model (0.10 ± 0.04, p = 0.01).

Chorus frogs, which also bred early, experienced a strong “year” effect (0.07 ± 0.03, p=0.008). It is possible that the warmer winter altered the decomposition of the litter in

! 77! our pools and altered larval growth rates by affecting nutrient availability or by exposing these species to compounds that they are sensitive to (Saenz et al. 2013).

Wood frogs were unique among our species in that “year” was overridden by

“density” in our top model, with density exerting a negative effect on growth rate (-0.016

± 0.007, p = 0.02). Berven (1990) clearly noted density-dependent effects on growth of wood frogs, with higher densities resulting in longer larval periods and being correlated with lower survival. Moreover, this further supports our current understanding of the effects of density on larval growth and development; in which growth rates at high densities are typically suppressed (Semlitsch and Caldwell, 1982; Berven, 1990; Scott,

1994).

Larval Community

The larval community model underscores the complex dynamics within the pool communities, where the importance of annual variation that the amphibian community exhibits and the composition of the dominant plant community, but also includes another predictor that commonly helps to structure aquatic communities: pH (Freda 1986; Freda and Dunson 1986). While the pH levels noted in Freda and Dunson (1986) as being critical in affecting larvae are not experienced in our pools (mean: 6.5±0.3, range: 5.9-

7.1), pH shows a differential effect by species. In our study, American toads (-2.52), gray treefrogs (-0.15) and chorus frogs (-0.42) exhibit negative coefficient to pH while

Jefferson salamanders (0.56), spotted salamanders (1.87), and wood frogs (1.67) exhibit positive coefficients to pH. Species negatively correlated with pH were positively correlated with percent oak litter and vice versa, despite these variables not being highly associated. Why species whose growth rate was positively correlated with the percent

! 78! oak litter (Jefferson salamander and wood frog) are negatively affected at the community level is not known, though it is possible that there are other interacting variables. What is interesting is that species typically associated with open canopy or early-successional type habitats – American toads, gray treefrogs, and chorus frogs – all were positively correlated with the percentage of oak litter in pools (Werner et al. 2007). It is possible that the activities present in the forest to enhance oak and hickory are also enhancing the habitat for these species.

Size and Mass at Metamorphosis

Size and mass at metamorphosis is considered to be an accurate approximation of subsequent adult fitness (Semlitsch et al. 1988). As such, conditions that increase size and/or mass could be considered beneficial whereas conditions that negatively affect size and/or mass could be considered deleterious. Interestingly, different models explained mass and SVL for each species, and these models differed from those best explaining growth rate.

Also of note, certain “guilds” – such as the more specialist vernal pool-breeding species, wood frogs and Jefferson salamanders - seem to exhibit similar responses to predictors from growth through metamorphosis that are in contrast to those of species that are often considered to be more site generalists, such as American toads and chorus frogs. For example, dissolved oxygen exerts strong positive effects on wood frogs and

Jefferson salamanders, whereas it exerts negative effects on American toads and chorus frogs. Surprisingly little is known about the performance of amphibian larvae and dissolved oxygen, and while the benefits of higher oxygen levels may be easily surmised,

! 79! how such levels could be deleterious to some species is uncertain and warrants further research (Skelly et al. 2002).

Our results suggest that the responses of larval amphibians to litter-based mesocosm studies are translatable to field-settings, with oak and hickory litter exerting significant positive effects, albeit not as exaggerated as those in which mesocosms were used (Skelly, 2002). While this study did not manipulate leaf litter entering the pools, we did place the pools in locations of differing community dominance, and as such the pools experienced a litter mixture comprised of numerous species. If certain plants in the community exert negative effects on amphibians, it is possible that those effects will be negated by other species in the mixture (Stoler and Relyea, 2011). It could be hypothesized however, that the effects documented in our study may become exacerbated if the community continues to shift and becomes homogenized through mesophication

(Melvin and Houlahan, 2012). Regardless, this provides valuable field-based support that large-scale shifts in the dominant plant communities may impact amphibian communities through altered habitat or resource availability.

Our findings suggest that management efforts to increase oak-hickory regeneration, such as selective thinning and prescribed burning, may aid in maintaining amphibian communities in forests (Hutchinson et al., 2008). Further, our research supports that the homogenization and mesophication that is being recorded in many eastern forests, particularly with regards to the increase in red maple, may negatively impact populations of pool-breeding amphibians. Enhancement of community heterogeneity on the landscape should be maintained and encouraged in order to provide the amphibian community with the necessary diversity of conditions and allochthonous

! 80! input to potentially buffer against the negative effects that some species can exert.

Regardless, further research must be conducted in systems that more accurately represent those that are being investigated, potentially through the utilization of in-ground mesocosms in forested settings where they are exposed to natural mixes and volumes of litter.

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! 84!

Figure 3.1. Four of fourteen constructed pools used in larval growth study, located at Vinton Furnace State Experimental Forest, Ohio. Pools measure approximately 5 meters diameter, with one of each pairing in a predominantly oak community and the other in a predominantly maple community.

! 85!

Figure 3.2. Floating leaf litter basket used to sample litterfall in constructed pools. Baskets were tethered to the margins of the pool and were deployed from September through December. One basket was used for each pool and accurately represented the oak-hickory component of the surrounding tree community.

! 86!

Figure 3.3: Efficacy of floating litterfall traps to represent the importance value of the 2 surrounding oak community as measured by percent of total litter sample captured (r adj. = 0.85).

! 87! Table 3.1. Total numbers of larvae sampled from species utilizing consructed pools in Vinton Furnace State Experimental Research forest from 2015-2016 used for growth analyses. * denotes species for which we conducted further analyses.

Species Total Sites Specimens Present

Jefferson salamander* 829 14 (Ambystoma jeffersonianum) American toad* 790 13 (Anaxyrus americanus) wood frog* 583 10 (Lithobates sylvaticus) chorus frog* 574 14 (Pseudacris crucifer + P. brachyphona) gray treefrog 217 9 (Hyla versicolor/chrysocelis) spotted salamander 217 8 (Ambystoma maculatum) green + pickerel Frog 71 8 (Lithobates clamitans + L. palustris) marbled salamander 32 1 (Ambystoma opacum) eastern newt 5 1 (Notophthalmus viridescens) bullfrog 3 1 (Lithobates catesbieanus)

! 88! Table 3.2. Models for growth rate of larval Jefferson salamanders (Ambystoma jeffersonianum) in constructed pools in Vinton Furnace State Experimental Forest in 2015 and 2016. Marginal r2 (r2 m) describes variance explained by fixed effects. Conditional r2 (r2 c) describes variance explained by fixed and random effects. Akaike parameter weights (ωρ). * denotes significant predictors (p<0.05).

2 2 Model Predictors r m r c AIC ΔAIC ωρ can4 year + % oak-hickory litter* 0.49 0.74 -1155.38 0 0.93 pH year* + pH 0.42 0.78 -1149.79 5.59 0.06 can3 year* + density + % oak-hickory litter* 0.49 0.74 -1147.20 8.18 0.02 ptemp year + pH + temp 0.42 0.78 -1142.20 13.18 0.00 ptemlit year*+pH*+temp*+%oak-hickory litter* 0.44 0.76 -1139.60 15.78 0.00

Table 3.3. Models for growth rate of larval American toads (Anaxyrus americanus) in constructed pools in Vinton Furnace State Experimental Forest in 2015 and 2016. Marginal r2 (r2 m) describes variance explained by fixed effects. Conditional r2 (r2 c) describes variance explained by fixed and random effects. Akaike parameter weights (ωρ). * denotes significant predictors (p<0.05).

2 2 Model Predictors r m r c AIC ΔAIC ωρ can4 year* + % oak-hickory litter 0.43 0.69 -1531.57 0 0.44 pH year* + pH 0.37 0.74 -1530.96 0.61 0.33 can2 year + pH + invert* 0.37 0.74 -1529.72 1.85 0.18 can3 year + % oak-hickory litter* + density* 0.38 0.74 -1526.85 4.72 0.04 ptemp year* + pH + temp 0.37 0.74 -1522.90 8.66 0.01

! 89! Table 3.4. Models for growth rate of larval chorus frogs (Pseudacris spp.) in constructed pools in Vinton Furnace State Experimental Forest in 2015 and 2016. Marginal r2 (r2 m) describes variance explained by fixed effects. Conditional r2 (r2 c) describes variance explained by fixed and random effects. Akaike parameter weights (ωρ). * denotes significant predictors (p<0.05) .

2 2 Model Predictors r m r c AIC ΔAIC ωρ can4 year* + % oak-hickory litter 0.17 0.45 -1159.89 0 0.39 can3 year* + canopy 0.16 0.45 -1158.67 1.21 0.21 pH year* + pH 0.16 0.46 -1158.56 1.32 0.20 can2 temp* + canopy* 0.04 0.73 -1158.42 1.47 0.19 ptemp year* + pH + temp 0.17 0.47 -1150.27 9.62 0.00

Table 3.5. Models for growth rate of larval wood frogs (Lithobates sylvaticus) in constructed pools in Vinton Furnace State Experimental Forest in 2015 and 2016. Marginal r2 (r2 m) describes variance explained by fixed effects. Conditional r2 (r2 c) describes variance explained by fixed and random effects. Akaike parameter weights (ωρ). * denotes significant predictors (p<0.05).

2 2 Model Predictors r m r c AIC ΔAIC ωρ can4 density* + % oak-hickory litter* 0.17 0.41 -1076.47 0 0.49 pH year + pH 0.10 0.48 -1076.25 0.22 0.44 can3 invert* + % oak-hickory litter + density* 0.19 0.39 -1072.39 4.08 0.06 ptemp year + pH + temp 0.11 0.49 -1068.03 8.44 0.01 can1 % oak-hickory litter *+ canopy + density 0.14 0.45 -1066.95 9.52 0.00

! 90! Table 3.6. Total numbers of larvae sampled from species utilizing constructed pools in Vinton Furnace State Experimental Research forest from 2015-2016. * denotes species underrepresented in sampling due to incongruous breeding season.

Species Total Sites Specimens Present

Jefferson salamander 967 14 (Ambystoma jeffersonianum) spotted salamander 168 8 (Ambystoma maculatum) marbled salamander* 23 1 (Ambystoma opacum) chorus frog 864 14 (Pseudacris crucifer + P. brachyphona) gray treefrog* 337 9 (Hyla versicolor/chrysocelis) American toad 12766 13 (Anaxyrus americanus) wood frog 2681 10 (Lithobates sylvaticus)

! 91! Table 3.7: Five most supported models for larval amphibian community in constructed pools in Vinton Furnace State Experimental Research forest from 2015-2016 with Akaike Information Criteria (AIC) scores, Δ AIC, and Akaike parameter weights (ωρ) * Indicates significant variables (p≤0.05).

Model Parameters ΔAIC drop1 ωρ P ΔAIC Can10 year* + % oak * + mean pH* 0 year: 63.5 0.92 0.02 oak: 28.4 pH: 9.6

Can3 year*+ mean pH + % oak + canopy 5.7 0.0 0.02 Can8 mean pH + % oak + invert + density 7.2 0.02 0.17 Can4 year* + % oak* 28.4 0.0 0.04 Can1 year + canopy + density + invert 36.1 0.0 0.08

! 92! Table 3.8: Most supported models for size and mass at metamorphosis for four amphibian species in constructed pools in Vinton Furnace State Experimental Forest with model parameters with associated coefficients and Akaike parameter weights (ωρ) * Indicates significant variables (p≤0.05).

Species Parameters ωρ

Jefferson SVL: DO2*(5.86 ± 0.76) + density + % oak* (3.01 ± 0.83) + 0.76 salamander invert*(-2.45 ± 0.85) + canopy* (-3.63 ±0.67) (Ambystoma jeffersonianum) Mass: DO2*(0.79 ± 0.09) + pH*(-0.85 ± 0.13) + 0.97 density*(-0.45 ± 0.09) + % oak*(0.65 ± 0.10)

wood frog SVL: DO2*(4.01 ± 0.19) + pH*(-8.90 ± 0.89) 0.32

(Lithobates + % oak* (3.76 ± 0.19) + canopy* (6.36 ± 0.76) sylvaticus)

Mass: DO2*(0.30 ± 0.03) + density + % oak*(0.32 ± 0.02) 0.99

+ invert*(0.08 ± 0.01) + canopy*(-0.06 ± 0.01)

chorus frog SVL: DO2*(-0.82 ± 0.09) + density* (-0.65 ± 0.10) 0.35

(Pseudacris + % oak* (0.28 ± 0.11) + canopy spp.)

Mass: DO2* (-0.03 ± 0.005)+ density*(-0.04 ± 0.008) 0.59

+ % oak*(0.02 ± 0.007) + invert* (0.13 ± 0.006) + canopy

American toad SVL: DO2* (-0.69 ± 0.06) + pH* (0.20 ± 0.05) 0.85

(Anaxyrus + invert* (-0.20 ± 0.07) + canopy* (0.28 ± 0.06) americanus)

Mass: DO2*(-0.013 ± 0.001) + pH*(0.031 ± 0.003) 0.99

+ density*(0.021 ± 0.003) + % oak* (0.006 ± 0.001)

! 93! Chapter 4: Amphibian Connectivity in a Heterogeneous Landscape: The Role of Constructed Ridge-top Pools

Abstract

The maintenance, persistence, and potential enhancement of amphibian populations depend upon the dispersal ability of a species and the connectedness of their habitat. Although most studies on dispersal and connectivity of amphibians focus on fragmented landscapes, it remains important to investigate these traits in less-impacted landscapes in order to conserve common (but vulnerable) species and better understand their capacity to respond in less-altered conditions, such as those exhibited by large nature preserves, state, and national forests. Studies focused on pool-breeding amphibians frequently take the “pond-as-patches” approach to metapopulation dynamics, treating ponds as discrete habitat patches. Efforts to increase the functional connectivity of these species may be achieved by increasing structural connectivity in the form of constructed breeding pools. In order to quantify colonization potential of amphibians in relatively intact forested landscapes, we constructed 14 ridgetop pools in a largely forested landscape in southeastern Ohio and recorded species colonization in those sites for two years. We utilized the species abundance data to model landscape connectivity based on three metrics: distance to natural wetlands, slope to natural wetlands, and area of natural wetlands. We found that Jefferson and spotted salamander (Ambystoma jeffersonianum and A. maculatum, respectively) and eastern newt (Notophthalmus viridescens)

! 94! connectivity was largely a function of distance and mean slope to wetlands, whereas anuran connectivity was more ubiquitous and did not relate strongly to our candidate model set. Our results suggest that creating isolated pools in locations that minimize distance and slope to the nearest wetland can facilitate connectivity in unfragmented landscapes with disjointed wetlands by enhancing the structural connectivity of amphibian populations. Importantly, these models of connectivity provide insight to the optimal number of pools to facility connectivity without over development, where the benefits of isolation and buffering of critical wetlands are preserved. Finally, our approach demonstrates that creating these pools can be easily integrated within current forest management settings and their size and ease of installation may provide additional wildlife benefits.!

.

! 95! Introduction

The degree of connectedness between breeding sites, coupled with their dispersal ability, is critical to the maintenance and persistence of a species on a landscape (Rothermel

2004). Pool-breeding amphibians, such as the mole salamanders (Ambystoma spp.) and wood frogs (Lithobates sylvaticus), may exhibit high rates of localized extinctions due to drought, disease, or other factors, even in relatively intact ecosystems (Marsh & Trenham

2001). In order for these species to persist at the landscape level, it is imperative that dispersal rates and connectivity between breeding sites allow for colonization of novel sites or recolonization of historic sites (Hanski 1998; Marsh & Trenham 2001). Habitat fragmentation only reduces the dispersal capabilities of these species and thus requires that alternate breeding sites be in closer proximity on the landscape (Guerry & Hunter

2002; Smith & Green 2005; Cushman 2006).

Dispersal and colonization are interrelated, and pool-breeding amphibians have been the subjects of debate with regards to their population structure being accurately described as truly functioning as a metapopulation or more as a patchily distributed population (Semlitsch 2000; Marsh & Trenham 2001; Smith & Green 2005).

Metapopulations exist when there is a “population” of unstable local populations that inhabit discrete patches which undergo local extinctions and subsequent recolonizations through dispersal, allowing a regional population to persist (Hanski 1998). Alternatively, these populations could be viewed as a “patchily distributed populations”, where breeding sites are treated as patches (a “ponds as patches” approach) and movements occur at high enough rates such at local breeding ponds fail to develop a significant degree of demographic independence, lacking true metapopulation structure (Marsh &

! 96! Trenham 2001; Smith & Green 2005; Petranka & Holbrook 2006) However, many pool- breeding amphibians are sampled while at breeding sites for convenience, despite most species spending comparatively little time at these locations (Wells 1977). As such, studies that focus heavily or exclusively within the “ponds as patches” framework will generally lead to pond-based explanations for distribution and abundance on the landscape, underestimating terrestrial habitat, landscape features, and dispersal (Marsh &

Trenham 2001). It is imperative to include landscape features that can either enhance or impede the movement of a species to/from breeding sites in order to gather a greater understanding of habitat connectivity on the landscape.

For clarification, connectivity refers to the extent to which certain landscapes impede or facilitate movement of individuals (Taylor et al. 1993). Further, landscape connectivity consists of two interrelated components as identified by O’Brien et al.

(2006): structural connectivity and functional connectivity. Structural connectivity represents the physical and spatial arrangement of patches of habitat, whereas functional connectivity represents the response of organisms to the physical/spatial arrangement of those habitat patches (i.e. structural connectedness). Crawford et al. (2016) found that increased functional connectivity of a population of Jefferson salamanders (Ambystoma jeffersonianum) was significantly related to allelic richness and heterozygosity, suggesting that dispersal among wetlands was common. Conversely, they also found that decreased connectivity resulted in increased genetic differentiation, suggesting that movement and dispersal was limited. While this is of particular importance in understanding population responses in fragmented landscapes, this may also be applied to a relatively intact landscape or system (such as national and state forests) where the goal

! 97! is to enhance populations through enhancing the functional connectivity of these discrete breeding sites. For example, Crawford et al. (2016) recommend that future management strategies for wetland-breeding salamanders requires strategic placement of wetlands on the landscape, particularly the forested distance to other breeding sites. The findings of our research will help further the understanding of species-specific movements across the landscape, including habitat features that may enhance connectivity.

Colonization by amphibians of novel breeding locations is not uncommon, although rates may be affected by the distance to nearest breeding sites, individual species dispersal capability, and landscape characteristics of intervening habitat(Pechmann et al. 2001; Crawford et al. 2016). Although many pool-breeding amphibians are philopatric to the breeding sites, they do exhibit some behavioral plasticity to disperse and colonize alternate sites (Pechmann et al. 2001; Gamble et al.

2007; Semlitsch 2008) For example, Breden (1987) found that 17% of adult Fowler’s toads (Anaxyrus fowleri) relocated each year, and Gill (1978) found that while adult red- spotted newts (Notophthalmus virdiscens) are nearly 100% faithful to breeding sites, the juveniles often disperse to other sites. By capitalizing on these dispersing individuals through the construction of strategically placed pools on the landscape, both structural and functional connectivity would be enhanced.

In order to assess aspects of connectivity and dispersal of pool-breeding amphibians, and as part of a larger study investigating the impacts of shifting plant communities on amphibians and the capacity of constructed ridge-top pools to provide pool-breeding amphibian habitat, we examined the degree of breeding site connectivity in a managed forest and the degree to which that affected subsequent colonization by

! 98! amphibians in constructed ridge-top pools. The alternative models that were tested are as follows: (1) connectivity is a function of distance to natural wetlands, (2) connectivity is an artifact of size of natural wetlands, (3) connectivity is an artifact of slope to natural wetlands, and (4) connectivity is an artifact of distance + slope + area of natural wetlands.

In a landscape where naturally existing pools typically occur within the basins of small streams that are separated by high ridges and steep slopes, we hypothesized that greater slopes would lessen the degree of connectivity between sites. Further, we hypothesized that the construction of pools on these ridge tops would increase the connectedness between populations in different stream basins, increasing the dispersal abilities of these species and thus, their potential range on the landscape.

Methods

Study Site

This study was conducted at Vinton Furnace State Experimental Forest (VFSEF), a research forest of approximately 4,900-hectares located in Vinton County of southeastern

Ohio. The study site is situated in the unglaciated region of the Allegheny Plateau approximately 50 kilometers north of the Ohio River. The area is highly dissected into narrow ridges, steep slopes, and small valleys, with elevations of 200 to 300 meters above sea level (Brockman 1998). Soils of VFSEF are unglaciated silt loams derived primarily from sandstones, siltstones, and shales. The region receives a mean precipitation of 1024 mm and has an annual mean temperature of 11.3°, with the coldest month (January) experiencing a mean temperature of -1.4° and the warmest month (July) experiencing a mean temperature of 22.7°. Dominant oak species at VFSEF are white

! 99! oak (Quercus alba L.)(21% basal area), black oak (Quercus velutina Lam.) (17% basal area), chestnut oak (Quercus montana Willd.)(15% basal area), and scarlet oak (Quercus coccinea Muenchh.) (12% basal area (Hutchinson et al. 2008).

Natural Vernal Pools: Locating and Amphibian Sampling

VFSEF was searched from December 2012-May 2014 for all naturally existing sites that harbored pool-breeding amphibians using satellite imagery and local knowledge coupled with ground-truthing. Sites that contained pool-breeding amphibian had their area estimated using satellite imagery with Google Earth Pro (version 7.1.5.1557) and were further surveyed using aquatic funnel traps or dip-netting to confirm breeding during this same period. Species were recorded as being either “present” or “absent” at each site.

Pool Construction: Experimental Design and Amphibian Sampling

Seven pairs (fourteen total) of 5-meter diameter pools approximately 0.75-meter in depth were constructed along a series of ridges in June 2014 using a small bulldozer provided by the Ohio Division of Forestry. Pools were not lined with any material as the sites exhibited a high enough percentage of clay in the native soil to retain water once the sites were sufficiently compacted by the bulldozer. Each pool was made to exhibit a gradual slope to the point of maximum depth. With the exception of utilizing a native grass and forb mix to initially stabilize soil and prevent excessive erosion at the point of entry for the bulldozer, no additional site manipulation occurred.

Breeding amphibians were then sampled from early March through mid-April of

2015 and 2016. We chose to survey during this period to maximize detectability for the pool-breeding amphibian species within the region as they migrated and emigrated from newly constructed pools. We sampled amphibians with aquatic funnel traps at a density

! 100! of 1 trap/m2 of surface area (5-6 traps/pool) and drift fence-pitfall arrays partially enclosing each pool. Aquatic traps were set and staked equidistant along the perimeter of each pool at a depth that did not exceed one-half the height of the trap (approximately 12-

15 cm), fully immersing the entry hole. Drift fence arrays only partially encircled each pool to ensure that colonizing amphibians were not dissuaded from utilizing pools. Each pool had three, 5-meter drift fences constructed of silt fabric arranged in a triangular pattern around each pool and were approximately 30 centimeters in height and were buried approximately 3-5 centimeters below the soil surface. Drift fences had a 5-quart bucket buried at either end that was equipped with a clear plastic funnel to prevent amphibians from escaping. Captured amphibians were measured (snout-vent length + tail length), weighed (g), sexed, and permanently marked with visible implant elastomer

(Northwest Marine Technology, Inc., Shaw Island, WA) to identify the initial location of capture and prevent re-sampling of previously captured individuals.

All sampling and handling protocols were approved by The Ohio State University

Institute for Animal Care and Use Committee (protocol # 2011A00000082-R1) and under

Ohio Division of Wildlife Permit 16-104.

Statistical Methods

We utilized ArcGIS (ArcGIS Desktop: Release 10.2 Redlands, CA: Environmental

Systems Research Institute) to determine coordinates of naturally occurring vernal pools and our created pools and determined the maximum slopes, mean slopes, and topographic distances along a straight line between naturally occurring wetlands using a 10-m digital elevation model (DEM) derived from the United States Geological Survey

! 101! (www.viewer.nationalmap.gov). This information was uploaded into RStudio Version

0.99.467 (RStudio Team, 2015) where further analyses were conducted (Figure 4.1).

We assessed wetland connectivity between 12 naturally occurring wetlands and

14 constructed pools using an incidence function that accounts for distance to all wetlands using a negative exponential dispersal function developed in Crawford et al.

(2016). Further, we modified aspects of this function (i.e. penalizing connectivity by slope) to expanding the usage to a different landscape and for multiple species in the community. The general form of the incidence function calculating connectivity (Ci) of wetland i derived from Crawford et al. (2016) is:

!! = ! exp!(−∝ !!")!!!!!!!!!!(1) !!!

! where α represents of an individual species, dij is the topographical distance !"#!!!"#$%&#'( between the constructed pool (i) and the natural source wetland (j). This tests the hypothesis that connectivity is a function of distance to wetlands. We modified Eq. 1 to assess alternative hypotheses regarding how the level of connectivity affects amphibian species counts at constructed pools. In Eqs. 2 & 3, we incorporated slope by dividing

(penalizing) our general function by the natural log of either the mean slope (!!"!") or

maximum slope (!!"!") to nearest wetlands to test the hypothesis that the slope to other wetlands will inhibit connectivity. In Equation 4 we multiplied (weighted) our general

2 function by incorporating the area (m ) of natural wetlands (Aj), testing the hypothesis the larger natural wetlands will attribute more to overall connectivity in the landscape.

Multiplying by the natural wetland area assumes that larger wetlands will have larger

! 102! populations of amphibians and thus increases the probability that the nearest constructed pool will be colonized.

exp!(−∝ !!") !! = ! !!!!!!!!!!(2) ln!(!!" ) !!! !"

exp!(−∝ !!") !! = ! !!!!!!!!!!(3) ln!(!!" ) !!! !"

!! = ! exp!(−∝ !!") !!!!!!!!(4) !!!

In Eqs. 5 & 6, we further modified our general formula of the incidence function by

dividing it by either the natural log of the mean slope(!!"!") or the maximum

slope(!!"!"), and multiplying each by the area (Aj) of the natural wetlands. This tests the hypothesis that connectivity is a function of not only distance, but also slope, and wetland size.

exp!(−∝ !!") !! = ! !!!!!!!!(5) ln!(!!" ) !!! !"

exp!(−∝ !!") !! = ! !!!!!!!!(6) ln!(!!" ) !!! !"

In order to specify α for each species, a literature search was conducted to determine the range of dispersal for each species. In general, we were liberal with our estimates of dispersal, assuming farther dispersal distances than the mean of those reported (Berven & Grudzien 1990; Marsh & Trenham 2001; Semlitsch 2002; Alex

! 103! Smith & M Green 2005; Petranka & Holbrook 2006; Gamble et al. 2007; Crawford et al.

2016).

We applied these six functions to the species abundance matrix from each constructed pool for the year of initial colonization (2015 and 2016, depending on species) using linear regression models with α optimized for each species. We specified a Poisson distribution , as counts comprise our data. Models were then evaluated using

Akaike Information Criteria (AICc) and the most appropriate models selected for further assessment (Burnham et al. 2002). As the best models are those with the lowest AIC values, we utilized Akaike weights to assess the likelihood of our other high-ranking models.

Results

Connectivity was modeled for six species in 2015 and eight species in 2016. Two additional species colonized our constructed pools in 2016 in sufficient numbers to model connectivity: four-toed salamanders and eastern newts. Tables are only presented for the

2016 sampling as the second season provided a means to further support or negate the

2015. Connectivity was most supported for Jefferson salamanders (α 1500 meters) with

Equation 2 (distance + mean slope) for both 2015 and 2016 (Table 4.1). Data for 2016 further supported the 2016 model with no other models emerging as competitive.

Connectivity was most supported for spotted salamanders (α 1500 meters) with Equation

2 (distance + mean slope) as well. This was the top model for our 2015 spotted salamander data, although there were several competing models. With the addition of

2016 data, Equation 2 strongly emerged as our top model with no competing models

! 104! (Table 4.2). There was no support for weighting connectivity by natural wetland area for either species of ambystomatid. Connectivity models for wood frogs (α 1500 meters) differed between years, with Equation 4 (distance + area) being the only to exhibit a significant effect (Table 4.3). Two competing models (Equations 6 and 5, respectively) exhibited a ΔAICc < 2, with both indicating an effect of slope, distance, and area on connectivity. Green frogs (α 2500 meters) exhibited site connectivity that was best supported by Equation 1 (distance only) for both 2015 and 2016, being further strengthened with 2016 data and with no competing models (Table 4.4). None of our equations were able to disentangle artifacts of connectivity for chorus frogs (α 500 meters) (Table 4.5) or American toads (α 2500 meters) (Table 4.6), with the largest

ΔAICc < 2.5 for chorus frogs and ΔAICc < 0.50 for American toads. Four-toed salamanders (α 1000 meters) were new to the constructed pools in 2016, with Equations

5, 6, and 4 emerging as our most supported models. Equations 5 & 6 were nearly identical (ΔAICc 0.01) and Equation 4 (ΔAICc 0.83) was also competitive (Table 4.7).

However, weighting by wetland area supports all top models (Equations 4-6). As with both ambystomatids, connectivity was best supported for eastern newts (α 2500 meters) by Equation 2 (distance + mean slope), with no competing models (Table 4.8).

Discussion

Even within largely protected, unfragmented landscapes, breeding sites can be a limiting feature for species on the landscape (Gorman et al. 2009). Our study, which utilized constructed ridge-top pools to explore artifacts of connectivity in a heterogeneous landscape, revealed a range of species-specific effects. Despite limited sample size, we

! 105! managed to detect a strong signal to support that distance to natural (source) wetlands and mean slope to the constructed pool are important in determining not only landscape connectivity, but also potential success of constructed wetlands. Further the constructed pools were colonized by a total of thirteen pool-breeding species within a two-year time frame, representing the full complement of pool-breeding species in the region (Table

4.9). These pools exhibited a diverse larval community and produced metamorphosed individuals of each of the species that we recorded as having colonized them (Graziano and Matthews, unpublished data). As such, we found that by increasing structural connectivity on the landscape with small, constructed ridge-top pools, we increased the functional connectivity of pool-breeding amphibians on the landscape, as evidenced by colonization of these species.

The maintenance, persistence, and potential enhancement of amphibian populations depends upon the dispersal ability of a species and the connectedness of their habitat (Gill 1978; Marsh & Trenham 2001; Rothermel 2004; Cushman 2006). With regards to created wetlands, many studies have established and corroborated how to build these pools to best suit amphibians (i.e. focusing on hydroperiod, slope, size), but relatively few establish where these pools should be constructed(Semlitsch 2000;

Pechmann et al. 2001; Denton & Richter 2013). Although most studies on dispersal and connectivity focus on fragmented landscapes, it remains important to investigate these traits in less-impacted landscapes as well, such as those exhibited by large nature preserves, state, and national forests. These tracts of land are likely to be essential in sustaining future populations of many organisms, and it is important to understand how amphibians traverse the landscape and how efforts to enhance habitat are received. Even

! 106! within large and relatively intact systems such as national forests, pool-breeding amphibians may exhibit high rates of localized extinctions due to drought, disease, or other factors (Marsh & Trenham 2001). In order to assure future persistence of these species, it may be necessary to augment the landscape with sites that provide additional breeding opportunities to buffer against stochastic events, environmental, and anthropogenic stressors (Marsh & Trenham 2001; Pechmann et al. 2001; Semlitsch 2002;

Griffiths et al. 2010; Crawford et al. 2016). Further, it is necessary to understand how focal species or the communities may respond to activities focused on increasing dispersal and colonization; as our study highlights, connectivity is species-specific and factors that affect it may be difficult to parameterize or otherwise model(Berven &

Grudzien 1990; Taylor et al. 1993).

For three of the four salamander species that we analyzed in our study (Jefferson salamander, spotted salamander, and eastern newt), landscape connectivity was most supported by Equation 2, specifying distance to natural wetlands and the mean slope to them. While Eq. 2 supported Jefferson salamander connectivity in 2015, strength for this model was increased with 2016 data. Interestingly, spotted salamanders exhibited a lag in colonization that was not well modeled by our connectivity functions for the 2015 year, but with the addition of 2016 data the models assumed nearly the same order of

Jefferson salamanders (with a single exception). Our findings have clear applied management strategies, as land managers can locate natural source wetlands within the dispersal distance noted for each species and determine the least slope to them with relative ease, increasing the chance of enhancing the functional connectivity of these species on the landscape should they decide to construct new wetlands (Table 4.10). The

! 107! fourth species of salamander to utilize our sites, the four-toed salamander, is listed as a special of concern, threatened, or endangered in several states (NatureServe 2015).

Despite not having results as clear as those with the aforementioned species, we did find strong evidence that wetland area is an important contributor to enhancing connectivity for this species, in addition to the distance and slope to those wetlands. While four-toed salamanders are capable of utilizing various sites for oviposition, the extent of these sites may be more important to maintaining populations of this species, and management focused on the creation or enhancement of a mosaic of wetlands may be an important determinant in future resilience of this species on the landscape(Marsh & Trenham 2001;

Semlitsch 2002; Crawford et al. 2016).

The connectivity functions for most of the anuran species in our study failed to estimate underlying drivers in our system. Green frogs were the only species that resulted in a clear top model that estimated connectivity, with the distance to natural wetlands

(Eq. 1) being the top model for 2015 data, being further strengthened by 2016 data.

Green frogs are not typically considered desirable species in vernal pools or constructed wetlands as they can quickly dominate the habitat and outcompete more desirable species

(i.e. wood frogs, ambystomatids)(Vasconcelos & Calhoun 2006) As such, from a restoration or management perspective, this not only underscores the importance of maintaining an ephemeral hydroperiod, but also supports the importance of isolated wetlands(Vasconcelos & Calhoun 2006; Denton & Richter 2013) Moreover, these results could be expanded to suggest limiting sites that green frogs utilize for reproduction (i.e. farm ponds that exhibit a permanent hydroperiod) in a landscape where efforts are focused on the enhancement of vernal pool obligates. Connectivity equations did not

! 108! perform as well for wood frogs in our landscape, with both 2015 and 2016 data having three competing models that were indiscernible from each other (ΔAICc < 2). However, for at least the first year of colonization, distance and slope (mean and maximum) comprised 80% of our model weights, suggesting that initial colonization may be a function of these landscape characteristics. With the addition of 2016 data, an alternate pattern emerged: one that clearly supported weighting connectivity by wetland area

(Equations 4-6). This could suggest that connectivity may initially be an artifact of distance and slope, but subsequent years are more strongly determined by their connectivity to more extensive natural pools, perhaps as a result of the more stochastic life history (boom and bust) observed in wood frogs (Berven 1990). Wood frogs have been the subject of considerable research, with most studies suggesting a species that requires large tracts of forest, exhibits a complex life history, and is prone to high fluctuations in population(Berven 1990; Berven & Grudzien 1990; Joly et al. 2003;

Herrmann et al. 2005). In light of this information, it would make sense that landscape connectivity for wood frogs is largely an artifact of not only distance to wetlands, but also the area of these wetlands (assuming that this area consists of numerous small breeding sites as opposed to one large site). Our equations did not clearly elucidate the underlying drivers dictating connectivity for either chorus frogs (spring peepers + mountain chorus frogs) or American toads. It is unclear why our equations performed so poorly for these species, although it could be inferred that there are more important drivers of dispersal and connectivity for these species, or hypothesized that these species exhibit significant annual variation in populations as a result of recruitment which negatively affected our ability to accurately model landscape connectivity for them (Trenham et al. 2003).

! 109! However, species that are as ubiquitous as American toads and chorus frogs in a fragmented landscape (i.e. agricultural and ex-urban regions) may exhibit dispersal abilities in an unfragmented landscape that nullify inter-pool distance and slope-effects due to their increased mobility compared to caudates, which are more reliably modeled.

Connectivity functions were more effective in modeling salamander than anuran abundance at our constructed pools likely as a result of differences in three traits: (1) longevity, (2) breeding strategies, and (3) dispersal/occupancy. Salamanders, particularly ambystomatids and newts, are relatively long-lived organisms when compared to many anuran species, namely wood and chorus frogs (Berven 1990; Newcomb Homan et al.

2003; Homan et al. 2007; Werner et al. 2007). As a result, there may be larger standing populations of the former on the landscape, the culmination of potentially decades of breeding events (Newcomb Homan et al. 2003; Homan et al. 2007). Anurans, such as wood and chorus frogs, are comparatively short-lived and individuals may only survive for one or two breeding events. To compensate, these species may exhibit “boom and bust” population cycles and as a result, dispersal and connectivity measures may exhibit considerable annual variation that we were unable to capture these dynamics of the two years of sampling (Gill 1978; Berven 1990; Crawford et al. 2016). Green frogs are longer-lived, exhibit higher plasticity in breeding sites, and have relatively simple life histories compared to the two aforementioned species, and as a result may lend their dispersal behavior and connectivity measures to being more easily modeled (Birchfield &

Deters 2005). Finally, our models assume random dispersal and colonization across the landscape. Anuran occupancy of breeding sites is likely the result of a nonrandom behavior: vocalization. Wood and chorus frogs, and American toads are all species that

! 110! exhibit gregarious breeding behavior, cueing in on sites due to the presence of calling males (Martof 1953). While some landscape characteristics may help determine the occupancy of a site, a calling male could potentially attract a disproportionate amount of individuals to a site, regardless of landscape characteristics (Berven 1981). While physical characteristics of the landscape may be appropriate and accurate in modeling connectivity and dispersal in salamanders, these traits may be insufficient to capture some of the underlying drivers in anuran dispersal and connectivity.

The ability of amphibians to disperse through the landscape and propensity to colonize novel breeding sites has been well documented(Berven & Grudzien 1990;

Pechmann et al. 2001; Birchfield & Deters 2005; Denton & Richter 2013). While this knowledge is often viewed through a negative lens with respect to fragmented landscapes, it should also be viewed as a powerful ally to biologists and land managers seeking to buffer existing amphibian populations or enhance current populations in large, forested landscapes. With increased habitat fragmentation on the landscape, robust amphibian populations in large tracts of relatively unfragmented forests should not be taken for granted, and ecologists should provide land managers with the necessary information as how to best manage these valuable sites for biodiversity.

In support of prior knowledge of pool-breeding amphibians, weighting our connectivity functions by wetland area was not a significant determinant in most of our models, with the exception of wood frogs and four-toed salamanders(Snodgrass et al.

2000). From a management and restoration perspective, this suggests that creating large wetlands may not be critical in conservation efforts. Rather, it would suggest that greater numbers of small wetlands be constructed or enhanced in an effort to provide multiple

! 111! breeding sites/opportunities to these species. Fortunately this is in line with many restoration and management guidelines: the maintenance or enhancement of landscape- level habitat heterogeneity (Carey 2003; Fahrig et al. 2011). In order to foster metapopulation dynamics, the creation of a mosaic of suitable breeding sites may be critical in maintaining robust amphibian populations. By capitalizing on dispersing individuals and strategically placing breeding sites on the landscape, the structural and functional connectivity of pool breeding amphibians may be enhanced.

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! 113! Marsh DM, Trenham PC. 2001. Metapopulation dynamics and amphibian conservation. Conservation biology 15:40–49. ! Martof BS. 1953. Territoriality in the Green Frog, Rana Clamitans. Ecology 34:165–174. ! Newcomb Homan R, Reed JM, Windmiller BS. 2003. Analysis of Spotted Salamander (Ambystoma maculatum) Growth Rates Based on Long-Bone Growth Rings. Journal of Herpetology 37:617–621. O’Brien D, Manseau M, Fall A, Fortin M-J. 2006. Testing the importance of spatial configuration of winter habitat for woodland caribou: An application of graph theory. Biological Conservation 130:70–83. ! Pechmann JHK, Estes RA, Scott DE, Gibbons JW. 2001. Amphibian colonization and use of ponds created for trial mitigation of wetland loss. Wetlands 21:93–111. ! Petranka JW, Holbrook CT. 2006. Wetland restoration for amphibians: should local sites be designed to support metapopulations or patchy populations? Restoration Ecology 14:404–411. ! Rothermel BB. 2004. Migratory Success of Juveniles: A Potential Constraint on Connectivity for Pond-breeding Amphibians. Ecological Applications 14:1535– 1546. ! Semlitsch RD. 2000. Principles for Management of Aquatic-Breeding Amphibians. The Journal of Wildlife Management 64:615. ! Semlitsch RD. 2002. Critical elements for biologically based recovery plans of aquatic- breeding amphibians. Conservation biology 16:619–629. ! Semlitsch RD. 2008. Differentiating Migration and Dispersal Processes for Pond- Breeding Amphibians. Journal of Wildlife Management 72:260–267. ! Snodgrass JW, Komoroski MJ, Bryan AL, Burger J. 2000. Relationships among isolated wetland size, hydroperiod, and amphibian species richness: implications for wetland regulations. Conservation biology 14:414–419. ! Taylor PD, Fahrig L, Henein K, Merriam G. 1993. Connectivity Is a Vital Element of Landscape Structure. Oikos 68:571. ! Trenham PC, Koenig WD, Mossman MJ, Stark SL, Jagger LA. 2003. Regional dynamics of wetland-breeding frogs and toads: turnover and synchrony. Ecological Applications 13:1522–1532. !

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! 115!

Figure 4.1. Shaded relief map of Vinton Furnace State Experimental Forest with locations of natural and constructed pools. Blue triangles represent constructed pool pairs and black stars denote locations of naturally occurring pools.

! 116! Table 4.1. Factors influencing site connectivity of Jefferson salamanders (Ambystoma jeffersonianum) for 2015-2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 1500m

Equation ( ) AICc ΔAICc AICcWt Cum.Wt LL (2)dist_mn.slope* 129.13 0.00 1.00 1.00 -62.02

(3)dist_mx.slope* 150.04 20.90 0.00 1.00 -72.47

(1)dist* 161.50 32.37 0.00 1.00 -78.20

(5)dist_mn.slope_area* 190.67 61.54 0.00 1.00 -92.79

(6)dist_mx.slope_area* 208.74 79.60 0.00 1.00 -101.82

(4)dist_area* 217.50 88.37 0.00 1.00 -106.21

Table 4.2. Factors influencing site connectivity of spotted salamanders (Ambystoma maculatum) for 2015-2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 1500m

Equation ( ) AICc ΔAICc AICcWt Cum.Wt LL (2)dist_mn.slope* 214.24 0.00 1 1.00 -104.57

(3)dist_mx.slope* 242.15 27.91 0.00 1.00 -118.53

(1)dist* 250.13 35.89 0.00 1.00 -122.52

(5)dist_mn.slope_area* 2 83.54 69.30 0.00 1.00 -139.22

(6)dist_mx.slope_area* 290.64 76.40 0.00 1.00 -142.78

(4)dist _area* 293.52 79.28 0.00 1.00 -144.21

! 117! Table 4.3. Factors influencing site connectivity of wood frogs (Lithobates sylvaticus) for 2015-2016. * indicates statistically significant model (p<0.10) using Poisson distribution. α = 1500m

Equation ( ) AICc ΔAICc AICcWt Cum.Wt LL (4)dist_area* 116.75 0.00 0.43 0.43 -55.83

(6)dist_mx.slope_area 117.81 1.06 0.25 0.68 -56.36

(5)dist_mn.slope_area 118.62 1.87 0.17 0.85 -56.77

(2)dist_mn.slope 120.75 4.00 0.06 0.90 -57.83

(3)dist_mx.slope 121.10 4.35 0.05 0.95 -58.01

(1)dist 121.14 4.95 0.05 1.00 -58.03

Table 4.4. Factors influencing site connectivity of green frogs (Lithobates clamitans) for 2015-2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 2500m

Equation ( ) AICc ΔAICc AICcWt Cum.Wt LL (1)dist* 150.02 0.00 0.94 0.94 -72.47

(3)dist_mx.slope* 155.61 5.58 0.06 1.00 -75.26

(2)dist_mn.slope* 171.03 21.00 0.00 1.00 -82.97

(5)dist_mn.slope_area 173.95 23.93 0.00 1.00 -84.43

(6)dist_mx.slope_area 175.76 25.73 0.00 1.00 -85.33

(4)dist _area 175.79 25.77 0.00 1.00 -85.35

! 118! Table 4.5. Factors influencing site connectivity of chorus frogs (Pseudacris crucifer + P. brachyphona) for 2015-2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 500m

Equation ( ) AICc ΔAICc AICcWt Cum.Wt LL (3)dist_mx.slope* 97.16 0.00 0.26 0.26 -46.03

(1)dist* 97.33 0.18 0.24 0.50 -46.12

(2)dist_mn.slope* 97.48 0.33 0.22 0.72 -46.20

(6)dist_mx.slope_area* 99.09 1.93 0.10 0.82 -47.00

(4)dist _area* 99.20 2.04 0.09 0.92 -47.05

(5)dist_mn.slope_area* 99.43 2.27 0.08 1.00 -47.17

Table 4.6. Factors influencing site connectivity of American toads (Anaxyrus americanus) for 2015-2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 2500m

Equation ( ) AICc ΔAICc AICcWt Cum.Wt LL (1)dist 129.83 0.00 0.19 0.19 -62.37

(3)dist_mx.slope 130.01 0.18 0.18 0.37 -62.46

(4)dist_area 130.14 0.31 0.16 0.53 -62.52

6)dist_mx.slope_area 130. 21 0.38 0.16 0.69 -62.56

(5)dist_mn.slope.area 130.26 0.43 0.16 0.85 -62.58

(2)dist_mn.slope 130.29 0.46 0.15 1.00 -62.60

! 119! Table 4.7. Factors influencing site connectivity of four-toed salamanders (Hemidactylium scutatum) for 2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 1000m

Equation ( ) AICc ΔAICc AICcWt Cum.Wt LL (5)dist_mn.slope_area* 61.33 0.00 0.29 0.29 -28.12

(6)dist_mx.slope_area* 61.34 0.01 0.29 0.58 -28.12

(4)dist_area* 62.17 0.83 0.19 0.78 -28.54

(2)dist_mn.slope* 63.63 2.30 0.09 0.87 -29.27

(3)dist_xn.slope* 63.88 2.54 0.08 0.95 -29.39

(1)dist* 64.88 3.55 0.05 1.00 -29.90

Table 4.8. Factors influencing site connectivity of eastern newts (Notophthalmus viridescens) for 2016. * indicates statistically significant model (p<0.05) using Poisson distribution. α = 2500m

Equation ( ) AICc ΔAICc AICcWt Cum.Wt LL (2)dist_mn.slope* 186.24 0.00 0.91 0.91 -90.58

(3)dist_mx.slope* 190.98 4.74 0.08 0.99 -92.95

(1)dist* 195.72 9.48 0.01 1.00 -95.31

(5)dist_mn.slope_area 242.75 56.51 0.00 1.00 -118.83

(6)dist_mx.slope_area 245.23 58.99 0.00 1.00 -120.07

(4)dist_area 246.11 59.86 0.00 1.00 -120.51

! 120! Table 4.9. Total adult amphibian captures in constructed pools for 2015-2016. *Denotes species omitted from statistical analyses. Numbers in ( ) indicate unique captures for each species. **Denotes species that chorused and oviposited in pools, but were not represented in captures.

Family Species Common Name Total Bufonidae Anaxyrus americanus American toad 70 (62) Hylidae Pseudacris brachyphona mountain chorus frog 66 (57) P. crucifer spring peeper 66 (65) Hyla versicolor/chrysocelis gray treefrog ** Ranidae Lithobates catesbeianus bullfrog *10 (10) L. clamitans green frog 162 (118) L. palustris pickerel frog *4 (4) L. sylvaticus wood frog 96 (83) Ambystomatidae Ambystoma jeffersonianum Jefferson salamander 206 (140) A. maculatum spotted salamander 267 (183) A. opacum marbled salamander *3 (3) Plethodontidae Hemidactylium scutatum four-toed salamander 40 (40) Salamandridae Notophthalmus viridescens eastern newt 124 (52)

! 121! Table 4.10. Matrix with values for distance (m) and mean slope (%) from constructed pools (Cons.) to respective naturally occurring pools at Vinton Furnace State Experimental Forest.

ARN.Poo & Cons.& l& ARS.Pool& Quarry& Beaver1& Powerline& Beaver2& Oxbow& Wolf.Pool& Tiny.Pool& RussellRd& Eagle.Com& Eagle.Pool& & Oak1& distance) 2664& 2959& 2129& 374& 2191& 1855& 1448& 2686& 2922& 2153& 443& 2345& mn)slope) 14& 16& 15& 14& 14& 12& 16& 14& 16& 11& 17& 17& Oak2& & distance) 1661& 1484& 2664& 2959& 2129& 374& 2191& 1855& 1448& 2686& 2922& 2153& mn)slope) 15& 16& 15& 15& 14& 14& 15& 17& 15& 15& 17& 17& Oak3& & distance) 443& 2345& 1661& 1484& 2664& 2959& 2129& 374& 2191& 1855& 1448& 2686& mn)slope) 13& 16& 12& 17& 17& 12& 15& 13& 17& 12& 16& 17&

122 &Oak4& distance) 2922& 2153& 443& 2345& 1661& 1484& 2664& 2959& 2129& 374& 2191& 1855&

! mn)slope) 12& 14& 13& 17& 16& 13& 17& 16& 16& 15& 16& 16& Oak5& & distance) 1448& 2686& 2922& 2153& 443& 2345& 1661& 1484& 2664& 2959& 2129& 374& mn)slope) 16& 15& 13& 15& 14& 9& 11& 12& 15& 12& 14& 13& Oak6& & distance) 2191& 1855& 1448& 2686& 2922& 2153& 443& 2345& 1661& 1484& 2664& 2959& mn)slope) 18& 18& 13& 16& 15& 12& 13& 13& 15& 11& 15& 13& Oak7& & distance) 2129& 374& 2191& 1855& 1448& 2686& 2922& 2153& 443& 2345& 1661& 1484& mn)slope) 17& 18& 12& 16& 16& 11& 15& 14& 17& 10& 17& 18& & continued

! Table 4.10 (continued). Matrix with values for distance (m) and mean slope (%) from constructed pools (Cons.) to respective naturally occurring pools at Vinton Furnace State Experimental Forest.

& Cons.& ARN.Pool& ARS.Pool& Quarry& Beaver1& Powerline& Beaver2& Oxbow& Wolf.Pool& Tiny.Pool& RussellRd& Eagle.Com& Eagle.Pool& & Map1& distance) 2664& 2959& 2129& 374& 2191& 1855& 1448& 2686& 2922& 2153& 443& 2345& mn)slope) 16& 16& 15& 15& 15& 12& 16& 14& 15& 10& 17& 16& Map2& & distance) 1661& 1484& 2664& 2959& 2129& 374& 2191& 1855& 1448& 2686& 2922& 2153& mn)slope) 15& 16& 14& 16& 14& 14& 15& 17& 15& 15& 17& 17& Map3& & distance) 443& 2345& 1661& 1484& 2664& 2959& 2129& 374& 2191& 1855& 1448& 2686& mn)slope) 13& 16& 12& 16& 17& 12& 16& 14& 17& 12& 16& 15&

123 &Map4& distance) 2922& 2153& 443& 2345& 1661& 1484& 2664& 2959& 2129& 374& 2191& 1855&

! mn)slope) 12& 16& 13& 17& 15& 14& 17& 16& 15& 15& 17& 17& Map5& & distance) 1448& 2686& 2922& 2153& 443& 2345& 1661& 1484& 2664& 2959& 2129& 374& mn)slope) 16& 14& 13& 14& 14& 8& 11& 11& 15& 11& 13& 12& Map6& & distance) 2191& 1855& 1448& 2686& 2922& 2153& 443& 2345& 1661& 1484& 2664& 2959& mn)slope) 15& 17& 13& 16& 16& 12& 14& 11& 16& 9& 17& 17& Map7& & distance) 2129& 374& 2191& 1855& 1448& 2686& 2922& 2153& 443& 2345& 1661& 1484& mn)slope) 17& 18& 12& 16& 16& 12& 15& 15& 17& 11& 13& 18& &

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