Influences of Fluvial Geomorphology on Aquatic-To-Terrestrial Hg Transport: Evidence from Protected and Urban Streams of Central Ohio, USA

Influences of fluvial geomorphology on aquatic-to-terrestrial Hg transport: evidence from protected and urban streams of central Ohio, USA

THESIS

Presented in Partial Fulfillment of the Requirements for the Degree Master of Science in the Graduate School of The Ohio State University

Lindsey E. Boaz, A.A.S., B.S.

Graduate Program in Environmental Science

The Ohio State University

2015

Master's Examination Committee:

Dr. Mažeika S.P. Sullivan, Advisor

Dr. Kristin L. Jaeger

Dr. Roman P. Lanno

Copyrighted by

Lindsey E. Boaz

2015

Abstract

Fluvial geomorphology is a driver of sediment supply and storage and may therefore affect the accessibility of sediment-bound mercury (Hg) to aquatic organisms, in addition to influencing aquatic food-web structure via physical habitat controls. In particular, benthic insects that in their larval form are aquatic but emerge into the terrestrial environment as winged adults have been shown to be both highly influenced by stream hydrogeomorphology as well as a vector of contaminant transport from aquatic- to-terrestrial systems. Therefore, the potential for fluvial geomorphology to influence both the aquatic-to-terrestrial export of Hg and the potential use of Hg as an aquatic- terrestrial food-web tracer is significant. I first evaluated Hg as a tracer for trophic structure and dynamics for linked stream-riparian systems under differing fluvial geomorphic conditions within a relatively protected watershed. To that end, I investigated potential differences in Hg dynamics and trophic structure between equilibrium (i.e., stable) and adjusting (i.e., unstable) stream reaches as well as the potential for Hg to model linked aquatic-terrestrial food webs at 20 stream-riparian study reaches in Big

Darby Creek watershed (OH, USA). At each reach, I conducted geomorphic surveys and sampled sediment, water, benthic and emergent aquatic insects, darters (small, insectivorous stream fish), and two riparian spider families (Tetragnathidae and

Pisauridae) for Hg and the stable isotopes 13C and 15N. I provide evidence that fluvial ii geomorphology may influence Hg availability and aquatic-terrestrial transport (via sediment storage and particle size). Although potential exists for the use of Hg as a food- web tracer, these factors – in addition to opportunistic feeding behavior among consumer groups – may mediate its effectiveness. I next evaluated the influences of fluvial geomorphology on aquatic-to-terrestrial Hg export across a gradient of geomorphic characteristics within an urbanized watershed, where the magnitude and ubiquity of geomorphic adjustment may be a controlling factor in Hg dispersal, bioaccumulation, and export across the aquatic-to-terrestrial boundary. In particular, I investigated the relationships between channel geometry, hydraulics, and amount of fine sediment and (1)

Hg loads in sediment, larval, and emergent aquatic insects and (2) aquatic-to-terrestrial contaminant transfer to riparian spiders (Pisauridae and Tetragnathidae). Although fluvial geomorphic characteristics were not related to Hg loads in sediment or larval aquatic insects, I found significant relationships between fluvial geomorphic characteristics and pisaurid spider Hg levels. Fluvial geomorphic characteristics also significantly influenced emergent insect abundance and tended to influence tetragnathid spider Hg levels and emergent insect community composition. Collectively, these results indicate that fluvial geomorphology may be an important player in governing aquatic-to- terrestrial contaminant transport via both sediment dynamics and trophic linkages, although the exact nature of these relationships is likely to differ among systems. In the protected system of my study, fluvial geomorphology also appears to mediate the use of

Hg as an aquatic-to-terrestrial food-web tracer through its effects on Hg accessibility and

iii community structure, the latter of which is evidenced in family-level differences in Hg storage.

Acknowledgements

I would like to thank the funding sources for this project: The Ohio State University,

Ohio Agricultural Research and Development Center; the Ohio Department of Natural

Resources, Division of Wildlife through the USFWS State Wildlife Grant Program; the

Ohio Biodiversity Conservation Partnership; and finally, the Environmental Science

Graduate Program for my fellowship nomination and the opportunity of a lifetime.

Thanks are extended to the Stream and River Ecology Laboratory members for field assistance. Cooperation from Franklin County Metro Parks as well as private landowners enabled the successful completion of this research. I would also like to extend acknowledgment to Dr. Andrea Grottoli and Yohei Matsui at the Stable Isotope

Biogeochemistry Laboratory, Ohio State University, for analysis of basal isotope signatures. Finally, I would like to thank my advisor, Dr. Mažeika Sullivan, for his guidance and tireless patience, and ultimately for the priceless opportunity to have done this research as part of his laboratory.

Vita

June 2004 ...... Northmor High School

2009...... A.A.S. Environmental Science, Safety, and

Health, Zane State College

2011...... B.S. Environment and Natural Resources,

The Ohio State University

2013 to present ...... Graduate Teaching Associate, School of

Environment and Natural Resources, The

Ohio State University

Field of Study

Major Field: Environmental Science

Table of Contents

Abstract ...... ii

Acknowledgements ...... v

Vita ...... vi

List of Tables…………………………………………………………………………...... ix

List of Figures ...... xi

Chapter 1: Background and Literature Review ...... 1

Chapter 2: Effects of fluvial geomorphology on the transport of mercury across stream- riparian boundaries and implications for its use as an aquatic-terrestrial food-web tracer

...... 31

Chapter 3: Influences of fluvial geomorphology on aquatic-to-terrestrial Hg export in small urban streams ………………….……………………………………………….….94

References ...... 148

Appendix A: Study Reach Description ...... 173

Appendix B: Geomorphic Parameters of Study Reaches ...... 175

Appendix C: Hg Levels for Samples from Study Reaches………………….………….180

vii

Appendix D: Stable Isotope Data (including estimates of trophic position and reliance on aquatic primary productivity)……………………………………..……………………184

viii

List of Tables

Table 1.1. Geomorphic and ecological effects of urbanization ...... 8

Table 2.1. Summary of descriptive statistics for geomorphic variables for all 20 Darby

Creek study reaches of Ohio, USA ………………………………...…………………....82

Table 2.2. Summary of descriptive statistics for Hg levels in water, sediment, invertebrates, and fish for all 20 Darby Creek study reaches of Ohio, USA ………....…83

Table 2.3. Summary of descriptive statistics for trophic position and reliance on benthic algal primary productivity of representative consumer groups for all 20 Darby Creek study reaches of Ohio, USA...... 84

Table 3.1. Summary of descriptive statistics for mercury (Hg) levels in abiotic and biotic samples across 12 urban study reaches of Columbus, Ohio, USA……...……………...135

Table 3.2. Principal components analysis (PCA) of hydrogeomorphic and water- chemistry variables……………………………...... ………………………………136

Table A.1. Description of study reaches for Darby Creek and Columbus urban streams………………………………………………………………………………….174

Table B.1. 1st-order hydrogeomorphic parameters for Darby Creek study reaches……176

Table B.2. 1st-order hydrogeomorphic parameters for Columbus urban study reaches..177

Table B.3. 2nd-order hydrogeomorphic parameters for Darby Creek study reaches…...178

Table B.4. 2nd-order hydrogeomorphic parameters for Columbus urban study reaches…………………………………………………………………………………..179

Table C.1. Hg levels of samples from Darby Creek study reaches…………………….181

Table C.2. Hg levels of samples from Columbus urban study reaches………………...183

Table D.1. Stable isotope data for darter samples from Darby Creek………...………..185

Table D.2. Stable isotope data for Pisauridae (2012) samples from Darby Creek...... 186

Table D.3. Stable isotope data for Pisauridae (2013) samples from Darby Creek……..187

Table D.4. Stable isotope data for Tetragnathidae (2012) samples from Darby Creek...188

Table D.5. Stable isotope data for Tetragnathidae (2013) samples from Darby Creek...189

Table D.6. Stable isotope data for Pisauridae (2012) samples from Columbus urban study reaches…………………………………………………………………………………..190

Table D.7. Stable isotope data for Pisauridae (2013) samples from Columbus urban study reaches…………………………………………………………………………………..191

Table D.8. Stable isotope data for Tetragnathidae (2012) samples from Columbus urban study reaches……………………………………………………………………………192

Table D.9. Stable isotope data for Tetragnathidae (2013) samples from Columbus urban study reaches…………………………………………………………………………... 193

List of Figures

Figure 1.1. Example of stream geomorphic setting (top) and its influence on habitat effects (bottom). From Montgomery (1999)...... 6

Figure 1.2. Conceptual illustration of the use of hydrogeomorphology in stream ecology…………………………………………………………………………………….7

Figure 1.3. Conceptual diagram of reciprocal food-web linkages and potential contaminant pathways. From Sullivan and Rodewald (2012)...... 11

Figure 1.4. Big Darby Creek Watershed in central OH, USA ...... 16

Figure 1.5. Map of study reaches in urban Columbus, OH, USA ...... 19

Figure 2.1. Coarse-level field indicators used to categorize adjusting and equilibrium reaches…………………………...……………………………………………………….86

Figure 2.2.Example of paired adjusting and equilibrium RPI reaches in Darby Creek,

Ohio.…………...……...………………………………………………………………….88

-1 Figure 2.3. Simple linear regression of benthic mean Hg (ng g ) levels on D16 and

D50.………………………………………………………………………..……………...90

Figure 2.4. Non-metric Multidimensional Scaling (NMS) ordination plots of benthic community composition data (scaled by variance) and emergent community composition data (scaled by variance)…………………………………………………..……………..91

Figure 2.5. Simple linear regression of mean Hg (ng g-1) levels on mean trophic positions of consumer groups ...... 93

Figure 2.6. Means and ranges of reliance on aquatic primary productivity, trophic positions, and Hg levels of secondary consumer groups ...... 95

Figure 2.7. Simple linear regression of mean Hg (ng g-1) levels on mean reliance on aquatic primary productivity of consumer groups...... 96

Figure 2.8. Log mean Hg (ng g-1) levels for the two most dominant benthic insect families and the two most dominant emergent insect families………………………...... 97

Figure 3.1. Conceptual model of expected fluvial geomorphic influences on in-stream Hg storage and aquatic-to-terrestrial export……..………………………………...... 143

Figure 3.2. Map of 12 study reaches distributed in 1st- and 2nd-order streams of

Columbus, OH, USA...... 145

Figure 3.3. Relationship between tetragnathid spider Hg levels (ng g-1) and proportion of filter feeding trichopterans…………………………………………….……………..146

Figure 3.4. Relationships between PC 1 “Stream Power and Nutrients” and emergent insect abundance and tetragnathid spider Hg...... 147

Figure 3.5. Relationship between PC 3 “Width-to-Depth Ratio” and pisaurid spider Hg levels………………………………………………………...………………………….148

Figure 3.6. Relationship between PC 5 “Dissolved Oxygen” and tetragnathid TP….…149

Figure 3.7. Relationship between tetragnathid spider abundance and abundance of benthic macroinvertebrates………….………………………………………………………..…150

xii

Figure 3.8. Non-metric Multidimensional Scaling (NMS) ordination plots of benthic macroinvertebrate community composition data (scaled by variance) and emergent insect community composition data (scaled by variance).……………………………….……152

Figure 3.9. Relationship between proportion of filter-feeding trichopterans and

D95.………………………………………………..…………………………………….153

xiii

Chapter 1: Background and Literature Review

Introduction

Multiple studies have shown that streams and their adjacent riparian zones are tightly linked through exchanges of carbon (C), organic material, and organisms (Nakano et al. 1999, Wiens 2002, reviewed in Baxter et al. 2005, Fausch et al. 2010). These reciprocal exchanges of energy are essential for functional, healthy ecosystems (Nakano and Murakami 2001, Newham et al. 2011, Sabo and Hagen 2012). Energy in ecosystems is usually assumed to flow from more to less productive habitats, providing a one-way path to the recipient system (Polis et al. 1997, Huxel and McCann 1998). Ecologists have long considered the stream system to be the recipient system for organic matter flow

(Likens and Bormann 1974), effectively dominated by its surrounding valley (Hynes

1975). However, recent research has highlighted the role of reverse flows of energy exchanges from aquatic to riparian zones in providing important trophic subsidies to riparian and terrestrial food webs (reviewed in Baxter et al. 2005). In particular, aquatic insects that emerge from the stream as adults (hereafter, emergent insects) can constitute an important component of the energy budgets of riparian consumers (e.g., Henschel et al. 2001, Nakano and Murakami 2001).

The contribution of emergent insects to riparian and terrestrial food webs suggests that rivers may also function as a source of lateral export of contaminants bioaccumulated 1 in aquatic food webs (Sullivan and Rodewald 2012). To date, however, there are comparatively few studies identifying the potential for contaminant flux from aquatic to terrestrial systems (but see Cristol et al. 2008 and Walters et al. 2008). Although lateral contaminant export has been studied in cases of extreme exposure (e.g., Superfund sites;

Walters et al. 2010) or accidental contamination events (Otter 2013), areas that are less heavily contaminated have not been the focus of attention. However, Hg especially has been shown to be widespread in aquatic systems due to atmospheric deposition, affecting even unindustrialized regions (reviewed in Fitzgerald et al. 1998). Additionally, anthropogenic effects on channel geomorphology in particular have not been extensively considered in mediating lateral fluxes of contaminants, yet prove significant in terms of sediment-bound contaminant loading and distribution within the stream (Rhoads and

Cahill 1999, O’Sullivan et al. 2012) as well as stream-riparian connectivity (Gore and

Shields 1995, Richards et al. 2002). Collectively, these features may affect the trophic linkages that function as aquatic-to-terrestrial contaminant vectors (Burdon and Harding

2008, Paetzold et al. 2008).

Understanding aquatic-to-terrestrial contaminant transfers also has relevance to food-web modeling. The use of contaminants (e.g., trace metals) as dietary tracers in aquatic systems (Soto et al. 2013, Speir et al. 2014), for example, has been proposed as a supplement to the popular approach of stable isotope analysis in modeling complex food- web structure (Peterson and Fry 1987). Stable isotopes offer advantages in efficiency and objectivity over traditional methods such as dietary analysis and feeding observations

(Post 2002), and integrate an organism’s diet over longer time periods than gut-content

2 analysis (Dalerum and Angerbjörn 2005, North et al. 2014, Williams et al. 2014).

However, they can also introduce uncertainty due to poor discrimination between algal- based and detrital-based C (Doucett et al. 1996), within-site variability in both algal based C (France 1996) and primary consumer nitrogen (N) (Anderson and Cabana 2007), and extent of isotopic fractionation among different consumers (Post 2002).

Accordingly, because trace metals can readily accumulate in aquatic sediments for decades (Chang and Cockerham 1994) and be highly accessible to benthic aquatic insects

(Sandor et al. 2001), trace metal distinction between aquatic and terrestrial sources may offer opportunities for model refinement when coupled with stable isotopes (Chouvelon et al. 2014), especially when considered via benthic insects emerging as adults and crossing the aquatic-to-terrestrial border.

Within this context, the influence of stream geomorphology – which has been shown to be a driver of multiple ecological patterns and processes in riverine systems

(Vaughan et al. 2009, Poole 2010, Sullivan 2013) – on emergent insect flux, and in turn, on aquatic-to-terrestrial flux of C and contaminants, remains unresolved. Building on patch dynamics concepts, where the stream system is considered a mosaic of abiotic and biotic processes (Pringle et al. 1988), stream geomorphology has been shown to strongly affect fish community assemblages (Walters et al. 2003, Sullivan et al. 2006) as well as organic matter and C dynamics (Walters et al. 2007, Sullivan 2013). Therefore, the potential for stream geomorphic properties to heavily influence contaminant flux via aquatic emergent insects is significant.

Stream geomorphology

Streams are increasingly viewed as connected to the surrounding landscape rather than embedded separately within it (Hesse and Sheets 1993, Sullivan et al. 2007, Fausch et al. 2010), and can exhibit a template of physical features which rivals the diversity of terrestrial variation and thus can be considered a landscape within its own right (Ward et al. 2002, Wiens 2002). Therefore, stream geomorphology is expected to play a pivotal role in governing the variation of diversity, habitat formation, and nutrient cycling.

At the local scale (i.e., reach 101 m), variations in geomorphology affect biotic habitat via patch dynamic processes (Pringle et al. 1988, Ramirez and Pringle 2001,

Walters et al. 2003, Thorp et al. 2006). Principally, Lane (1995)’s relation of fluvial hydraulics postulates that sediment discharge, bedload particle size, stream flow, and stream slope are characteristic determinants of both catchment-scale (Lu et al. 2006,

Montgomery 1999, Mirus and Loague 2013) and local-scale (Frissell 1986, Montgomery and Buffington 1997) geomorphic conditions. When lotic systems are viewed in the four dimensions of longitudinal, lateral, vertical, and temporal (Ward 1989), these characteristic determinants interact to form local geomorphology that can be studied at various spatial scales. Since the overall geomorphic setting depends on these interactions, anthropogenic influences on water and sediment affect observable geomorphic characteristics (Pizzuto et al. 2000) via alterations to Lane’s relation of fluvial hydraulics (1955). These adjustments are heavily influenced by altered hydrology in urban areas in large part due to increased area of impervious surfaces (Paul and Meyer

2001, Wenger et al. 2009) and associated hydropeaking and erosion (Moglen et al. 2004,

Walsh et al. 2012, Navratil et al. 2013). Urbanization also affects sediment regimes resulting in channel widening (Allmendinger et al. 2007) and streambed incision and aggradation (Ziliani and Surian 2012).

Within the watershed, local geomorphic processes affect disturbance regimes that influence aquatic biota, conceptualized in the Process Domains Concept (Montgomery

1999). The Process Domains Concept provides an alternative perspective to the River

Continuum Concept (Vannote et al. 1980) and patch dynamics notions (e.g., Pringle et al.

1988). Montgomery (1999) suggests that although some physical factors influencing aquatic-riparian ecosystems are continuum-driven, the River Continuum Concept does not take discrete spatial or temporal differences in geomorphic effects into account; additionally, patch dynamics fail to predict or identify areas characterized by different patch-forming processes. Therefore, an understanding of the spatio-temporal geomorphic setting promotes greater understanding of community structure both within and between watersheds (Figure 1.1).

Figure 1.1 Example of stream geomorphic setting (top) and its influence on habitat effects (bottom). From Montgomery (1999).

At the broadest river network perspective, Thorp et al. (2006) suggest that the relationship of patch dynamics to geomorphic processes is nested and hierarchical in nature (Figure 1.2). Additionally, this level of perspective reveals the geomorphic template as predicting areas of concentrated resource exchange, e.g. at tributary junctions and meander bends (Sabo and Hagen 2012). As a result of this broader perspective of stream ecology, fluvial geomorphology can be considered a driving force of biocomplexity (defined as structural and functional phenomena resulting from

6 interactions of biota and the physiochemical environment, Thorp et al. 2006) and is suggested to be a priority for river management interests (Vaughan et al. 2009). Sidle and Onda (2004) note this “hydrogeomorphology” as a multidisciplinary interest and predict research challenges that will extend throughout the next century.

Hydrology Geomorphology

Habitat structure (flow, substrate size) determining biotic community

Potential for biogeochemical cycling

Ecology

Figure 1.2. Conceptual illustration of the use of hydrogeomorphology in stream ecology.

Urban Streams

Common impacts to streams draining urban areas have been conceptualized in the

“urban stream syndrome” (Meyer et al. 2005), the negative effects of which are not limited to the immediate urban area (Grimm et al. 2008). Adding to the complexity is the variability in fluvial geomorphic impacts based on the degree of urbanization (Poff et al.

2006). Walsh et al. (2005) note that although most studies attempt to correlate in-stream ecological metrics with total catchment imperviousness, variation in these relationships highlight the need for further research in the underlying mechanisms of urban–ecological processes. Examples of the effect of urbanization on stream geomorphology and ecology follow (Table 1.1):

Table 1.1. Geomorphic and ecological effects of urbanization.

Geomorphic effects Ecological effects - Altered hydrography, i.e., high - Reduction of retentive structures flows that peak more rapidly (e.g., large organic debris) for (“flashiness”) (Paul and Meyer periphyton, resulting in reduced 2001) primary production (Coleman and Dahm 1990) - Increased frequency in bankfull - Changes in the stream’s typical discharge may result in channel regime of organic matter widening and increased decomposition due to increased sediment discharge from eroded flows and influx of banks to downstream (Hammer contaminants (Chadwick et al. 1972, Arnold et al. 1982, 2006) Trimble 1997) - Altered channel stability - Decreased rate of nutrient resulting in differing potentials uptake from the water column for bank restabilization by benthos (Meyer et al. 2005) (Henshaw and Booth 2000) - Often reduced (Konrad and - Water table lowered below Booth 2005) but sometimes microbially active soil zones augmented (Lerner due to incision (Groffman et al. 2002) baseflows 2003) - Bed sediment instability - Frequent algal scour by high (Uehlinger et al. 2002) flows (Murdock et al. 2004)

The differences in stream geomorphology between urban and forested streams can be significant (reviewed in Paul and Meyer 2001, Wenger et al. 2009). Development of

8 the watershed in an urban area produces significant changes to stream geomorphology

(Hammer 1972, Arnold et al. 1982, Walsh et al. 2005, Allmendinger et al. 2007) that in turn influence biotic processes. Research suggests that aquatic larval insects selectively choose microhabitats (Rae 2013) and adjust their behavior to the hydraulic environment produced by local geomorphology (Oldmeadow et al. 2010). For example, Ramirez and

Pringle (2001) observed changes in macroinvertebrate drift composition corresponding to a break in geomorphic gradient. Brussock and Brown (1991) demonstrated that physical habitat variations at the reach level control macroinvertebrate distributions in a manner that obscures the predictions of the longitudinally-based River Continuum Concept.

As an example of the combined effects of stream geomorphic variation and anthropogenic modification on biotic communities, Gurtz and Wallace (1984) showed that finer (and thus more mobile) patches of channel substrate experienced the greatest reduction in macroinvertebrate populations from disturbance due to logging. In addition to reducing benthic density and diversity (Lenat 1981, Waters 1995, Descloux et al.

2013), fine sediments have a greater surface area for contaminant accessibility than coarse particles (Inoue 1980) and may store more trace metals than the overlying water

(Driscoll et al. 1995, Wang et al. 2004). In general, lower velocity and discharge do not mobilize sediment as do greater flows (Adenlof and Wohl 1994, Lisle 1995); therefore, reduced baseflows common to many urban streams (Konrad and Booth 2005) may contribute to increased contaminant access via sediments. Consequently, relationships linking stream geomorphic characteristics to sediment characteristics and aquatic larval

9 assemblages suggest that geomorphic variability will also influence aquatic emergent insects, and in turn, aquatic-to-terrestrial contaminant transport.

Contaminants and aquatic-terrestrial linkages

Fitzgerald et al. (1998) suggested that the increase of mercury (Hg) levels in aquatic environments could be attributed to atmospheric deposition from remote industrialization, rather than natural background sources. These industrial species of Hg have variable deposition and dispersal patterns, and are notorious for bioaccumulating and subsequently biomagnifying in aquatic biota such as periphyton (Brooks et al. 2012), macrophytic vegetation (Gentès et al. 2013), aquatic insect larvae (Azevedo-Pereira et al.

2012), and fish (Wren and MacCrimmon 1986). Riparian bird species are affected as well, as evidenced by Hg bioaccumulation from fish to ospreys (Hakkinen and Hasanen,

1980) and insects to tree swallows (Hallinger et al. 2011). Biomagnification up the freshwater food chain in this way may in some cases result in toxic levels in higher-level consumers, including humans (Driscoll et al.2007). Cabana et al. (1994) concluded that longer food chains result in higher levels of Hg toxicity than shorter food chains, which poses special risk to humans on account of the tendency to consume top-level predators.

Fluvial geomorphology also exhibits potential to influence Hg bioaccumulation through storage effects in sediment, the accessibility of which has been shown to elevate aquatic insect body loads (Richman and Milani 2010, Sinclair et al. 2012). These effects make

Hg of worthwhile consideration in a study of geomorphic impacts on aquatic contaminants.

Recently, increasing attention has been paid to the potential transport of contaminants from aquatic-to-terrestrial systems through the energetic pathways that link these two systems (Sullivan and Rodewald 2012, Figure 1.3). For example, Walters et al.

(2008) showed a relationship between aquatic insects and the export of toxic-level PCBs to riparian spiders. However, the relationship extends beyond a single predator and prey;

Northam et al. (2012) traced Hg from aquatic mayflies to spiders to riparian songbirds.

The intensity of contaminant effects can be context-dependent, as demonstrated by Jones et al.’s (2013) study of bottom-up versus top-down controls on insect-mediated mercury flux.

Figure 1.3. Conceptual diagram of reciprocal food-web linkages and potential contaminant pathways (e.g., energy flows as represented by arrows) in a stream-riparian ecosystem. Arrows represent energy-flow pathways. Solid arrows represent strong pathways (e.g., primary predator-prey relationships); dashed arrows represent weaker pathways (e.g., occasional or facultative predator-prey relationships). Thicker arrows represent aquatic-to-terrestrial fluxes of emergent aquatic insects. From Sullivan and Rodewald (2012). 11

Because trace metals can readily accumulate in aquatic sediments, becoming highly accessible to benthic insects (Sandor et al. 2001) that emerge to the riparian zone, analysis of trace metals such as Hg might be particularly useful in tracing aquatic-to- terrestrial energy fluxes (Cristol et al. 2008, Walters et al. 2008, Sullivan and Rodewald

2012). Aquatic sediments can act as decades-long reservoirs for trace metals (Chang and

Cockerham 1994), increasing the utility of Hg as a food-web tracer in linked aquatic- terrestrial systems. A combined approach investigating the influence of fluvial geomorphology on Hg accessibility and the biotic community within the contexts of contaminant export and food-web tracing will therefore contribute to better understanding these aquatic-terrestrial linkages and thus the fluvial ecosystem as a whole.

Describing and quantifying aquatic-terrestrial food webs

Given their intimate physical, biological, and chemical connections with the adjacent landscapes, stream and river food webs can be highly complex (Calizza et al.

2012, Parker and Huryn 2013, Ishikawa et al. 2014; see also Polis and Strong 1996).

Stream-riparian food webs incorporate a suite of taxa from both aquatic and terrestrial environments that cross the land-water interface as part of their foraging and life-cycle activities and can rely heavily on energy subsidies [aquatic-to-terrestrial energy fluxes, and vice versa (Nakano and Murakami 2001, Fausch et al. 2010)] from adjacent habitats.

Adaptive foraging behavior of both high- and low-trophic-level consumers (reviewed in

Valdovinos et al. 2010) and regional species diversity of the players within trophic levels

(Villegér et al. 2008, Baiser et al. 2013) contribute additional complexity to the challenge

12 of accurately modeling food webs. In some contexts, consumers may feed randomly on prey rather than preferentially ((Brind’Amour and Dubois 2013), adding another level of difficulty in tracing a clear trophic pathway.

Stable isotope analysis has emerged in recent decades as a valuable tool in describing complex food-web structures (Peterson and Fry 1987). Use of the commonly encountered stable isotope ratios of carbon (δ13C) and nitrogen (δ15N), in particular, have been widely applied ecologically to more fully resolve spatial and temporal trophic interactions within food webs (reviewed in Layman et al. 2011). In general, the isotopic ratios [of heavy (15N or 13C) to light (14N or 12C) isotopes] of a consumer reflect its dietary sources while giving an indication of trophic position, allowing quantification of an organism’s food sources and the origin of nutrients for primary producers. These

0 ratios are expressed in terms of parts per thousand ( /00) deviation from an established standard (e.g., Pee Dee belemnite limestone for δ13C; atmospheric nitrogen for δ15N), as follows:

13 15 δ C or δ N = [(Rsample/Rstandard) – 1] × 1,000 where R = 13C/12C or 15N/14N.

A common application for use of stable isotopes in depicting trophic interactions within food webs is the estimation of trophic position (TP). Consumer body tissue differentially uptakes isotopically-heavy versus light N, resulting in fractionation of nitrogen isotopes and allowing for δ15N to be used as a continuous measure of trophic position as follows:

15 15 TP = λ + (δ Nconsumer – δ Nbase)/Δn

13 where λ = trophic position of the selected isotopic baseline and Δn = fractionation rate of

0 N (generally assumed to be constant at 3.4 /00 increase across each trophic level, Post

2002).

An additional component of stable isotope analysis involves quantifying the sources of energy (e.g., aquatic versus terrestrially derived) supporting a given consumer.

To partition the relative contribution of different resource pools, stable isotope mixing models have been developed based on the isotopic signatures of resources compared with their presence in the tissue of consumers. The Bayesian family of mixing models

(Parnell et al. 2010) is particularly effective as it attempts to incorporate uncertainty in model parameters (e.g., source pool variation and variation in trophic fractionation) which reflects variability inherent to natural systems. Mixing models are helpful in food webs such as linked aquatic-terrestrial systems, where consumers may acquire energy from separate sources with distinct sets of basal resources (e.g., aquatic versus terrestrially derived). The TP model needs adjustment to accommodate such distinction, as follows (Post 2002):

15 15 15 TP = λ + (δ Nconsumer – [δ Nbase1 × α + δ Nbase2 × (1 – α)]/Δn where α = the proportion of N acquired from baseline one. If it is assumed that the movement of C and N through the food web is similar, a simple 2-end-member linear mixing model is sufficient to estimate α:

13 13 13 13 α = (δ Cconsumer – δ Cbase2)/(δ Cbase1 – δ Cbase2).

For these analyses, software packages such as SIAR (Stable Isotope Analysis in

R; Parnell and Jackson 2013) are effective at handling variability in sources using

Bayesian mixing models. Outputs generated from Bayesian mixing models are probability distributions, the results of which can be used in subsequent analyses (Parnell et al. 2010, Araujo et al. 2011).

Study Systems

Big Darby Creek is designated by the Ohio Department of Natural Resources as a

State and National Scenic River, draining approximately 1450 km2 of central Ohio from its headwaters in Logan County, OH to its confluence with the Scioto River in Pickaway

County, OH (Figure 1.4). Coarse glacial deposits of gravel and cobble are commonly distributed in the streambed; combined with cool-temperature, groundwater-fed headwaters tributaries and the natural stream gradient, Big Darby Creek contains extensive stretches of designated Exceptional Warmwater Habitat (OEPA 2006).

Although the stream is known to provide habitat for several state and federally listed endangered species, Ohio EPA has identified threats to some stream segments as runoff from urban areas (e.g., Columbus to the northeast) and agricultural land (~ 73% of watershed land use), combined with poor stream bank management (OEPA 2006). Thus,

Big Darby Creek provides an opportunity to study Hg as a food-web tracer within the context of a protected system not heavily impaired by Hg contamination and with a diverse array of biota, yet still experiencing instability in fluvial geomorphology along some segments.

! !

! !!! ! !

! !! !

0 25 50 100 Kilometers

Figure 1.4. Big Darby Creek Watershed in central OH, USA.

By contrast, reaches located within 1st- and 2nd- order tributaries of the Olentangy

River, Scioto River, and Blacklick Creek are heavily modified urban systems (Figure

1.5). The Olentangy River is located in central Ohio, spanning approximately 150 km from Crawford County to its confluence with the Scioto River in downtown Columbus.

Although a portion of the Olentangy (north of Columbus) has been designated as a State

Scenic River by the Ohio Department of Natural Resources, the Ohio EPA cites problems primarily due to point source pollution, runoff from urban areas, and poor stream bank

16 management including channelization (OEPA 2007). The Ohio EPA (2007) also notes changes in hydrologic regime by destabilization of channels, resulting in greater sedimentation in downstream reaches and increasingly intermittent tributary stream flow

(of which our 1st- and 2nd- order study reaches are considered a part). Of note, 13 of 18 sites assessed in the Olentangy Tributaries Headwater group do not achieve Ohio EPA’s scores for aquatic life use (OEPA 2007). At the time of this writing, Ohio EPA’s Total

Maximum Daily Load (TMDL) report for the Scioto River was still in progress; however, study reaches considered tributaries to the Scioto River are expected to exhibit similar features (e.g., management activity, sources of impairment) as those to the Olentangy

River, given their relative proximity within the same urban area. Although the Scioto

River is the receiving stream for the Olentangy River and is therefore a higher-order stream, all tributaries chosen were 1st - or 2nd - order streams, making them of comparable stream order across watersheds of differing size.

Blacklick Creek is located within the Big Walnut Creek watershed, conjoining (in addition to Alum Creek) with Big Walnut Creek at Three Rivers Park on the east side of

Columbus. This area is located in a sub-ecoregion of level- to rolling-terrain till plains

(OEPA 2005). The Ohio EPA (2005) lists impairments to Blacklick Creek as pathogens, nutrient enrichment, and siltation; sources of impairment include home sewage treatment systems, minor municipal point sources, contaminated sediments, and land development.

Although still located within urban Columbus, study reaches within tributaries to

Blacklick Creek contain ~ 28 km (out of a total ~ 45 km) in attainment status for Ohio

EPA’s designated Warmwater Habitat (OEPA 2005; although distinct from Exceptional

Warmwater Habitat of Big Darby Creek, OEPA 2006). Therefore, differences in physical habitat and physiochemical attributes exist among reaches chosen in tributaries to Blacklick Creek, Olentangy River, and the Scioto River, though all tributaries are contained under the designation of largely impaired urban stream systems.

Figure 1.5. Map of study reaches in urban Columbus, OH, USA. The 12 reaches are designated with blue dots.

Summary and Objectives

The overarching goal of this research was to investigate fluvial geomorphic effects on Hg export via aquatic-terrestrial trophic linkages, both as a potential food-web tracer and a vector of contamination. Although stable isotopes have proven effective food-web tracers in many instances, limitations exist for their use, which warrant consideration of supplemental tracing methods (Soto et al. 2013). Trace metals have been suggested as an aquatic food-web tracer (Soto et al. 2011); however, effective use of these substances in food-web modelling is not fully understood. Understanding the influence of fluvial geomorphology on Hg accumulation and aquatic-to-terrestrial transport is critical to understanding its use as a potential food-web tracer, as well as its export as a contaminant to the riparian zone. Within this framework I have developed two main research objectives, each of which corresponds to a chapter within this thesis and represent a manuscript for future publication.

Objective 1: Investigate the influences of stream geomorphic condition [i.e., adjusting reaches (unstable, characterized by current or recent changes in sediment size distribution and/or channel geometry) vs. reaches in dynamic equilibrium (stable, characterized by minor or no changes in sediment or channel dimensions)] on Hg dynamics and the potential utility of Hg as a food-web tracer for 20 linked stream-riparian food webs.

Objective 2: Explore the relationship between fluvial geomorphology (using a suite of parameters related to sediment size and distribution, channel geometry, and stream

19 power) and aquatic-to-terrestrial Hg export via linked energetic pathways in 12 small urban streams, considering mechanisms related to trophic position and Hg levels of riparian consumers (Tetragnathidae and Pisauridae), benthic and emergent insect community composition and Hg levels, and abundances of insects and Tetragnathidae spiders.

I anticipate that this work will contribute to understanding the potential for use of

Hg as a food web tracer in a protected aquatic-terrestrial system. Although stable isotopes have proven effective in modeling food webs over past decades, trace metals have potential to more fully resolve trophic relationships in some situations, with Hg as a particularly ubiquitous tracer candidate within this context. Of the studies that have been conducted in this vein, most have focused primarily on areas of high contamination (e.g.,

Superfund sites, accidental spills) rather than a protected area as in this study.

Additionally, I anticipate this work will contribute to more fully understanding effects of fluvial geomorphology on contaminant export within small urban streams. Although urban streams are known to experience elevated contamination levels relative to their more natural counterparts, traditional management and restoration actions have primarily focused on physical habitat structuring without explicit consideration of how contaminants may be laterally exported from the stream. Managers of urban streams will therefore benefit from a better understanding of fluvial geomorphic effects on aquatic-to- terrestrial contaminant transfer, and how management decisions affecting fluvial geomorphology may have implications for contaminant export from the stream.

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Chapter 2: Effects of fluvial geomorphology on the transport of mercury across stream- riparian boundaries and implications for its use as an aquatic-terrestrial food-web tracer

Lindsey E. Boaz1,2, S. Mažeika P. Sullivan1, and Kathleen A. Hossler1

1School of Environment & Natural Resources, The Ohio State University, 2021 Coffey

Rd., Columbus, OH 43210

2Corresponding author email: [email protected]

Abstract

Fluvial geomorphology, as a driver of sediment supply and storage, may affect both aquatic food-web structure as well as the accessibility of sediment-bound mercury

(Hg) to aquatic organisms. In particular, benthic insects that in their larval form are aquatic but emerge into the terrestrial environment as winged adults have been shown to be both highly influenced by stream hydrogeomorphology as well as a vector of contaminant transport from aquatic-to-terrestrial systems. Therefore, the potential for fluvial geomorphology to influence both the bioaccumulation of Hg as well as its transport across the aquatic-terrestrial boundary may be significant, yet remains unexplored. We hypothesized that sediment-bound Hg accessibility and trophic interactions would differ between stream reaches in geomorphic dynamic equilibrium

(i.e., stable) and stream reaches undergoing geomorphic adjustment (i.e., unstable), leading to differences in the magnitude of aquatic-terrestrial Hg transport. In turn, we expected that fluvial geomorphology would mediate the use of Hg as a food-web tracer for linked stream-riparian food webs. At 20 stream-riparian study reaches in Big Darby

Creek watershed (OH, USA), we conducted geomorphic surveys and sampled sediment, water, benthic and emergent aquatic insects, darters (small, insectivorous stream fish), and two riparian spider families (Tetragnathidae and Pisauridae) for Hg and the stable isotopes 13C and 15N. We provide evidence that fluvial geomorphology may influence

Hg availability and aquatic-terrestrial transport (via sediment storage and particle size).

Specifically, we found that benthic insects showed significantly higher mean Hg levels at adjusting than at equilibrium reaches, and that measures of fine sediment (D16) and

32 median particle size (D50) were strong negative predictors of benthic insect Hg. This work also suggests that influences of fluvial geomorphology might mediate use of Hg as an aquatic-to-terrestrial food-web tracer.

Introduction

Food webs, which describe the network of trophic interactions through which energy flows within an ecosystem and include properties such as species composition and connectivity (Pimm 1984, Polis and Strong 1996), offer insight into the structure, function, and stability of ecosystems (Thébault and Loreau 2003, Sabo et al. 2009,

Thompson et al. 2012, Alva-Basurto and Arias-González 2014). Food-web structure and composition can be highly context dependent, suggesting that environmental variability is a key determinant (Weinstein et al. 2005, Schaal et al. 2009, Royan et al. 2013). In streams, ecosystem size, disturbance, and resource availability have been cited as critical environmental determinants of food-web descriptors such as food-chain length (Sabo et al. 2009, Takimoto et al. 2012). Fluvial geomorphology has also emerged as a potentially important factor driving stream food-web dynamics (Sullivan 2013, Bellmore and Baxter

2014, Christian and Allen 2014). Montgomery (1999), for example, suggested that local- scale geomorphic processes may determine habitat and disturbance regimes that influence stream communities (i.e., Process Domains Concept). Walters et al. (2003) found that geomorphic variables best predicted stream fish assemblages. Sullivan (2013) found that stream-reach morphology, including features such as substrate size and sediment storage elements (e.g., large wood, pools), was an important predictor of the assimilation and distribution of C in food webs of mountain streams of Idaho.

Mercury is a widespread trace metal, of which its organic form – methylmercury

(MeHg) – is a known bioaccumulant in aquatic food webs (Chen et al. 2005, Clayden et al. 2013, Lavoie et al. 2013). Methylmercury was shown by Cabana et al. (1994) to

34 relate to trophic position in lakes, as its slower rate of elimination relative to uptake is similar to the preferential retention of 15N over 14N at higher trophic levels. Total mercury (hereafter, “Hg”) is often easier to measure as it does not require further speciation and can be used as a proxy for MeHg. Studies of insectivorous birds and amphibians, for example, have shown the majority of total Hg concentrations in blood to be MeHg [e.g., > 95% (Evers et al. 2005, Rimmer et al. 2005) and ~ 71% (Bergeron et al.

2011), respectively]. Similarly, MeHg comprises a high proportion of total Hg in aquatic emergent insects (Haines et al. 2003, Chételat et al. 2008, Buckland-Nicks et al. 2014).

Stream geomorphology may influence Hg movement through food webs via multiple mechanisms. Firstly, sediment supply versus transport, which is an underpinning of many stream geomorphic classification systems (Frissell et al. 1986,

Montgomery and Buffington 1997, Sutfin et al. 2014), controls overall sediment storage and may therefore affect the bioavailability of sediment-bound Hg to aquatic organisms

(Boudou and Ribeyre 1997). Within this context, fluvial geomorphic influence at the reach scale is particularly important (Hassan et al. 2005, Dubinski and Wohl 2007, Eaton and Church 2011). Streams can also be characterized by channel dimensions and derivatives (e.g., width-to-depth ratio, of note to this study), streambed gradient, and particle size distribution (Rosgen 1994), with relevance for sediment aggradation and associated Hg storage (Miller et al. 1998). Secondly, the relationships between carbon assimilation and fluvial geomorphology (e.g., Walters et al. 2007, Sullivan 2013) may implicate channel morphology in the distribution and assimilation of Hg into stream food webs.

Thirdly, geomorphically-driven shifts in the instream biotic community can influence food-web structure via biotic assemblages (Sullivan et al. 2004, 2006, Walters et al. 2007, Sullivan and Watzin 2009) and hence trophic accumulation and transfer of

Hg. Channel morphology may also influence transfers of energy and contaminants across the aquatic-terrestrial boundary (see Sullivan and Rodewald 2012). For instance, lateral channel adjustment that alters the aquatic-terrestrial interface may shift stream-riparian connectivity (Sullivan et al. 2006) and thus affect transfer of Hg to the riparian zone. In

Vermont, geomorphically stable reaches were associated with higher in-stream habitat condition and a greater relative abundance of sensitive macroinvertebrate taxa than their unstable counterparts (Sullivan et al. 2004), which could have implications for Hg transport to riparian predators through differences in aquatic emergent insect flux between stable and unstable reaches (although emergence was not measured in this study).

Food webs are increasingly used to link stream and riparian ecosystems through the energetic pathways that cross aquatic-terrestrial boundaries (e.g., reviewed in Baxter et al. 2005, Sullivan and Rodewald 2012). Over the past few decades, stable isotope analysis has emerged as a powerful tool to model complex food-web structure (Peterson and Fry 1987), including in linked stream-riparian systems (e.g., Scharnweber et al. 2014,

Paradzayi and Sullivan 2015). The stable isotope ratios of carbon (δ13C) and nitrogen

(δ15N), in particular, have been widely applied ecologically. In general, the isotopic ratios of a consumer reflect its dietary sources while giving an indication of trophic

36 position, allowing quantification of an organism’s food sources and the origin of nutrients for primary producers.

Nonetheless, limitations of stable isotope analysis exist. In aquatic systems, these can include poor discrimination between algal-based and detrital-based C (Doucett et al.

1996), within-site variability of algal based C (France 1996), within-site differences between baseline primary consumer N (Anderson and Cabana 2007), and variability in extent of isotopic fractionation among different consumers (Post 2002). These limitations suggest the use of additional and/or alternative food-web tracers to more fully resolve trophic relationships (Soto et al. 2013).

Whereas Hg has been used as a proxy for exposure risk of predator from prey

(Casazza et al. 2014) and young from parent organism (through conversion of tissues to embryonic material, Lyons et al. 2013, Saxton et al. 2013), it has not been extensively considered as an energy tracer across the aquatic-terrestrial boundary (but see Cristol et al. 2008, Walters et al. 2008, Sullivan and Rodewald 2012), yet considerable potential exists for its use as a dietary tracer (Stewart et al. 2008, Speir et al. 2014). Organisms may be exposed to trace metals through environmental media (e.g., water, sediment, atmospheric deposition) and/or diets, resulting in food-web implications for biomonitoring efforts such as greater bioaccumulation in more productive systems

(Ofukany et al. 2014) and taxon-dependent strengths of monitoring particular contaminants (Soto et al. 2011). In addition to biomonitoring, trace metal accumulation may offer opportunities for model refinement when coupled with stable isotopes

(Chouvelon et al. 2014). Further, because trace metals can readily accumulate in aquatic

37 sediments and be highly accessible to benthic aquatic insects (Sandor et al. 2001) that emerge to terrestrial systems as winged adults (hereafter “emergent insects”), analysis of trace metals might be particularly useful in tracing aquatic-to-terrestrial energy fluxes

(Sullivan and Rodewald 2012).

Although the use of trace metals, including Hg, as aquatic food-web tracers in conjunction with stable isotopes has considerable promise, it is not without limitations.

As with isotopes, adequate discrimination among trace metal levels in food sources must still exist in order for trophic levels to be distinguished (Soto et al. 2013). Cabana et al.

(1994) caution that within a complex aquatic food web, it may be difficult to differentiate variation in biomagnification through food chains of variable length from variation in bioavailability at the base of the food chain. Additionally, differences in biotic retention may be taxon- (Cid et al. 2010, Buckland-Nicks et al. 2014, Kraus et al. 2014) or life stage- (Sarica et al. 2005, Cid et al. 2010, Kraus et al. 2014) dependent. As such, the effectiveness of Hg as an aquatic food-web tracer may be directly tied to variability in local community composition. However, aquatic sediments can act as decades-long reservoirs for trace metals (Chang and Cockerham 1994), aiding their distinction from terrestrial measurements and increasing the utility of Hg as a food-web tracer in linked aquatic – terrestrial systems.

In this study, we investigated the influences of stream geomorphic condition

[adjusting reaches (i.e., unstable, characterized by current or recent changes in sediment size distribution and/or channel geometry) vs. reaches in dynamic equilibrium (i.e., stable, characterized by minor or no changes in sediment or channel dimensions)

38 following Sullivan et al. 2004 and Sullivan et al. 2006] on Hg dynamics and the potential utility of Hg as a food-web tracer in linked stream-riparian food webs. We hypothesized that (1) adjusting reaches would lead to greater Hg availability to aquatic organisms via greater sediment storage and smaller particle size, resulting in elevated Hg levels in biota relative to equilibrium reaches; additionally, sedimentation at these reaches would influence community structure and therefore, trophic interactions and Hg transfer from the aquatic to the terrestrial zone and (2) variability in trophic interactions, including reliance on nutrition derived from primary production (i.e., benthic algae),would mediate the capacity of Hg to describe food-web structure across the aquatic-terrestrial boundary, as reflected in consumer bioaccumulation patterns. For ten pairs of adjusting-equilibrium stream reaches (i.e., 20 reaches total) in the Darby Creek watershed of central Ohio,

USA, we used the naturally-abundant stable isotopes 13C and 15N as a reference method to assess trophic position of and the contribution of aquatic primary production (e.g., benthic algae) to stream-riparian consumers (darters, riparian spiders of the families

Pisauridae and Tetragnathidae). We complemented this approach with Hg analysis of water, sediment, and food-web consumers (benthic larval insects, aquatic emergent insects, darters, Pisauridae, and Tetragnathidae) as an exploratory model of food-web structure.

Material and Methods

Study reaches and categorical stability assessments

We selected twenty study reaches in Big Darby Creek, a National and State

Scenic River draining 1,441 km2 of central-southern Ohio, USA. Following a paired- study design, we selected ten reaches representing geomorphic dynamic equilibrium

(stable) riffle-pool interfaces (RPIs) and ten reaches that represented adjusting (unstable)

RPIs. In the field, coarse-resolution indicators of channel stability were used to assign

RPI geomorphic condition using a suite of indicators of channel adjustment. Stream reaches exhibiting evidence of adjustment (e.g., degradation, aggradation, over-widening, changes in planform) were categorized as “adjusting”, while reaches exhibiting no or minor signs of adjustment were categorized as “equilibrium” (Figure 2.1; see also Bey and Sullivan 2014). Following Sullivan et al. (2006), channel degradation/incision was inferred from exposed till, fresh substrate in the stream bed, recently abandoned terraces along banks, fresh vertical faces along banks, and an inactive (or poorly connected) floodplain. A high degree of embeddedness, loss of diversity in flow and depth regimes, high sediment deposition, incomplete riffles, increased channel bed exposure during low flow, partial riffles and/or runs, presence of mid-channel or diagonal bars, and high width-to-depth ratio were used as evidence for aggradation. We used fallen trees, many bank overhangs, erosion on both sides of the banks, and deposition of mid-channel bars as indicators of channel widening. Channel avulsions, flood chutes or chute cut-offs, formation of islands, partial or transverse riffles and/or runs, and newly formed channel bars were used as principal evidence for changes in planform. In some cases, (e.g., high width-to-depth ratio and mid-channel bars for both widening and aggradation), one

40 feature suggested more than one adjustment process. An example of an adjusting- equilibrium RPI pair is given in Figure 2.2.

Each RPI constituted a study reach and consisted of the flow sequence from the top of the riffle to ~ 5 m into the downstream pool, representing critical habitat for darters

(Matthews 1985, Walters et al. 2003, Rashleigh et al. 2005) and benthic macroinvertebrates (Feminella 1996, Halwas et al. 2005, Beauger et al. 2006). Study reaches spanned ~ 100 m on average from upstream to downstream. Reaches were selected such that each pair was located within the same larger stream segment to minimize differences in the near-shore zone (10 – 15 m), water quality, and drainage area to the degree possible.

Following study-reach selection, coordinated surveys of quantitative hydrogeomorphic measurements were conducted during the summer and early autumn of

2011 and 2012. Fish assemblage data for a companion study in the same system (Bey and

Sullivan 2014) were collected at this time. Tetragnathid spider abundance surveys were conducted, along with collection of tetragnathid and pisaurid spider tissue samples, sediment, water, and baseline stable isotope samples (periphyton and detritus) in the summer and early autumn of 2012 and 2013. Benthic and emergent insect communities were sampled in the summer and early autumn of 2013.

Quantitative hydrogeomorphic surveys

Following initial coarse adjusting vs. equilibrium characterizations of RPIs, quantitative hydrogeomorphic assessments were conducted for each of the 20 stream

41 reaches following procedures outlined in Cianfrani et al. (2004). We investigated five parameters that would potentially indicate aggradation and, consequently, Hg access and habitat alteration due to particle retention: width-depth ratio (mean bankfull width ÷ mean bankfull depth), D16 and D50 (particle size for which 16% and 50% of the particles are finer, respectively, with D16 used as a measurement of fine sediment), velocity, and discharge. Bankfull width (m), cross-sectional area (m2) and mean depth (m) were measured at two lateral transects across each RPI, then averaged per RPI. Bed grain size was estimated for each RPI by Wolman’s pebble count method (1954), with 100 randomly selected clasts measured at each of the two lateral transects using a gravelometer. Estimates for velocity (m s-1) and discharge (m3 s-1) were generated using the Reference Reach Spreadsheet developed by Mecklenburg (2006) from field measurements of channel slope, bankfull width, cross-sectional area, mean depth, and bed grain size (see Appendix B for geomorphic formulae). Channel slope (m m-1) was determined from longitudinal surveys of each RPI, using a precision laser (LL300

SPECTRA Self Leveling Laser, Trimble Construction Tools Division, Dayton, Ohio).

From these data, fluvial geomorphic parameters selected for specific analysis were found to be different between adjusting and equilibrium RPIs, with the exception of D16.

Adjusting RPIs had significantly higher width-to-depth ratios (p = 0.047) and trended towards lower D50, velocity, and discharge (p = 0.067, 0.055, and 0.067, respectively,

Table 2.1).

Sediment and water samples

A sediment core of ~ 10 cm and a grab water sample were collected at each RPI for Hg analysis. Both water and sediment samples were composited from several locations across the reach to capture reach-scale variability. Sediment was frozen in plastic sleeves before processing. Water samples were unfiltered as sample degradation was considered of less risk to samples than contamination by filtration, given that water samples were sent for analysis within 24 h of collection.

Biotic surveys and samples

Pyramidal emergence traps were used to collect aquatic emergent insects following Alberts et al. (2013). Traps covered a stream area of 1 m2 and were constructed of a PVC frame with square base, 18 × 16-mm mesh side panels and a collection jar at top. Two traps were set at each reach, at the top and bottom of each riffle, with sample collections made at five and ten days after deployment. Larval benthic aquatic invertebrates were collected at the top and bottom of each riffle, using a 600-cm2

Surber sampler with a 500-µm-mesh net. In the laboratory, larval and emergent insects were subsequently identified to family using Merritt, Cummins, and Berg (2008) and enumerated. Fish assemblage data as well as darter tissue samples (3 – 5 individuals per reach from Etheostoma blennioides and Etheostoma caeruleum) were collected during a companion study at the same study reaches (Bey and Sullivan 2014). Specimens were frozen until processing.

Surveys of riparian tetragnathid spiders (long-jawed orb weavers) were conducted along two 15-minute transects at each RPI (one on each bank) from ~ 9:00 pm – 12:00

43 am following Meyer and Sullivan (2013) and focused on webs within ~ 1 m of the stream edge and up to ~ 2 m in height (Williams et al. 1995). Tetragnathidae (8 – 10 individuals) and Pisauridae (fishing spiders, 2 – 3 individuals) were also collected for stable isotope and Hg analysis.

Periphyton (i.e., benthic algae) and detritus samples were collected to establish baseline isotopic signatures. Periphyton was collected using a nylon brush from fifteen rocks selected at random from each riffle (Finlay et al. 1999). Detrital samples were collected by hand and frozen in plastic sleeves until processing. Coarse particulate matter

(CPOM) was removed from periphyton samples, while fine particulate matter (FPOM; particles <1 mm) was sieved from detritus samples.

Stable isotope analysis

Tissue samples from darters, Tetragnathidae, and Pisauridae, in addition to periphyton and detrital samples, were analyzed for the stable isotopes 13C and 15N. For

Tetragnathidae and Pisauridae, tissue was combined from multiple individuals into a single composite sample in order to capture variance (Lancaster and Waldron 2001). For darters, plugs of skinless dorsal muscle tissue were removed for analysis (Pinnegar and

Polunin 1999).

All samples were oven dried at 55ºC (~ 48 h), ground using a mortar and pestle

(tissue and periphyton) or ball grinder (detritus), weighed, and packed in tin capsules before sending for analysis. Samples were analyzed by continuous-flow isotope-ratio mass spectrometry (EA-IRMS) at the Washington State University Stable Isotope Core

(Pullman, Washington) using the standards Vienna Pee Dee Belemite for C and

13 atmospheric N2 for N. Isotopic composition of samples was reported in conventional  C and 15N notation. Analytical precision was 0.2‰ and 0.08‰ for determination of 13C and 15N, respectively.

Hg analysis

Tissue samples from all consumer groups, as well as sediment and water samples, were analyzed for Hg. Preparation of tissue samples for the non-insect consumers (i.e., darters, Tetragnathidae, Pisauridae) is described under “Stable isotope analysis”. For the insect consumers (i.e., benthic and emergent insects), individuals from the two most numerically dominant benthic insect families and the two most numerically dominant emergent insect families for each reach were composited for a single tissue sample per family; tissue preparation was as described for the non-insect consumers. All samples were analyzed at the Service Testing and Research Laboratory, The Ohio State

University (Wooster, Ohio) using cold vapor atomic fluorescence to analyze for total Hg

(ng g-1, dry weight). Instrument calibration (CETAC M8000 mercury analyzer, CETAC

Technologies, Omaha, Nebraska) was performed with NIST-traceable 100 mg L-1 mercury standard (SPEX CertiPrep, Metuchen, New Jersey). Samples were digested using a 1:1 mixture of trace metal grade perchloric acid and nitric acid. Lab reagent blanks (manufactured in-house; 3% hydrochloric acid added to nano-pure water) and

NIST Standard Reference Materials were analyzed for quality assurance during

45 invertebrate sample and sediment sample sequences (1547 peach leaves and 2709a San

Joaquin soil, respectively).

Trophic position and reliance on aquatic primary productivity

Trophic positions (TP) of spiders and darters were estimated using the two-source food-web model from Post (2002): TP = λ + {δc – [δaq × αaq + δte × (1-αaq)]} / Δn, where λ is the trophic position of the basal food sources (i.e., 1 for primary producers); δc is the

15 15 δ N signature of the consumer; δaq and δte are the δ N signatures of the two aquatic and terrestrial basal food sources, i.e., periphyton and detritus, respectively; αaq is the proportional reliance of the consumer on algal primary productivity; and Δn is the enrichment in δ15N per trophic level (i.e., 3.4 ‰; Post 2002).

Consumer reliance on aquatic primary productivity (i.e., αaq) was estimated using a two end-member, two isotope Bayesian mixing model solved with the R software package SIAR (Stable Isotope Analysis in R; Parnell and Jackson 2013). SIAR is equipped to handle variability in sources, consumers and trophic fractionation factors

(Parnell et al. 2010). δ13C and δ15N data were used to estimate the contribution from each basal food source (i.e., periphyton and detritus) to the consumer. Because the trophic fractionation factor, or isotopic enrichment between the consumer and basal food sources, was not known a priori (αaq was needed to estimate TP), the model was run multiple times for each consumer using TP estimates based on αaq = 0, 0.1, …, 0.9, 1. The TP estimates were used to generate trophic fractionation factors by subtracting one and multiplying by the per trophic step fractionation (i.e., 3.4 ‰ ± 0.98 ‰ for δ15N and 0.39

13 ‰ ± 1.3 ‰ for δ C; Post 2002). The best αaq estimate was then selected by weighted minimum sum-of-squares (weighted by variance in δ15N and δ13C) between the measured consumer δ15N and δ13C signatures and consumer δ15N and δ13C signatures estimated from αaq, trophic fractionation and the basal food source signatures.

Statistical analysis

To investigate the influence of geomorphic adjustment on Hg availability and the biotic community, we first compared abiotic and biotic Hg levels and benthic and emergent insect family richness (along with tetragnathid spider abundance) between adjusting and equilibrium RPIs using the Student’s paired t-test. We then evaluated the influence of specific geomorphic variables (i.e., width-depth ratio, D16, D50, velocity, and discharge) on Hg body loads in consumers using regression analysis, based on an unpaired design in which we included all 20 reaches. For this phase of the analysis, RPIs were considered sufficiently independent from each other, given that study reaches were separated by at least one riffle-pool sequence (~ 200 – 300 m) and that fine-scale hydrogeomorphic controls on fish assemblages, for example, have been shown to be expressed at the reach level (Sullivan et al. 2006).

The impact of reach geomorphic condition (i.e., adjusting or equilibrium) on benthic community composition and emergent insect community composition was assessed qualitatively using Non-metric Multidimensional Scaling (NMS) and more formally (without distorting data into lower dimensions) with Permutational Multivariate

Analysis of Variance (PERMANOVA; Anderson 2001). NMS was performed on

Jaccardian distance matrices (scaled by variance to provide more equal weighting for less abundant families), which are generally preferred for abundance data so that double absences do not contribute toward distance determination (Legendre and Legendre 1998).

In NMS, distance matrices are rank-ordered and the solution (in this case a 2-dimensional ordination) determined iteratively by minimization of a stress criterion (Kruskal, 1964).

In PERMANOVA, between and within group variances are compared about group centroids in p-dimensional space (where p is the number of benthic or aquatic emergent families). The test statistic is a pseudo F-ratio of mean between group sum-of-squares over mean within group sum-of-squares.

Hg was evaluated as a food-web tracer using simple linear regression as a predictor for TP and αaq for all three non-insect consumer groups (darters, Pisauridae and

Tetragnathidae). Additional analyses included analysis of variance (ANOVA) to analyze differences in αaq among consumer groups, followed by Levene’s test for homogeneity of variance to test for significant differences in variances among these results. The

Student’s paired t-test was used to assess potential differences in mean, maximum, and standard deviation of TP and reliance on aquatic primary productivity between adjusting and equilibrium RPIs.

All simple linear regressions and t-tests were performed using JMP® Version

10.0 Statistical Discovery Software (SAS Institute, Cary, North Carolina). The statistical analyses ANOVA, NMDS and PERMANOVA were performed in R 2.15.1 (R Core

Team 2012); NMDS and PERMANOVA additionally required the R-package VEGAN

(Oksanen et al. 2013). Logarithmic transformations were used, where necessary, to

48 normalize data before analysis (Zar 1984). For all tests, significance was determined at α

= 0.05, with α = 0.10 considered evidence of a trend as per Bocharova et al. (2013).

Results

In this study, sediment appeared to be the primary route of Hg exposure to organisms, as mean sediment Hg levels across all reaches were elevated compared to Hg levels in water (20.1 ± 15.5 ng g-1 Hg compared to 0.0028 ± 0.0023 ng g-1 Hg, Table 2.2).

Across all reaches, benthic insects had significantly lower mean Hg levels than emergent

-1 -1 insects (113.6 ± 56.9 ng g vs. 181.3 ± 110.4 ng g , t32 = -2.83, p = 0.008, Table 2.2).

Darters, as the group with lowest mean Hg levels across all reaches (46.2 ± 14.3, Table

2.2), appeared to reflect the lower Hg levels of their benthic prey. Of the two groups of riparian spiders, Tetragnathidae showed a significantly higher Hg accumulation than

-1 -1 Pisauridae (228.7 ± 77.2 ng g vs. 175.9 ± 55.6 ng g , t19 = 2.86, p = 0.010, Table 2.2).

Across all reaches, mean TP for consumer groups was highest for darters (3.9 ± 0.4), followed by Tetragnathidae (3.0 ± 0.5), and then Pisauridae (2.9 ± 0.7, Table 2.3). Mean reliance on aquatic primary productivity was less distinct among consumer groups over all reaches, comprising 0.24-0.28 of diet (Table 2.3). There were no significant differences between adjusting and equilibrium RPIs in mean, maximum, or standard deviation of TP and reliance on aquatic primary productivity (p > 0.05, Table 2.3).

Influences of fluvial geomorphology on Hg availability and invertebrate community composition

Hg levels in sediment and water were not significantly different by geomorphic condition (i.e., between adjusting and equilibrium reaches; t9 = -1.00, p = 0.343 and t8 =

0.31, p = 0.767, respectively). However, benthic insects showed significantly higher mean Hg levels at adjusting than at equilibrium reaches (t9 = 2.94, p = 0.017), while aquatic emergent insects tended to have higher mean Hg levels at equilibrium reaches (t7

= -1.95, p = 0.092). When evaluated against specific stream geomorphic variables, consumer Hg levels exhibited significant relationships for benthic insects only. In particular, D16 (as a measure of fine sediment) was a strong negative predictor of benthic

2 2 insect Hg (R = 0.37, F = 10.75, p = 0.004; Figure 2.3a), as was D50 (R = 0.33, F = 8.85, p = 0.008; Figure 2.3b).

Benthic insect family richness was significantly higher at equilibrium reaches than adjusting reaches (t9 = -3.54, p = 0.006). However, geomorphic state did not significantly influence benthic insect community composition (PERMANOVA: F1,18 =

1.32, p = 0.117; Figure 2.4a), emergent insect community composition (PERMANOVA:

F1,18 = 0.749, p = 0.959; Figure 2.4b), or emergent insect family richness (t9 = 0.53, p =

0.607). Tetragnathid spider abundance was ~ 17% greater at equilibrium than adjusting

RPIs (t17 = -2.28, p = 0.036).

Hg as a food-web tracer

We observed an overall negative relationship between Hg levels and TP across our representative non-insect consumer groups (Pisauridae, Tetragnathidae, and darters)

(R2 = 0.24, F = 16.37, p < 0.001; Figure 2.5). Within consumer group, however, Hg

50 increased with TP for darters (R2 = 0.21, F = 3.54, p = 0.082), but exhibited no relationship for Tetragnathidae (R2 = 0.00, F = 0.14, p = 0.716) or Pisauridae (R2 = 0.11,

F = 2.21, p = 0.155; Figure 2.5). Two-source mixing models indicated that mean reliance on benthic algae (i.e., aquatic primary productivity) was similar for all groups (ANOVA:

F2,55 = 1.06, p = 0.354). Pisauridae, however, exhibited a greater range in both reliance on aquatic primary productivity and in TP than other consumer groups (Figure 2.6a,b), although variances were not significantly different among Pisauridae, Tetragnathidae, and darters (Levene’s test: F2,55 = 0.18, p = 0.669 and F2,55 = 0.49, p = 0.521, respectively). In contrast, Hg levels were most variable among Tetragnathidae, followed by Pisauridae, then darters (Levene’s test: F2,56 = 11.90, p < 0.001, Figure 2.6c). Hg was not related to aquatic primary productivity across or within the non-insect consumer groups (Figure 2.7).

Among aquatic insect consumers, benthic insects tended to have lower Hg levels than emergent insects [t17 = 1.97, p = 0.065], mainly because of significant differences in

Hg by family [F2,53 = 19.49, p < 0.001; Figure 2.8]. It should also be noted, however, that within family, benthic larval insects tended to have higher Hg levels than emergent adults [e.g., Hydropsychidae; t6 = -2.23, p = 0.067; Figure 2.8]. Also of note, of the two most dominant benthic insect families sampled, non-emergent Elmidae showed significantly lower Hg levels than emergent Hydropsychidae (t18 = 3.53, p = 0.002).

However, Elmidae exhibited significantly higher Hg levels at adjusting reaches than at equilibrium reaches (t17 = 2.45, p = 0.030), whereas Hg levels of Hydropsychidae were invariant (t17 = 0.28, p = 0.784).

Discussion

Fluvial geomorphology is a driver of sediment supply and storage (Montgomery and Buffington 1997) and may therefore affect the accessibility of sediment-bound Hg to aquatic organisms (Boudou and Ribeyre 1997). Fluvial geomorphic effects may also extend to the instream biotic community, influencing the overall food-web structure via the physical habitat template (Walters et al. 2003, Sullivan 2013) and hence trophic accumulation and transfer of Hg and other trace metals, eventually crossing the aquatic- terrestrial boundary (Sullivan and Rodewald 2012). Therefore, the potential for geomorphic condition to influence both the transport and use of Hg, recently proposed as a food-web tracer (Soto et al. 2011, 2013), is significant.

Whereas Hg has been found to increase with trophic position in aquatic and riparian systems (Cabana et al. 1994, Stewart et al. 2008, Keller et al. 2014, Lodenius et al. 2014), we observed a negative relationship between Hg and TP across all three non- insect consumer groups (i.e., Tetragnathidae, Pisauridae, and darters). Thus, our study of linked stream-riparian food webs along a protected river of central Ohio, USA suggests that as a stand-alone tracer, expected patterns of Hg bioaccumulation (increasing Hg with

TP) may be obscured by complexities in the physical environment that are associated with shifts in community composition, Hg bioaccumulation, and trophic dynamics. Our findings provide important evidence that the use of Hg as a tracer within a more refined context, for example targeting specific portions of the food web and being especially

52 cognizant of family-level differences in Hg storage, may shed light on trophic relationships and the movement of Hg between aquatic and terrestrial systems.

Fluvial geomorphology as a mechanism driving Hg dynamics in linked stream food webs

Fluvial geomorphology may influence Hg bioaccumulation through its impacts on both abiotic (sediment-Hg dynamics and storage) and biotic (community composition) characteristics and their interactions. Alterations in stream geomorphology also influence the physical aquatic-terrestrial interface [e.g., via bank failure (Neller 1988, Booth 1990), channel widening (Allmendinger et al. 2007), and stream-floodplain connectivity (Gore and Shields 1995, Richards et al. 2002)] and, therefore, aquatically-derived Hg transfer to terrestrial consumers. We studied the contribution of fluvial geomorphology as a mediator of Hg uptake by consumers from two perspectives: (1) qualitative channel adjustment and (2) the influence of specific quantitative geomorphic parameters on Hg levels of benthic and emergent insects, insectivorous stream fish (darters in this study), and riparian spiders (Pisauridae and Tetragnathidae).

The anoxic conditions that develop in aquatic sediments enable methylating bacteria to convert Hg to its more bioavailable form (Morel et al. 1998). Even when there is very little Hg detected in the water column, as was the case in our study and others (e.g., see Daniel and McCreadie 2014 for a brief overview of reasons for low Hg in stream water), Hg loads may exist in sediments and organismal tissue on the order of

10s to 100s ng g-1 (Daniel and McCreadie 2014, Table 2.2). This is likely explained by rapid attachment of source Hg from the water column to dissolved organic matter sorbed

53 to sediments (Driscoll et al. 1995, Wang et al. 2004). In our study reaches, pH was basic

(Bey and Sullivan 2014); however, acidic conditions have been linked to decreasing adsorption rate to sediments (George and Batzer 2008) and increased algal-sourced Hg

(Jardine et al. 2012). Therefore, fluvial geomorphology can exert control on Hg storage and accessibility through both sediment-Hg dynamics and sediment storage.

We anticipated that Hg availability due to sediment access and, correspondingly, bioaccumulation would be higher at adjusting vs. equilibrium RPIs. We presumed that both increased aggradation and smaller associated particle size (Table 2.1, although adjusting reaches trended towards smaller D50 only) would increase Hg accessibility to consumers, as Hg tended to be stored on sediment surfaces rather than in the water column (20.1 ± 15.5 ng g-1 Hg versus 0.0028 ± 0.0023 ng g-1 Hg, respectively; Table 2.2) and smaller particle size results in greater surface area for Hg accessibility (Inoue 1980).

Similarly, lower velocity and discharge at adjusting RPIs (Table 2.1) implies less potential to mobilize sediment relative to higher velocity and discharge measured at equilibrium reaches (Adenlof and Wohl 1994, Lisle 1995) and thus may lead to increased

Hg access. Areas of pronounced sediment storage such as wetlands (Sinclair et al. 2012) and near-shore sediment deposits (Richman and Milani 2010) with high Hg levels have been linked to high insect Hg body loads. In addition to visual signs of channel aggradation, many adjusting RPIs had significantly higher width-depth ratios and trended towards smaller particle size (p = 0.047 and 0.067, respectively, Table 2.1). As sediment

Hg concentrations did not significantly differ between adjusting and equilibrium RPIs, differences in Hg levels of biota at RPIs of differing geomorphic condition were likely

54 influenced by sediment size and distribution (and thus, accessibility) rather than Hg concentrations within Darby Creek.

Among the various food-web members considered in our study, geomorphic characteristics emerged as significant only for Hg levels of benthic insects. Percentage of fine sediment exerted the strongest influence on benthic insect Hg levels, with a decrease

2 in D16 correlating to an increase in Hg (R = 0.37, F = 10.75, p = 0.004, Figure 2.3a).

When categorized by geomorphic condition, benthic insects at adjusting RPIs had significantly higher Hg levels than at equilibrium reaches (p = 0.017). We interpret this result as evidence that benthic insects likely exhibit the closest link with RPI geomorphic parameters due to substrate contact, with higher-level consumers showing unexpected patterns due to foraging behavior (see discussion on Hg as a tracer, below) and the sampled emergent insect community evidencing variability most likely through family- level differences in shedding contaminants through metamorphosis (reviewed in Kraus et al. 2014).

We also hypothesized that fluvial geomorphology would influence insect community composition, with implications for Hg bioaccumulation and transfer from the aquatic to terrestrial zone. For aquatic insects, the link between geomorphology and biota appears strongest at the benthic larval stage, as benthic insect family richness was significantly higher at equilibrium reaches (p = 0.006), although community composition did not significantly differ by geomorphic condition. Larger particle size evidenced by equilibrium RPIs (i.e., D50, Table 2.1) likely provides better substrate habitat for a more diverse benthic insect community, although this was not reflected in the emergent insect

55 community (no significant difference by geomorphic condition for family richness or community composition).

Although in our study system adjusting reaches trended towards smaller D50 only, many studies have shown a reduction in benthic insect diversity with increasing percentages of fine sediment (Lenat 1981, Waters 1995, Descloux et al. 2013). Bjornn et al. (1977) observed a negative relationship between Ephemeroptera, Plecoptera, and

Trichoptera (EPT) and an increase in fine sediment, with the notable exception of

Baetidae (Angradi 1999). These changes in assemblage may be primarily due to a homogenization of habitat resulting from the filling in of interstitial spaces of large particles (Lemly 1982) and can influence a shift in feeding groups from filter-feeders towards collector-gatherers (Relyea et al. 2000), although we did not observe this change.

As we determined from our study (Figure 2.8) and consistent with other research

(Cid et al. 2010, Buckland-Nicks et al. 2014, Kraus et al. 2014), benthic invertebrate community composition may directly influence aquatic-to-terrestrial Hg transport due to notable differences in family-level storage. For example, increased aggradation and decreased particle size may make Hg more accessible to the benthic invertebrate community; however, the same geomorphic characteristics might influence the community towards one in which non-emergent taxa such as oligochaetes and other non- insect taxa make up a larger proportion (Angradi 1999, Zweig and Rebeni 2001), thereby reducing Hg transfer to riparian consumers. In our study, Elmidae, which spend their full life-cycle as collector-gatherers on the substrate (Elliott 2008) and are therefore closely associated with sediment, represented one of the two most dominant benthic insect taxa.

Elmidae had significantly higher Hg levels at adjusting reaches (p = 0.030), but are non- emergent. However, Hydropsychidae, which spend relatively less time in sediment contact as filter-feeding retreat builders (Wallace 1974) and emerge from the aquatic environment as adults, showed no difference between geomorphic condition of RPIs

[although it should be noted that independent of geomorphic condition, Hydropsychidae exhibited significantly higher Hg levels than Elmidae (p = 0.002, Figure 2.8)]. Increased

Hg access through sediment storage and particle size might therefore be mediated by a shift toward non-emergent taxa, especially in contexts where aggradation is more pronounced than in our protected system.

Additionally, Bey and Sullivan (2014), within the same study system, found that fish species richness (including but not limited to darters) was significantly greater at equilibrium than adjusting RPIs. Geomorphic adjustment has been shown to structure fish assemblages (Sullivan et al. 2006, D’Ambrosio et al. 2014) with a more diverse community favored by less geomorphically-altered streams. Although broad categories such as stream order exert controls on fish assemblages (Thornbrugh and Gido 2010,

Yoon et al. 2011), small-scale habitat variables such as substrate type greatly contribute to local species abundance (Smith and Kraft 2005).

Fluvial geomorphology may also affect the strength of aquatic-terrestrial linkages through direct influences on those biota constituting important trophic connections between the two systems. Riparian arthropods can be susceptible to hydroecological modifications (Ellis et al. 2001, Manderbach and Hering 2001, Sadler et al. 2004) and function as a link between the aquatic and riparian zones (Baxter et al. 2005, Paetzold et

57 al. 2005). Paetzold et al. (2008) found that richness and abundance of riparian spiders were primarily affected by availability of exposed gravel above the typical high-water mark, suffering negative consequences in cases of channelization and the resulting inundation frequency and higher-peaking hydroperiods. In a headwater system, Iwata

(2007) found that Tetragnathidae were distributed in relation to aquatic insect emergence rates, noting that geomorphology affects organic matter storage via pools around which

Tetragnathidae tended to have higher abundances. Akamatsu and Toda (2011) found that

Nephila clavata, a riparian orb-weaving spider of the family Nephilidae, did not change its reliance on aquatic subsidies during periods of high flooding (a possible consequence of hydrogeomorphic modification, Paetzold 2008), but experienced stunted growth after large floods due to depressed aquatic prey abundance.

For both Tetragnathidae and Pisauridae, stream connectivity to the riparian zone is an important consideration, with modifications that reduce physical connectivity between the stream and the riparian zone likely to have negative consequences.

Pisauridae are found to rely on a combination of upland flood refugia and exposed channel bed in order to effectively forage on aquatic prey (Greenwood and McIntosh

2008); a disconnect between the two was shown to reduce spider abundance and richness in a heavily channelized stream (Paetzold et al. 2008). Tetragnathidae, as web-weavers, are limited by the proximity of riparian vegetation to food sources. We found significantly higher numbers of Tetragnathidae at equilibrium reaches, characterized by lower width-to-depth ratios (and thus potentially higher aquatic-riparian connectivity via proximity of riparian vegetation to the wetted perimeter), than at adjusting reaches (p =

0.036). Alterations to geomorphology, in affecting the strength of the aquatic-terrestrial trophic link, may therefore affect Hg transport across the aquatic-terrestrial boundary.

Hg as a food-web tracer

The two sources of dietary input to streams – aquatic primary producers and terrestrial leaf litter – may exhibit similar Hg levels (Tsui et al. 2009, although some aquatic primary producers can have much higher levels), but are typically distinct relative to their bioavailability to organisms. Although leaf litter stores atmospherically- deposited Hg (St. Louis et al. 2001), the levels of which can increase with decomposition

(Demers et al. 2007), the proportion of total Hg as the more bioavailable form (MeHg) is typically low (<10%, Hall and St. Louis 2004; but see Skyllberg et al. 2003).

Hg contamination is of particular concern in aquatic systems because MeHg is produced primarily in aquatic, anoxic sediments by Hg-methylating bacteria (Morel et al.

1998, Benoit et al. 2003). Once produced, MeHg is acquired by higher trophic-level organisms primarily from their food sources (Morel et al. 1998, Watras et al. 1998). In this study, however, there was no relationship between Hg and reliance on benthic algae

(Figure 2.7). This lack of relationship may in part stem from the generally low reliance on aquatic primary productivity for the consumers of this study (mean reliance 0.24 to

0.28; Table 2.3), but is also consistent with other aquatic-based studies. Tsui et al.

(2009), for example, suggested that dependence on algal versus terrestrial sources cannot entirely explain Hg accumulation, as the two macroinvertebrate groups in that study with strongest affinities for algal and terrestrial sources, respectively, exhibited similar MeHg

59 levels. Similarly, a study by Jardine et al. (2012) showed no clear differences in Hg concentrations of organisms at the same trophic level despite differing food sources in neutral pH (7.0 – 8.0) waters. Note that the alkaline waters typical of our study system

(pH range of 8.3 – 9.2; Bey and Sullivan 2014) may have also influenced the lack of relationship between Hg and reliance on aquatic algal C. Mattieu et al. (2013), for example, observed a negative effect of water alkalinity on Hg bioaccumulation in bass, whereas Hg appears to bioaccumulate more readily in acidic waters (Chasar et al. 2009,

Scudder et al. 2009, Jardine et al. 2012) and may thus better distinguish between aquatic and terrestrial basal sources under those conditions. Collectively, our results suggest Hg more strongly discriminates among trophic positions than between aquatic and terrestrial food sources.

As we found similar reliance on aquatic primary productivity for all consumer groups in this study, we conclude that consumer differences in Hg accumulation were related to Hg levels in the aquatic prey base rather than reliance on aquatic versus terrestrial basal resources. Among the aquatic invertebrate prey sampled, Hg levels were significantly lower in benthic larval insects than in emergent adult insects (p = 0.008).

Within a single taxon [based on a subset of those study reaches from which both benthic and emergent members of the same family (Hydropsychidae) were obtained], Hg tended to decrease from benthic to emergent insects. Hg depuration during metamorphosis has also been documented in several previous studies (Sarica et al. 2005, Cid et al. 2010,

Buckland-Nicks et al. 2014, Kraus et al. 2014). It is important to note that for this study, benthic versus emergent insect differences in Hg loads were based on the two dominant

60 families sampled for each study reach, which is unlikely to be reflective of the entire benthic and emergent insect communities, yet represents a numerically important food- web component.

Comparing aquatic insect Hg levels to those of riparian spiders and darters suggests that Tetragnathidae, which exhibited the highest mean Hg levels (229 ± 77 ng g-

1 ), may be more reliant on energetic pathways through emergent insects, which is consistent with findings of multiple studies (Sanzone et al. 2003, Wesner 2012). The lower mean Hg levels (46 ± 14 ng g-1 ) of darters supports a reliance on benthic macroinvertebrates (French and Jude 2001, Gillette 2012). Pisauridae, with intermediate mean Hg levels (176 ± 56 ng g-1) may be reliant on both pathways (Williams 1979,

Greenwood and McIntosh 2008). Collectively, these trophic linkages could account for the decrease in Hg with increasing TP in riparian spiders and darters (e.g., darters were at a significantly higher TP position than Tetragnathidae, but had significantly lower Hg levels; Tables 2.2 and 2.3).

Benthic larval insects, which form the base of stream food webs, however, can exhibit marked differences in Hg storage by family (Cid et al. 2010). This pattern was echoed in our results, where we found that the difference in Hg levels between benthic and emergent insects was driven by differences in family-level Hg storage.

Chironomidae (one of the dominant emergent families), for example, had the highest Hg levels (245.3 ± 122.3 ng g-1), while Elmidae (one of the dominant benthic families) had the lowest Hg levels (89.5 ± 52.9 ng g-1) among the sampled insect families. Differences in Hg levels by insect taxon have been observed in other aquatic-based studies (Tsui et al.

2009, Cid et al. 2010, Buckland-Nicks et al. 2014) and may depend on feeding strategies

(Tsui et al. 2009), degree of contact with sediment (Kolaříková et al. 2012), and physiological processes such as early instar development (Cid et al. 2010). These findings, along with results from the current study, suggest that taxon-level differences in

Hg accumulation might limit its use as a food-web tracer across multiple consumer groups. Of note, and concordant with our findings of decreased Hg with increasing TP across all consumer groups, a study of wetland macroinvertebrates by George and Batzer

(2008) found much lower Hg levels in larger-bodied organisms of higher trophic position than in amphipods; the authors suggest that bioaccumulation pathways in macroinvertebrates may be more complex than previously thought.

Within specific consumer groups, however, Hg might still present a valid tracer.

For example, darter Hg and trophic position were weakly related (p = 0.080). Chasar et al. (2009) observed a narrower Hg range in fish versus invertebrates, noting that invertebrates are likely associated with local variability in substrate type, flow, and redox conditions, while more mobile and longer-living fish are less susceptible to such local variability reflected in Hg accumulation. In the case of darters in our study, bioaccumulation of Hg appears to be a comparable trophic-level tracer to stable isotope analysis.

Although Tetragnathidae appeared to reflect higher levels of Hg in concordance with higher levels of Hg in emergent insects, there was no correlation between Hg levels and TP. As web weavers, Tetragnathidae may feed on either aquatic or terrestrial insects; therefore, it is possible for them to obtain Hg from both sources. Although the focus of

62 this study was on the aquatic zone as source of Hg, the terrestrial zone may also be an important contributor to Hg accumulation in Tetragnathidae. Skyllberg et al. (2003) found that the potential for inorganic Hg to be converted to bioavailable MeHg was greatest in the riparian corridor due to peaty stream-bank soils; correspondingly, Tsui et al. (2014) noted the potential for riparian invertebrates to store terrestrially-derived Hg and return it to aquatic predators. Therefore, although terrestrial food sources could account for some variability, it is necessary to explore other potential contributors to Hg variability in tetragnathid spiders.

Variability in tetragnathid spider diet within the emergent insect compartment [as different tetragnathid species may spatially segregate construction of their webs according to prey selectivity (Tahir et al. 2012)] may be a more tenable mechanism explaining the lack of relationship between Hg and TP. Our composite Tetragnathidae samples for each study reach may have obscured variability in species-specific prey selections. Similarly, streams experiencing nutrient enrichment may cause

Tetragnathidae to be more selective on smaller-bodied individuals with aquatic-based diets (due to more manageable emergent insect body size and consumption efficiency;

Davis et al. 2011). Total phosphorus in streamwater from our study was 0.114 ± 0.009 mg L-1 (Sullivan, unpublished data) with 0.08 mg L-1 as Ohio Environmental Protection

Agency’s target for Exceptional Warmwater Habitat for wadeable streams (OEPA 2006).

Thus, Tetragnathidae could be preferentially feeding on small-bodied emergent insects such as Chironomidae, which in this study exhibited higher Hg loads (Figure 2.8; see also

Kolaříková et al. 2012) although this link has not been established empirically. Finally,

63 periodic flooding common to the system may reduce the numbers of emergent insects available as riparian prey (Greenwood and McIntosh 2008), which may account for variability in Tetragnathidae diet over the sampling period (e.g., highly variable emergent events rather than predictable pulses, preventing steady bioaccumulation through selectively feeding on higher-Hg-storing families).

Darter reliance on benthic macroinvertebrates (French and Jude 2001) as a constant source of Hg accumulation contrasts with the opportunistic prey-switching behavior of Pisauridae, which depredate emerging insects from the water surface or shoreline, small fish, and terrestrial prey (Greenwood and McIntosh 2008, Nyffeler and

Pusey 2014). Pisauridae feed opportunistically, but aquatic prey makes up a large portion of their diet, especially stoneflies and caddisflies that remain in riverbank habitat after emerging (Williams 1979). In our study, Hydropsychidae had significantly lower Hg levels than the smaller-bodied and presumably less accessible Chironomidae (that tend to emerge most often from the open channel, Malison et al. 2010), which likely accounts for lower bioaccumulation of Hg in Pisauridae relative to Tetragnathidae, though substantially higher than benthic-feeding darters (Figure 2.6c).

Relative to lower levels of darter Hg, the trend of decreasing Hg with increasing

TP (Figure 2.5) may be explained by Pisauridae (also known as “fishing spiders”) feeding on fish (in this study, darters were used as a proxy for small fish) occupying a higher TP but with lower Hg levels relative to other consumer groups considered in our study. A review of global fish predation on spiders by Nyffeler and Pusey (2014) implicated

Pisauridae in > 75% of reported incidences, noting that most fish captured were 2–6 cm-

64 long representatives of the most common taxa occurring in their respective geographic areas. The most common species across our study reaches were greenside darter, 13.6% of assemblage; sand shiner, 11.6% of assemblage; and banded darter, 11.5% of assemblage (Bey and Sullivan 2014). Higher pisaurid spider TP and lower Hg body loads may suggest that Pisauridae in our system are more dependent on small fish than emergent insects. Adding to potential variability in the relationship between Hg and TP is the exceptionally high rate of cannibalism among Pisauridae, including both pre-

(Johnson and Sih 2005) and post- (Schwartz et al. 2014) copulatory cannibalism. Larger females commonly cannibalize the smaller males, which in turn are more likely to depredate high-Hg-level emergent insects on the riverbank than low-Hg-level fish due to body size (Zimmermann and Spence 1989). These females may experience greater Hg accumulation and increased TP; therefore, although cannibalism alone does not explain the pattern in this study, the trend of decreasing Hg with increasing TP may have been more pronounced in the absence of cannibalism.

Conclusion

Geomorphic condition (i.e., adjusting vs. equilibrium) of study reaches –

with notable differences between paired reaches in particle size, width-to-depth ratio, velocity, and discharge – was strongly related to benthic insect Hg levels. Darters appear to readily bioaccumulate Hg likely due to a relatively consistent diet of benthic insects, which are closely tied to sediment Hg availability. Geomorphic condition was also associated with benthic insect community composition, which can alter trophic pathways

65 and Hg movement through aquatic food webs. Given these findings, geomorphic assessments, which have been found to be closely linked to biotic communities and thus have been proposed as a complement to habitat assessments (Sullivan et al. 2004,

Sullivan et al. 2006, Bey and Sullivan 2014) may also be a valuable tool in evaluating streams for Hg contamination. For example, geomorphic condition assessments may be used as a coarse-level management tool targeting potential Hg hotspots, followed by finer-resolution characteristics of the system at hand, such as seasonal variability in Hg dynamics (e.g., Daniel and McCreadie 2014), .

In addition, Hg might have considerable promise as a food-web tracer, although its use may be mediated by influences of fluvial geomorphology. Using contaminants to trace links between aquatic and terrestrial food webs has been shown to be very useful in cases of extreme exposure (e.g., Superfund sites; Walters et al. 2010) or accidental contamination events (Otter 2013). However, since Hg is widespread in aquatic systems via atmospheric deposition, affecting even unindustrialized regions (reviewed in

Fitzgerald et al. 1998), its potential use as a food-web tracer within remote and protected contexts is valuable, yet remains largely unexplored. Food-web modeling using specific families rather than the composite emergent or benthic insect community may be advantageous, and help further refine both the movement of Hg in stream food webs as well as its export to terrestrial systems. Future studies will be aided by efforts to distinguish those appropriate scenarios for which Hg can effectively supply information to more comprehensively understand food webs in aquatic-to-terrestrial contexts.

Acknowledgements

Research support was provided by state and federal funds appropriated to The

Ohio State University, Ohio Agricultural Research and Development Center. This research was also supported by the Ohio Department of Natural Resources, Division of

Wildlife through the USFWS State Wildlife Grant Program and the Ohio Biodiversity

Conservation Partnership. Thanks are extended to the Stream and River Ecology

Laboratory members for field assistance. Cooperation from Franklin County Metro Parks as well as private landowners enabled the successful completion of this research.

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Table 2.1. Summary of descriptive statistics for geomorphic variables for the all 20 Darby Creek study reaches of Ohio, USA as well as broken out by adjusting and equilibrium reaches.

Overall Adjusting Equilibrium

Min Med Max 푋̅ SD 푋̅ SD 푋̅ SD

Width-to- 9.1 35.4 58.3 34.5 13.5 39.8 14.9 29.2 10.1

Depth

Ratio

D16 (mm) 12.0 23.0 61.0 25.7 11.6 24.9 14.8 26.4 8.1

D50 (mm) 21.0 45.0 120.0 54.9 25.4 50.2 25.6 59.5 25.7

Velocity 1.2 2.4 6.1 2.9 1.5 2.3 1.5 3.4 1.5

(m s-1)

Discharge 18.7 84.6 626.4 125.8 133.5 80.5 61.6 171.1 171.0

(m3 s-1)

Table 2.2. Summary of descriptive statistics for Hg levels in water, sediment, invertebrates, and fish for the all 20 Darby Creek study reaches of Ohio, USA as well as broken out by adjusting and equilibrium reaches.

Overall Adjusting Equilibrium

Hg (ng g-1) Min Med Max 푋̅ SD 푋̅ SD 푋̅ SD

Biota

Tetragnathidae 123.1 207.0 415.4 228.7 77.2 230.6 88.3 226.8 69.0

Pisauridae 93.0 168.9 292.4 175.9 55.6 163.9 50.3 187.9 60.6

Darters 17.0 47.0 70.0 46.2 14.3 49.5 16.0 42.6 12.0

Benthic 39.7 104.3 247.1 113.6 56.9 127.3 58.3 98.4 52.7 insects Emergent 34.4 151.1 553.6 181.3 110.4 137.3 81.8 215.9 120.1 insects Water < 0.1 < 0.1 < 0.1 < 0.1 < 0.1 < 0.1 <0.1 < 0.1 < 0.1

Sediment 6.4 14.8 67.0 20.1 15.5 17.5 8.8 22.6 20.3

Table 2.3. Summary of descriptive statistics for trophic position and reliance on benthic algal primary productivity of representative consumer groups for the all 20 Darby Creek study reaches of Ohio, USA as well as broken out by adjusting and equilibrium reaches. Values for aquatic primary productivity are given as percentages of total diet.

Overall Adjusting Equilibrium

Min Med Max 푋̅ SD 푋̅ SD 푋̅ SD

Trophic Position (TP)

Tetragnathidae 2.1 2.9 4.0 3.0 0.5 3.1 0.5 2.8 0.4

Pisauridae 1.4 2.8 4.4 2.8 0.7 2.9 0.7 2.8 0.6

Darters 3.2 3.9 5.0 3.9 0.4 4.0 0.5 3.8 0.4

Reliance on Aquatic Primary Productivity (αaq), expressed as a proportion

Tetragnathidae 0.11 0.24 0.64 0.24 0.10 0.24 0.11 0.24 0.09

Pisauridae 0.12 0.25 0.56 0.28 0.11 0.27 0.10 0.29 0.12

Darters 0.15 0.25 0.36 0.26 0.06 0.26 0.06 0.27 0.07

a 3 2 5 . . 8 . 7 1 4 . . . 1

6 .

b 11

. Figure 2.1. Coarse-level field indicators used to categorize (a) adjusting and (b) equilibrium reaches. Geomorphic condition was evaluated by assessing four channel adjustment processes: degradation/incision, aggradation, widening, and change in planform. Vertical adjustments include degradation/incision (lowering of channel bed elevation through scour) and aggradation (elevation of channel bed via sediment accumulation), with lateral adjustment occurring through channel widening (due to erosion of banks) or change in planform (formation of a new channel direction). In (a), field indicators of vertical adjustment may include (but are not limited to, see text for more complete list), for degradation, exposed till and fresh substrate (1), recently abandoned terraces (2), fresh vertical faces along banks (3); and for aggradation, high degree of substrate embeddedness (4), high width-to-depth ratios (5), incomplete riffles (6), and presence of mid-channel bars (7). Field indicators of lateral adjustment may include, for over-widening, high width-to-depth ratios (5), deposition of mid-channel bars (7), and bank overhangs (8). See text for description of change in planform. In (b), field indicators of equilibrium reaches include, for example, low degree of embeddedness (9), riffle completeness (10), and small width-to-depth ratios (11). Dash-dot lines indicate level of sediment. Dotted lines indicate level of water. 85

Thalweg not aligned with planform

Exposed substrate Incomplete riffle

Island formation

Highly-connected floodplain Vegetated channel bars Thalweg aligned with planform

Complete riffle

Figure 2.2. Example of paired (a) adjusting and (b) equilibrium RPI reaches in Darby Creek, Ohio. Example indicators of geomorphic conditions are shown. In (a), the thalweg is misaligned with channel planform and the channel is beginning to split around a recently formed island. Exposed substrate is visible in the riffle, which is not complete from bank to bank. These visual indicators suggest the reach is aggrading and changing planform. In (b), the thalweg is well-aligned with the channel planform, the riffle is complete without visible exposed substrate, channel bars are well vegetated, and the channel has high connectivity with the floodplain during high flows. These indicators suggest the reach is in a state of geomorphic dynamic equilibrium.

) a

log, ng log,ng g

(

ct Hg Hg ct Benthic inse Benthic

D (log, mm) 16

) b

- Benthic insect Hg (log, ng insect ng (log, Hg g Benthic

D50 (log, mm)

-1 Figure 2.3. Relationships between benthic mean Hg (ng g ) and (a) D16 and (b) D50. Regression statistics are as follows: (a), y = 285.9 – 54.4Log(x) with R2 = 0.37 (p = 0.004); and (b), y = 306.6 - 49.2Log(x) with R2 = 0.33 (p = 0.008).

NMS2

NMS 1

NMS2

NMS 1

Figure 2.4. Non-metric Multidimensional Scaling (NMS) ordination plots of (a) benthic community composition data (scaled by variance) and (b) emergent community composition data (scaled by variance). Adjusting reach scores are indicated by unshaded circles and equilibrium reach scores are indicated by shaded squares. The stress level was 19% in (a) and 22% in (b). The blue, dashed lines indicate variable loadings (i.e., family level abundance; only the most significant families are indicated).

)

- Consumer Hg (ng g Hg Consumer

Mean Trophic Position

Figure 2.5. Simple linear regression of mean Hg (ng g-1) levels on mean trophic positions of non-insect consumer groups. Data points are coded by consumer group: crosses = darters; squares = Tetragnathidae; and x’s = Pisauridae. Regression lines are also indicated: solid line = all consumer groups; long dashes = darters; short dashes = Tetragnathidae; and dashes with dots = Pisauridae. Regression statistics are as follows (in order from strongest to weakest): all consumer groups, y = 351.6 - 59.37 x with R2 = 0.24 (p = 0.0002); darters, y = -1.345 + 12.01 x with R2 = 0.21 (p = 0.082); Pisauridae, y = 256.0 - 27.92 x with R2 = 0.11 (p = 0.155); and Tetragnathidae, y = 192.2 + 12.14 x with R2 = 0.00 (p = 0.716).

Figure 2.6. Means and ranges of (a) reliance on aquatic primary productivity (expressed as proportion), (b) trophic positions, and (c) Hg levels of non-insect consumer groups. Box-and-whiskers indicate the medians (central bar), 1st and 3rd quartiles (box boundaries), and 95% CI (“whiskers”) for each group; outliers are plotted as open circles. The two spider families (Tetragnathidae, Pisauridae) are shaded.

a Reliance on aquatic primary productivity primary aquatic on Reliance Tetragnathidae Pisauridae Darters

Trophic position Trophic

Tetragnathidae Pisauridae Darters

continued

Figure 2.6 continued

)

Hg levels (ng g Hg

Tetragnathidae Pisauridae Darters

)

- Consumer Hg (ng g Hg Consumer

Reliance on aquatic primary productivity (log, expressed as proportion of 1)

Figure 2.7. Simple linear regression of mean Hg (ng g-1) levels with mean reliance on aquatic primary productivity (expressed as a proportion and log-transformed) for non- insect consumer groups. Data points are coded by consumer group: crosses = darters; squares = Tetragnathidae; and pluses = Pisauridae. Regression lines are also indicated: solid line = all consumer groups; dashes = darters; short dashes = Tetragnathidae; and dashes with dots = Pisauridae. Regression statistics are as follows (in order from strongest to weakest): all consumer groups, y = 140.8 - 13.72x with R2 = 0.00 (p = 0.749); darters, y = 25.76 - 15.33x with R2 = 0.06 (p = 0.374); Pisauridae, y = 212.1 + 27.15x with R2 = 0.03 (p = 0.505); and Tetragnathidae, y = 232.1 + 2.343x with R2 = 0.00 (p = 0.968).

)

ng ng g

Hg levels (log, Hg

Benthic Insects Emergent Insects

Figure 2.8. Log mean Hg (ng g-1) levels for the two most dominant benthic insect families and the two most dominant emergent insect families (ELM = Elmidae spp., HYD = Hydropsychidae spp., and CHI = Chironomidae spp.). The box-and-whiskers indicate the medians (central bar), 1st and 3rd quartiles (box boundaries), and 95% CI (“whiskers”) for each group; outliers are plotted as open circles. The two benthic families (Elmidae, Hydropsychidae) are shaded.

Chapter 3: Influences of fluvial geomorphology on aquatic-to-terrestrial Hg export in

small urban streams

Lindsey E. Boaz1,2, S. Mažeika P. Sullivan1, and Kathleen A. Hossler1

1School of Environment & Natural Resources, The Ohio State University, 2021 Coffey

Rd., Columbus, OH 43210

2Corresponding author email: [email protected]

Abstract

Although mercury (Hg) contamination is common in stream ecosystems, mechanisms governing bioavailability and bioaccumulation in fluvial systems remain poorly resolved as compared to lentic systems. In particular, streams in urbanized catchments are subject to multiple stressors that may contribute to Hg distribution, bioaccumulation, and export across the aquatic-to-terrestrial boundary. We hypothesized that fluvial geomorphic characteristics related to channel geometry, hydraulics, and sediment size and distribution would influence (1) Hg loads in sediment and benthic larval and adult aquatic emergent insects and (2) aquatic-to-terrestrial contaminant transfer to common riparian spiders (Pisauridae and Tetragnathidae) via changes in aquatic insect Hg body loads as well as in aquatic insect community composition.

Channel geomorphic characteristics were not significantly related to Hg loads in sediment or benthic and emergent insects. However, we found significant relationships between geomorphic characteristics (e.g., stream power and width-to-depth ratio) and riparian spider Hg levels and emergent insect abundance, suggesting that via these mechanisms fluvial geomorphology may play an important role in aquatic-to-terrestrial contaminant Hg transfer.

Introduction

As cross-system energetic linkages via movement of organisms and trophic pathways are considered crucial components of functional stream–riparian ecosystems

(Nakano and Murakami 2001, reviewed in Baxter et al. 2005, Newham et al. 2011, Sabo and Hagen 2012), they might also result in adverse impacts through the movement and bioaccumulation of environmental contaminants. The aquatic-to-terrestrial linkage is of particular concern because many contaminants are known to accumulate and/or become bioavailable within aquatic sediments (Morel et al. 1998, Hudson-Edwards and Taylor

2003, Heim and Schwarzbauer 2013). Hence, the transfer of energy from aquatic to terrestrial environments may provide a vector for contaminant export (Menzie 1980,

Walters et al. 2008, Sullivan and Rodewald 2012).

Notably, aquatic insects that emerge from the stream as adults (hereafter, emergent insects) can constitute an important component of the energy budgets of riparian consumers (Henschel et al. 2001, Nakano and Murakami 2001). Observations of prey use by riparian spiders (Henschel et al. 2001) and birds (Iwata et al. 2003), for example, have reflected diets comprised of 54% and 67–82% aquatic prey, respectively.

In addition to their nutritional importance, emergent insects are of particular concern as vectors of contaminants from aquatic to terrestrial systems because of their life history, with larval stages often occurring in the benthos where exposure to bioavailable contaminants can be substantial (Soto et al. 2011, Daniel and McCreadie 2014). The proclivity of aquatic insects to bioaccumulate sediment-associated contaminants as well as their role in subsidizing riparian food webs suggests that streams and rivers may

96 function as sources of lateral contaminant export from aquatic to terrestrial food webs

(Sullivan and Rodewald 2012). This lateral transfer of contaminants may introduce decades-old pollution into terrestrial food webs (Walters et al. 2008, 2010), and result in negative ecological consequences even at trace levels (e.g., Rowse et al. 2014).

In spite of the strong associations between fluvial geomorphology and fish assemblages (Walters et al. 2003, Sullivan et al. 2006, Bey and Sullivan 2014), riparian spider density and distribution (Iwata 2007, Paetzold et al. 2008), and in particular, aquatic macroinvertebrate density and community composition (Sullivan et al. 2004,

Muehlbauer and Doyle 2012, D’Ambrosio et al. 2014), the influence of stream geomorphology on emergent insect flux, and in turn, aquatic-to-terrestrial contaminant export, remains unresolved yet quantitatively important. Benthic larval insects, for example, selectively choose microhabitats (Rae 2013) and adjust their behavior to the hydraulic environment governed by local stream geomorphic characteristics (Oldmeadow et al. 2010). Pedersen et al. (2014) found that in particular, substrate heterogeneity induced by fluvial geomorphology was a primary driver of macroinvertebrate diversity.

Further, Iwata et al. (2007) showed a direct link between riparian spiders and geomorphology wherein abundances of spiders were significantly higher near detritus- storing pools where macroinvertebrates tended to drift, feed, and emerge. Therefore, relationships linking stream geomorphic characteristics to benthic assemblages suggest that geomorphic variability will also influence emergent insects, and in turn, aquatic-to- terrestrial contaminant transport.

Fluvial geomorphology may play a particularly influential role in urban systems, as these streams experience altered hydrology and geomorphology (Paul and Meyer

2001, Wenger et al. 2009). Since channel geomorphology is controlled by interactions among discharge, sediment size and load, and stream slope (Lane and Richards 1997), changes in the catchment (e.g., removal of vegetation, construction of impervious surfaces, land grading, etc.) that alter hydrology and/or sediment dynamics can lead to quantitative hydrogeomorphic alterations (Pizzuto et al. 2000, Paul and Meyer 2001,

Meyer et al. 2005, Walsh et al. 2005, Hogan et al. 2014). Relative to undeveloped catchments, urban streams tend to be characterized by a flashier and less predictable hydrograph and increased erosion (Moglen et al. 2004, Walsh et al. 2012, Navratil et al.

2013). Both channel degradation (i.e., incision) and aggradation (Ziliani and Surian 2012) can eventually lead to changes in cross-sectional geometry (e.g., widening; Allmendinger et al. 2007). The alteration of both sediment supply and water velocity can change general reach-scale streambed gradient (Pizzuto et al. 2000) and sediment-size distribution (Finkenbine et al. 2000, Pizzuto et al. 2000). Collectively, these hydrogeomorphic adjustments can contribute to reduced benthic macroinvertebrate density and diversity (Lenat 1981, Waters 1995, Descloux et al. 2013) through homogenization of habitat (Lemly 1982, Angradi 1999). Chadwick et al. (2006) observed a decline in macroinvertebrate richness and detrital processing for streams in watersheds with over 40% total impervious area, implicating altered flow regimes (and metal content) as primary factors. In Ohio, biological criteria for fish and macroinvertebrates

98 were negatively impacted by urban peak flows when compared to reference streams

(Coleman et al. 2011).

Along with impacts to aquatic biota, urban fluvial geomorphology may also affect aquatic-to-terrestrial energy and contaminant export. For example, increased aggradation and decreased particle size associated with sedimentation may increase the proportion of non-emergent taxa such as oligochaetes within the benthic invertebrate community

(Angradi 1999, Zweig and Rebeni 2001), thereby reducing prey subsidies to riparian consumers. Additionally, peak flows that scour the streambed can negatively influence the benthic (and, correspondingly, emergent) insect communities and have been shown to decrease streamside spider richness and abundance (Paetzold et al. 2008). Riparian spiders are an important aquatic-to-terrestrial link, as they can depend on aquatic insects for a large portion of their diets (~ 61% in Collier et al. 2002; see also Henschel et al.

2001, Fausch et al. 2010) and their densities tend to track those of emergent insects

(Iwata et al. 2007).

In addition to altered geomorphology, urban streams can also experience heavy metal contamination (Chakrapani and Subramanian 1993, 1996, Murray 1996, Singh et al. 1997). Total mercury (hereafter, Hg), in particular, is a widespread contaminant found in many aquatic ecosystems, especially within urban areas subject to both point sources and atmospheric deposition (Van Metre 2012). Stream geomorphology may directly affect transport of contaminants, including Hg, as sediment-bound contaminants may be differentially distributed due to reach-scale hydrogeomorphic variability that affect sediment sorting (Rhoads and Cahill 1999). After dispersal, Hg becomes bioavailable

99 through methylation within the anoxic environment of aquatic sediments, where it can then bioaccumulate in the tissues of periphyton (Brooks et al. 2012), macrophytic vegetation (Gentès et al. 2013), aquatic insect larvae (Azevedo-Pereira et al. 2012), and fish (Wren and MacCrimmon 1986). Once incorporated into aquatic consumers, Hg can then cross the aquatic-to-terrestrial boundary, as evidenced in studies of Hg bioaccumulation from fish to ospreys (Hakkinen and Hasanen, 1980) and emergent insects to insectivorous birds (Hallinger et al. 2011, Alberts et al. 2013, Rowse et al.

2014) and spiders (Walters et al. 2008).

Within this framework, we investigated the influence of fluvial geomorphology on aquatic-to-terrestrial Hg flux for 12 stream-riparian reaches in urban Columbus, Ohio,

USA. Based on a gradient of geomorphic adjustment represented by the study reaches within our system, we hypothesized that hydrogeomorphic characteristics related to channel geometry, hydraulics, and sediment size and distribution would influence (1) Hg loads in sediment, larval, and emergent insects and in turn (2) aquatic-to-terrestrial contaminant transfer to riparian spiders of the families Pisauridae and Tetragnathidae via changes in benthic and emergent insect Hg body loads and community composition (see

Figure 3.1 for details). Using the naturally abundant stable isotopes 13C and 15N, we also explored relationships between Hg concentrations in riparian spiders and their relative trophic position to help elucidate potential effects of geomorphology on Hg accumulation in consumers that can constitute a significant aquatic-to-terrestrial linkage. We provide evidence that fluvial geomorphology plays an important role in aquatic-to-terrestrial

100 contaminant transfer via combined effects on riparian spider and emergent insect abundance and Hg levels.

Material and Methods

Study reaches

We selected 12 study reaches within the Columbus Metropolitan Area, Ohio,

USA. Reaches (defined as 20 to 30 times bankfull width) were selected to represent a gradient of channel geomorphic conditions and were 1st or 2nd order tributaries of the

Olentangy River, Scioto River, or Blacklick Creek (Figure 3.2). Reaches were considered sufficiently independent from each other, as each was located within a separate subcatchment of its respective basin. Riparian vegetation among all reaches was relatively consistent in composition and coverage, with the possible exception of one reach (Cole) which experienced a heavy loss of coverage due to its location in a residential backyard. Riparian coverage at other reaches was extensive because of the low stream order of the study reaches, although the width of the riparian corridor was variable. Dominant tree vegetation included Norway maple, American hackberry, pawpaw, box elder and American elm; dominant shrubs included honeysuckle, multiflora rose, and spice bush.

Following study-reach selection, hydrogeomorphic surveys were conducted during the summer and early autumn of 2012. Benthic and emergent insect communities as well as sediment and water were sampled in the summer and early autumn of 2013.

Surveys of riparian Tetragnathidae (long-jawed orb weavers) abundance were conducted,

101 along with collection of Tetragnathidae and Pisauridae (fishing spiders) for tissue samples and periphyton and detritus for baseline stable isotope samples in the summer and early autumn of 2012 and 2013. Dissolved oxygen (DO) and pH were recorded using a YSI 600 QS multi-parameter water-quality monitor (Yellow Springs, Ohio) from a companion study. High DO in overlying waters has been shown to reduce methylated Hg flux from sediments to the water column (Henry et al. 1995), while acidic conditions have been linked to decreasing adsorption rate to sediments (George and Batzer 2008,

Jardine et al. 2012).

Hydrogeomorphic surveys

Hydrogeomorphic surveys were conducted for each of the 12 stream reaches following procedures outlined in Cianfrani et al. (2009). Channel slope (m m-1) was determined from longitudinal surveys of each reach, using a precision laser (LL300

SPECTRA Self Leveling Laser, Trimble Construction Tools Division, Dayton, Ohio).

Bankfull width (m), cross-sectional area (m2), and mean depth (m) were measured at three lateral transects across each reach, then averaged per reach. Bed grain size was estimated for each reach following Wolman’s pebble count method (1954), with 100 randomly selected clasts measured at each of the three lateral transects using a gravelometer from which D16, D50 and D95 were generated (particle size for which 16%,

50% and 95% of particles are finer, respectively, with D16 used as a measure of fine sediment). Metrics characterizing channel morphology and hydraulic conditions including velocity (m s-1), discharge (m3 s-1), Froude number, relative roughness, shear

102 stress (kg m-2), and width-to-depth ratio (mean bankfull width ÷ mean bankfull depth) were generated using the Reference Reach Spreadsheet developed by Mecklenburg

(2006) from field measurements of channel slope, bankfull width, cross-sectional area, mean depth, and bed grain size. Formulae for metrics are given in Appendix B.

Drainage area (km2) was calculated using the StreamStats GIS database (United States

Geological Survey, Reston, Virginia).

Sediment and water samples

A composite sample of three sediment cores (~10 cm) and a composite grab water sample (~ 5 cm in depth) were taken from upstream, middle, and downstream locations within the reach. Sediment was frozen in plastic sleeves before processing. Water samples were unfiltered as sample degradation was considered less of a risk to samples than contamination by filtration, given that water samples were sent for analysis within

24 h of collection. All sediment and water samples were sent to The Ohio State

University’s Service Testing and Research (STAR) Laboratory, Wooster, Ohio.

Sediment was tested for total Hg (ng g-1). Water was tested for Hg, total dissolved solids

(TDS, mg L-1), total nitrogen (TN, mg L-1), and total phosphorus (TP, mg L-1). TDS could relate to Hg supply as dissolved organic material has been shown to effectively transport Hg (Tsui and Finlay 2011), while N and P could be limiting nutrients in less- productive urban streams.

Biotic surveys and samples

103

Pyramidal emergence traps were used to collect emergent insects following

Meyer and Sullivan (2013). Traps covered a stream area of 1 m2 and were constructed of a PVC frame with square base, 18 × 16-mm mesh side panels and a collection jar at top.

Three emergence traps were set within a 30-m section (corresponding to longitudinal spider surveys) in the middle of each reach, spanning all flow habitats present (riffle, run, and pool) with one trap placed in each flow type if present. Sample collections were made at five and ten days after deployment and combined for a composite sample.

Within the same section, larval benthic aquatic macroinvertebrates were collected using a

600-cm2 Surber sampler with a 500-µm mesh net and composited over the three flow conditions. In the laboratory, benthic and emergent insects were subsequently identified to family and assigned a functional feeding group (if possible based on family-level identification) using Merritt, Cummins, and Berg (2008) and enumerated.

Surveys of riparian tetragnathid spiders were conducted along 30-m transects

(running longitudinally along each reach at the same locations as the aquatic insect collections) for 15-minutes per transect between ~ 9:00 pm – 12:00 am following Meyer and Sullivan (2013) and focused on webs within ~ 1 m of the stream edge and up to ~ 2 m in height (Williams et al. 1995). Tetragnathidae (8–10 individuals) and Pisauridae (2–

3 individuals) were also collected for stable isotope and Hg analysis.

Periphyton (i.e., benthic algae) and in-stream detritus samples were collected to establish baseline isotopic signatures. Periphyton was collected using a nylon brush from fifteen cobbles selected at random across each reach (Finlay et al. 1999). Three detrital samples were collected by hand from areas of typical flow across the reach for a

104 composite sample and frozen in plastic sleeves until processing. Coarse particulate matter (CPOM) was removed from periphyton samples, while fine particulate matter

(FPOM; particles <1 mm) was sieved from detritus samples.

Stable isotope analysis

Tissue samples from Tetragnathidae and Pisauridae, in addition to periphyton and detrital samples, were analyzed for the stable isotopes 13C and 15N. For Tetragnathidae and Pisauridae, tissue was combined from multiple individuals into a single composite sample in order to minimize variance (Lancaster and Waldron 2001).

All samples were oven dried at 55ºC (~ 48 h), ground using a mortar and pestle

(tissue and periphtyon) or ball grinder (detritus), weighed, and packed in tin capsules before sending for analysis. Spider tissue samples were analyzed by continuous-flow isotope-ratio mass spectrometry (EA-IRMS) at the Washington State University Stable

Isotope Core (Pullman, Washington) using the standards Vienna Pee Dee Belemite for C and atmospheric N2 for N. Isotopic composition of samples was reported in conventional

13C and 15N notation. Typical analytical precision was 0.2‰ and 0.08‰ for determination of 13C and 15N, respectively. Periphyton and detritus samples were analyzed by EA-IRMS at the Ohio State University Stable Isotope Biogeochemistry

Laboratory (Columbus, Ohio) under similar standards, methods, and notation.

Hg analysis

105

Tissue samples from all consumer groups, as well as sediment and water samples, were analyzed for Hg. For benthic macroinvertebrates, the sample was a composite of all families collected. For emergent insects, individuals from the two most numerically dominant families for each reach were composited for a single tissue sample per family.

Preparation of tissue samples for spiders is described under “Stable isotope analysis”, and the same procedure was followed for benthic and emergent insects. All samples were analyzed at the STAR Laboratory using cold vapor atomic fluorescence to analyze for total Hg (ng g-1, dry weight). Instrument calibration (CETAC M8000 mercury analyzer,

CETAC Technologies, Omaha, Nebraska) was performed with NIST-traceable 100 mg L-

1 mercury standard (SPEX CertiPrep, Metuchen, New Jersey). Samples were digested using a 1:1 mixture of trace metal grade perchloric acid and nitric acid. Lab reagent blanks (manufactured in-house; 3% hydrochloric acid added to nano-pure water) and

NIST Standard Reference Materials were analyzed for quality assurance during invertebrate sample and sediment sample sequences (1547 peach leaves and 2709a San

Joaquin soil, respectively).

Trophic position

Trophic positions (TP) of Tetragnathidae and Pisauridae were estimated using the two-source food-web model from Post (2002): TP = λ + {δc – [δaq × αaq + δte × (1-αaq)]} /

Δn, where λ is the trophic position of the basal food sources (i.e., 1 for primary

15 15 producers); δc is the δ N signature of the consumer; δaq and δte are the δ N signatures of the two aquatic and terrestrial basal food sources, i.e., periphyton and detritus,

106 respectively; αaq is the proportional reliance of the consumer on algal primary

15 productivity; and Δn is the enrichment in δ N per trophic level (i.e., 3.4 ‰; Post 2002).

(Parnell et al. 2010). δ13C and δ15N data were used to estimate the contribution from each basal food source (i.e., periphyton and detritus) to the consumer. Because the trophic fractionation factor, or isotopic enrichment between the consumer and basal food sources, was not known a priori (αaq was needed to estimate TP), the model was run multiple times for each consumer using TP estimates based on αaq = 0, 0.1, …, 0.9, 1.0. The TP estimates were used to generate trophic fractionation factors by subtracting one and multiplying by the per trophic step fractionation (i.e., 3.4 ‰ ± 0.98 ‰ for δ15N and 0.39

Statistical analysis

Key physicochemical gradients were identified by Principal Component Analysis

(PCA) of fifteen hydrogeomorphic parameters chosen to represent channel geometry

(easily measured in the field) and hydraulics (derived from in-field measurements and

107 selected as important criterion for macroinvertebrate habitat, Feminella 1996, Beauger et al. 2006), and five water-chemistry parameters. The fifteen hydrogeomorphic parameters included bankfull width, D16, D50, D95, discharge, drainage area, Froude number, relative roughness, max depth, mean depth, shear stress, slope, velocity, width-to-depth ratio, and cross-sectional area; the five water-chemistry parameters included DO, pH, TDS, TN,

TP. Principal component axes (PCs) with eigenvalues > 1 were retained as predictors in linear regression for Hg levels in sediment, water, benthic and emergent insects, and pisaurid and tetragnathid spiders. Retained PC axes were also used as predictors for TP estimates of Tetragnathidae and Pisauridae, as well as for abundances of all taxonomic groups. Water quality parameters were additionally used as predictors in linear regression for sediment and water Hg levels.

The impact of geomorphic characteristics on benthic macroinvertebrate and emergent insect community compositions was assessed qualitatively using Non-metric

Multidimensional Scaling (NMS) and more formally (without distorting data into lower dimensions) with Permutational Multivariate Analysis of Variance (PERMANOVA;

Anderson 2001). NMS was performed on Jaccardian distance matrices (scaled by variance to provide more equal weighting for less abundant families), which are generally preferred for abundance data so that double absences do not contribute toward distance determination (Legendre and Legendre 1998). In NMS, distance matrices are rank- ordered and the solution (in this case a 2-dimensional ordination) determined iteratively by minimization of a stress criterion (Kruskal, 1964). In PERMANOVA, between and within group variances are compared about group centroids in p-dimensional space

108

(where p is the number of benthic macroinvertebrate or aquatic emergent families). The test statistic is a pseudo F-ratio of mean between group sum-of-squares over mean within group sum-of-squares.

All linear regressions and PCA were performed using JMP® Version 10.0

Statistical Discovery Software (SAS Institute, Cary, North Carolina). NMDS and

PERMANOVA were performed in R 2.15.1 (R Core Team 2012) using the package

VEGAN (Oksanen et al. 2013). Logarithmic transformations were used, where necessary, to normalize data before analysis (Zar 1984). Data from one study reach (Rush Run) was not included in the analyses due to difficulty in obtaining representative biotic samples across the sampling period (extreme geomorphological shifts across sampling seasons, lack of sufficient invertebrate biomass for Hg and isotope analysis, emergent trap failure).

For all tests, significance was determined at α = 0.05, with α = 0.10 considered evidence of a trend as per Bocharova et al. (2013) and Rowse et al. (2014).

Results

Hg patterns in biotic and abiotic samples

In this study, sediment appeared to be the primary route of Hg exposure, as mean

Hg levels in sediment across all reaches were 31.82 ± 14.48 ng g-1 compared to 0.002 ±

0.001 ng g-1 of Hg in the water column (Table 3.1). Among the aquatic consumers, Hg levels were significantly higher in the emergent insects (267.0 ± 110.0 ng g-1) than in the

-1 benthic macroinvertebrates (85.23 ± 30.12 ng g , t11 = -5.52, p < 0.001, Table 3.1). Hg levels of the riparian consumers were intermediate between those of the aquatic

109 consumers, at 176.97 ± 43.36 ng g-1 for pisaurid spiders and 235.39 ± 88.18 ng g-1 for tetragnathid spiders (Table 3.1). There were no significant relationships in Hg levels between consumers and abiotic samples (p > 0.05).

Benthic macroinvertebrate and emergent insect communities were highly variable across all reaches, with a mean abundance of 93.08 ± 83.73 individuals for emergent insects and 46.75 ± 30.20 individuals for benthic macroinvertebrates (Table 3.1).

Compared with benthic macroinvertebrates and emergent insects, the mean abundance of tetragnathid spiders across all reaches was least variable at 38.13 ± 16.57 individuals

(Table 3.1). Mean trophic position for Tetragnathidae and Pisauridae across all reaches was 2.65 ± 0.54 and 2.01 ± 0.51, respectively (Table 3.1). Tetragnathid and pisaurid spider Hg levels were also evaluated against their respective TPs, as well as benthic and emergent insect abundances; however, these relationships were also not significant (p >

0.05). Relationships between water-chemistry variables and sediment and water Hg were not significant (p > 0.05). Tetragnathid spider Hg levels tended to decrease with increasing proportion of benthic filter feeders (p = 0.081, Figure 3.3).

PCA of fluvial geomorphology and water-chemistry parameters

PCA of the fifteen hydrogeomorphic characteristics and five water-chemistry parameters identified six axes with eigenvalues > 1 (Table 3.2). The first PC explained ~

30% of the variance; we labeled this PC “Stream Power and Nutrients” (velocity [0.864,

+], discharge [0.864, +], shear stress [0.760, +], total phosphorus [0.702, +], total nitrogen

[0.700, +], Froude number [0.607, +]). The second PC (hereafter “Channel Geometry”)

110 was driven by cross-sectional area and mean water depth ([0.756, +] and [0.740, +], respectively) and accounted for ~ 28% of the variance. “Width-to-Depth Ratio”, the third

PC, explained ~ 13% of the variance ([0.890, +]). The remaining variance was captured by PCs 4, 5, and 6 (~ 12%, 7%, and 5% of variance, respectively). However, of these, only PC5 (“Dissolved Oxygen”) exhibited significant relationships with data in our analysis and was retained for further consideration.

Influences of fluvial geomorphology on Hg levels and aquatic insect communities

“Stream Power and Nutrients” was a strong positive predictor for emergent insect abundance (p = 0.040, Figure 3.4a) and was weakly related to tetragnathid spider Hg (p =

0.098, Figure 3.4b). We did not see evidence of increasing Hg with increasing TP of

Tetragnathidae and Pisauridae, or significant relationships between TP and the PC axes.

“Width-to-Depth Ratio” tended to influence Hg body loads in Pisauridae (p = 0.068,

Figure 3.5). “Dissolved Oxygen” was a strong positive predictor of tetragnathid spider

TP (p = 0.032, Figure 3.6). There were no significant relationships between “Channel

Geometry” and Hg body loads or abundances.

Of note, although abundance of Tetragnathidae was not correlated to emergent insect abundance (p = 0.226), tetragnathid abundance tended to increase with increasing abundance of those benthic macroinvertebrate families that emerge as adults (as opposed to the entire benthic sample, p = 0.093, Figure 3.7). Additionally, the abundance of emergent insects and Hg body loads of tetragnathid spiders displayed parallel responses

111 to shifts in “Stream Power and Nutrients”, suggesting a potential trophic linkage (Figure

3.4a,b).

The influence of geomorphic adjustment and water chemistry (as expressed by the retained PCs) was additionally evaluated for both benthic macroinvertebrate community composition and emergent insect community composition. Of these relationships, there was evidence of a trend between PC5 “Dissolved Oxygen” and emergent community composition (PERMANOVA, F1,9 = 1.421, p = 0.083); specifically, this trend appeared to be one of increasing abundance of Chironomidae with increasing DO (Figure 3.8b). All other relationships were nonsignificant. However, there was evidence for a fluvial geomorphic influence on benthic functional feeding guilds, as increasing D95 was a strong positive predictor for proportion of the benthic macroinvertebrate community categorized as filter-feeding trichopterans (the families Hydropsychidae and

Philopotamidae, p = 0.040, Figure 3.9).

Discussion

Fluvial geomorphology in urban systems often differs from that of less disturbed stream systems via adjustments in channel geometry, hydraulics, and sediment size and distribution (Paul and Meyer 2001) and can influence contaminant loadings via sediment transport and storage (Koiter et al. 2013, Lecce and Pavlowsky 2014), resulting in implications for aquatic-to-terrestrial contaminant export. Although we did not find that sediment and larval insect Hg were influenced by hydrogeomorphic parameters, hydrogeomorphology was related to both spider Hg levels and emergent insect

112 abundance. Notably, both Tetragnathidae Hg and emergent insect abundance increased with stream power and nutrient loads. Pisauridae Hg increased with greater width-to- depth ratio. Our findings provide important evidence that fluvial geomorphology in urban systems can influence Hg transfer to the riparian zone, primarily via effects on the abundance of emergent insects and Hg body loads in riparian spiders.

Fluvial geomorphology and in-stream Hg

Contrary to our expectations, we observed no significant relationships between hydrogeomorphic factors and either sediment Hg levels or benthic macroinvertebrate Hg body loads. Furthermore, benthic insect Hg was unrelated to sediment Hg. The lack of relationship between fluvial geomorphic characteristics or sediment Hg and benthic insect Hg may be in part an artifact of our composite benthic insect samples (used in order to obtain sufficient biomass for Hg testing). We observed significant family-level differences in Hg storage in Big Darby Creek (Chapter 2, see also Tsui et al. 2009, Cid et al. 2010, and Kolaříková et al. 2012). Therefore, composite Hg testing may introduce variability in Hg levels due to the shifts in insect taxonomic composition across study reaches. Future studies may benefit from measuring benthic insect Hg within one or two families, although this could present some difficulty in small heavily-urbanized streams where benthic insect abundance may be very low.

Hg bioaccumulation can vary spatially in stream networks (Tsui et al. 2009), potentially due to variation in in-stream processes that regulate the production of highly bioavailable methylmercury (MeHg). Although we only measured total Hg (i.e., not

113

MeHg), we observed considerable variability in sediment Hg among study reaches (12.3

– 56.0 ng g-1, Table 3.1), although not as variable as Hg sediment contamination considered across a larger catchment of the region [e.g., 31 – 137 ng g-1 (Alberts et al.

2013) and 7 – 99 ng g-1 (Rowse et al. 2014), both in the Scioto and Olentangy Rivers].

Multiple studies suggest that urban streams may experience high Hg levels, but low methylation rates (Hurley et al. 1995, Rice 1999, Chalmers et al. 2014). Dissolved organic carbon (DOC) and total suspended solids (TSS) have both been shown to be positively related to Hg and MeHg in both dissolved and particulate phases in streamwater (Balogh and Meyer 1997, Balogh et al. 2004). Although we did not measure

DOC or TSS, for TDS, we observed no relationship with sediment or water Hg, but a positive relationship with Hg levels in Pisauridae (p = 0.035, Figure 3.5, Table 3.2). Low amounts of natural organic matter and depressed biogeochemical cycling, often typical of urban streams (Miller 2013, Chalmers et al. 2014, Silva-Junior et al. 2014), may have led to low Hg methylation (Wiener et al. 2003) in spite of relatively high Hg levels in sediment in some reaches (although note that water Hg levels were relatively low across our reaches, comparable to more undisturbed or higher-order urban systems).

Methylation also depends upon the type and quality of organic matter input. For example, Lambertsson and Nilsson (2006) found that areas relatively unaffected by Hg contamination had a net production of methylated Hg as high as that in Hg-polluted areas, due to high inputs of fresh organic matter from a well-established riparian zone

(see also Lu et al. 2013). Imberger et al. (2011) found that catchment-scale hydrology in urban systems can obscure reach-level organic inputs, resulting in organic matter export

114 rather than retention. Similar hydrologic and chemical differences across the reaches of this study, in addition to taxon-related variability in invertebrate Hg levels, may have resulted in idiosyncratic biogeochemical cycling of Hg, driving the apparent lack of relationship between sediment and water Hg levels and consumer bioaccumulation.

We expected that hydrogeomorphic effects on Hg loadings and subsequent export might also be mediated through impacts on benthic macroinvertebrate community composition. Although we observed no significant hydrogeomorphic or water-chemistry influence on benthic macroinvertebrate community composition (Figure 3.8a), there was some evidence of functional shifts. For example, we observed a decrease in proportion of the benthic macroinvertebrate community categorized as filter-feeding trichopterans (the families Hydropsychidae and Philopotamidae) with decreasing D95 (p = 0.040, Figure

3.9). This result is concordant with Gerth et al. (2013), who observed an increase in trichopteran abundance with increasing substrate size in their large-scale study of macroinvertebrate assemblage patterns across wadeable streams of the contiguous 48

United States. In a related study of the protected Darby Creek watershed of central Ohio, filter feeder Hydropsychidae exhibited significantly lower Hg levels than substrate- associated Chironomidae (Chapter 2, see also Kolaříková et al. 2012); concordantly, other studies have found marked family-level differences in Hg storage (Tsui et al. 2009,

Cid et al. 2010, Buckland-Nicks et al. 2014). Fluvial geomorphology may therefore prompt shifts in feeding guilds that may result in differential uptake of Hg among benthic macroinvertebrates. In this study, however, there was no significant relationship between proportion of benthic filter feeders and benthic Hg levels (p = 0.117).

115

Fluvial geomorphology and aquatic-to-terrestrial Hg transport

Emergent insects constitute an important link to terrestrial food webs, as they are depredated by a suite of riparian consumers including both Tetragnathidae (Sanzone et al.

2003, Wesner 2012) and Pisauridae (Williams 1979, Greenwood and McIntosh 2008).

We hypothesized that fluvial geomorphic characteristics would influence the composition of the emergent insect community. Among the PCs, “Dissolved Oxygen” influenced a trend in the emergent community towards increasing Chironomidae with increasing DO

(Figure 3.8b). We also observed a decrease in proportion of Hydropsychidae and

Philopotamidae (in the benthic life-stage) with decreasing D95 (p = 0.040, Figure 3.9).

These shifts in macroinvertebrate composition are expected to have implications for aquatic-to-terrestrial contaminant transfer, due to family-level storage differences (see discussion of benthic community, “Fluvial geomorphology and in-stream Hg”, above). In a related study on Big Darby Creek, emergent Chironomidae were found to have significantly higher Hg levels than emergent Hydropsychidae (e.g., 214.5 ng g-1 versus

107.1 ng g-1, respectively, Chapter 2); hence, an emergent community dominated by

Chironomidae rather than Hydropsychidae, etc. could result in overall greater Hg export to the riparian zone. Of note, there was a trend of decreasing Tetragnathidae Hg levels with an increase in the proportion of benthic filter feeders (p = 0.081, Figure 3.3), although this trend was not reflected in Pisauridae Hg levels.

In this study, Tetragnathidae Hg levels were unrelated to emergent insect Hg levels. However, both tetragnathid spider Hg and emergent insect abundance increased

116 with stream power and nutrient availability (p = 0.098, Figure 3.4b, and p = 0.040, Figure

3.4a, respectively; Table 3.2), suggesting that aquatic prey availability, rather than prey contaminant burdens, was more important to Hg export in our urban systems. Urban streams often experience reduced baseflows when compared to streams in more natural settings, due to reduced groundwater recharge (Shaw 1998, Ferguson and Suckling 1990,

Kennen et al. 2008). Consequently, streams of greater discharge (highly loaded onto PC1

“Stream Power”, Table 3.2) may have been better insulated from reduced baseflows than those of lesser discharge, particularly during dry periods, resulting in greater potential to support higher emergent insect abundances. Thus, an increase in Tetragnathidae Hg with increasing stream power could be explained by more favorable fluvial habitat conditions

(i.e., constant water flow and delivery of food to aquatic organisms rather than stress- inducing dry periods) to support higher abundance of emergent insect prey (thus increasing opportunity for Hg accumulation).

There is also evidence to suggest an increase in Hg bioavailability with increasing productivity (Jackson 1986, Tsui et al. 2010), although we did not observe significant relationships between nutrients and water or sediment Hg levels. Additionally, there were no significant relationships between nutrient availability (“Stream Power and

Nutrients”) and either Tetragnathidae or Pisauridae Hg levels (p = 0.160 and 0.161, respectively.) However, Tetragnathidae TP did significantly increase with DO (p =

0.032, Figure 3.6, Table 3.2). Well-aerated stream sections can provide better habitat for the macroinvertebrate community (Halwas et al. 2005, Beauger et al. 2006), potentially contributing to higher TP of Tetragnathidae feeding on emergent insects as a food

117 subsidy. For example, greater habitat quality and heterogeneity could support higher trophic levels within the insect community itself via increased diversity (Pedersen et al.

2014, St. Pierre and Kovalenko 2014), thus elevating tetragnathid spider Hg levels through feeding on emergent insects (although we observed no significant differences in community composition among reaches).

Hg levels in Pisauridae tended to increase with width-to-depth ratio (p = 0.068,

Figure 3.5, Table 3.2). As mobile predators, Pisauridae forage in the near-shore zone

(Williams 1979) and depend on large cobbles and gravel bars as shown in Greenwood and McIntosh (2008), where exposed substrate was found to be the most limiting factor to Pisauridae even in areas of high disturbance. In this respect, widening streams – which in our study were characterized by considerable exposed bankfull channel bed – could allow increased access to aquatic prey for Pisauridae, thus contributing to greater Hg accumulation. Further, Pisauridae Hg levels may have also been responding to the increase in TDS, which increased with width-to-depth ratio and can function as loci for

Hg transport within streams, especially the DOM fraction (Driscoll et al. 1995, Wang et al. 2004). Consequently, streams experiencing higher TDS loads may make Hg more available to riparian consumers through diet (although Hg levels in emergent insects were not significant along this axis in our study) and contact with medium [note: Pisauridae forage from the water itself (Schmidt 1957), while Tetragnathidae do not].

As adjusting streams become widened and aggraded (i.e., higher width-to-depth ratio), they can also become increasingly disconnected from the surrounding floodplain

(Gore and Shields 1995, Richards et al. 2002). Hence, while increasing width-to-depth

118 ratio may positively impact Pisauridae foraging within the bankfull channel, the potential subsequent stream–riparian disassociation is expected to adversely affect Tetragnathidae, as the typical wetted width of widened channels was often meters away from the streambank, and thus web-building spiders would be limited by the lack of overhanging

(the water) vegetation. Although there was no significant relationship between width-to- depth ratio and tetragnathid spider abundance in our study (p = 0.278, Table 3.2), other studies have evidenced negative effects on riparian spiders; for example, Paetzold et al.

(2008) demonstrated a reduction in spider abundance and richness with greater disconnect between the stream and riparian zone. Similarly, Burdon and Harding (2008) observed a strong decrease in riparian spider webs with increasing distance from the stream. Our findings relative to the effects of width-to-depth ratio and TDS on Hg levels in Pisauridae, combined with other work on riparian spider abundance, suggest that fluvial geomorphology may increase Hg export to the riparian zone through effects on

Pisauridae, but attenuate Hg export by a possible reduction in Tetragnathidae abundance.

Taken in concert, these findings suggest that Hg export from the aquatic-to-terrestrial zone via trophic pathways may be taxon specific, with groups such as Pisauridae, functioning as vectors of Hg storage and export in systems with an appreciable stream- riparian disconnect.

Although Hg was expected to increase with TP for Tetragnathidae and Pisauridae, we did not observe these patterns in our study. Both Tetragnathidae (Henschel et al.

2001) and Pisauridae (Williams 1979, Nyffeler and Pusey 2014) are known to feed opportunistically on both aquatic and terrestrial prey, the latter of which may also

119 contribute to Hg uptake (Tsui et al. 2014; see also Skyllberg et al. 2003). Both families prey heavily on emergent insects (Sanzone et al. 2003, Greenwood and McIntosh 2008), with Pisauridae mainly restricted to prey in the immediate shore zone such as trichopterans and plecopterans (Williams 1979), while Tetragnathidae construct webs in the riparian zone to capture more dispersed prey such as Chironomidae (often emerging from the open channel, Malison et al. 2010). However, differences in Hg levels of emergent insects are heavily expressed at the family level (Cid et al. 2010, Buckland-

Nicks et al. 2014, Kraus et al. 2014), which likely accounts for variability in Hg levels within Tetragnathidae and Pisauridae. In addition, Pisauridae may feed opportunistically on small fish (4 – 6 cm, Nyffeler and Pusey 2014), further contributing to variability in

Hg levels, depending on the propensity of Pisauridae to forage on fish among reaches.

However, small fish are not expected to be a major food source for Pisauridae in our study system based on TP (range 1.0 – 2.8, Table 3.1), compared with a TP range of 1.4 –

4.4 in the larger Big Darby Creek (Chapter 2). We interpret the lack of relationship between TP and Hg in riparian spiders to suggest that these spiders are likely also feeding on terrestrial prey sources with little to no Hg body loads.

Despite our expectations that geomorphic characteristics related to substrate particle size and distribution, channel geometry, and stream power would be associated with Hg distribution and bioavailability, our results suggest that in our study system, aquatic-to-terrestrial contaminant transfer was influenced more strongly by the combined effects on aquatic emergent insect communities and riparian consumer Hg levels.

Additional research investigating the potential role of stream geomorphic change on in-

120 stream Hg dynamics [including, for example, biogeochemical cycling, Hg transport and retention at the sub-reach scale (10-1 m, as the current research was conducted at 101-102 m), and quantifying residence time of Hg-bound sediments] will be an important step in further describing hydrogeomorphic impacts on Hg flux. Nonetheless, we have provided initial evidence that the consequences of in-stream channel change resulting from urbanization can extend beyond the stream channel boundary and influence the distribution and magnitude of Hg contamination in terrestrial consumers. This finding not only has direct relevance for linked aquatic-terrestrial food-webs, but also has broader implications for human health as aquatic organisms can be vectors of contaminant transfer to human food sources (Sullivan and Rodewald 2012). As such, studies of contaminants in stream systems will benefit from consideration of fluvial geomorphic influences within a linked aquatic-to-terrestrial context across a range of spatial scales and ecosystem contexts.

Acknowledgements

Research support was provided by state and federal funds appropriated to The

Ohio State University, Ohio Agricultural Research and Development Center. This research was also supported by the Ohio Department of Natural Resources, Division of

Wildlife through the USFWS State Wildlife Grant Program and the Ohio Biodiversity

Conservation Partnership. Thanks are extended to the Stream and River Ecology

Laboratory members for field assistance. We would also like to acknowledge Dr. Andrea

Grottoli and Yohei Matsui at the Stable Isotope Biogeochemistry Laboratory, The Ohio

121

State University, for analysis of basal isotope signatures. Cooperation from Franklin

County Metro Parks as well as private landowners enabled the successful completion of this research.

122

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Table 3.1. Summary of descriptive statistics for mercury (Hg) levels in water, sediment, aquatic benthic macroinvertebrates, emergent insects, spider, emergent insect, and benthic macroinvertebrate abundances, and trophic position (TP) of riparian tetragnathid and pisaurid spiders across 12 urban study reaches of Columbus, Ohio, USA.

Min Med Max 푋̅ SD

Biotic Hg (ng g-1)

Tetragnathidae 130.0 213.0 454.0 235.0 88.2

Pisauridae 128.0 160.0 243.0 177.0 43.4

Benthic insects 41.5 88.0 130.0 85.2 30.1

Emergent insects 126.0 246.0 475.0 267.0 110.0

Water Hg (ng kg-1) 0.2 1.7 4.3 1.9 1.0

Sediment (ng g-1) 12.3 32.6 56.0 31.8 14.5

Abundance (# individuals)

Benthic insects 13 36 112 46.8 30.2

Emergent insects 11 73 292 93.1 83.7

Tetragnathidae 16 38 71.5 38.1 16.6

Trophic Position (TP)

Tetragnathidae 1.6 2.6 3.6 2.7 0.5

Pisauridae 1.0 2.1 2.8 2.0 0.5

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Table 3.2. Principal components analysis (PCA) of hydrogeomorphic and water- chemistry variables including principal components (PC) with eigenvalues >1, % variance captured, and each PC’s loadings for each variable shared with the PCA axis. Bold font represents the loadings driving each axis. Geomorphic PC 1 PC 2 PC 3 PC 4 PC 5 PC 6 and water “Stream “Channel “Width-to- “Dissolved quality Power and Geometry” Depth Oxygen” parameters Nutrients” Ratio” Loading Loading Loading Loading Loading Loading Bankfull 0.47 0.67 0.56 0.09 0.02 -0.03 width (m) D16 (mm) -0.61 -0.13 0.25 0.29 -0.55 0.16 D50 (mm) -0.59 -0.03 0.44 0.32 -0.43 0.28 D95 (mm) 0.35 0.41 -0.19 0.63 0.02 -0.52 Discharge 0.86 0.40 0.13 0.23 -0.01 0.09 (m3 s-1) Dissolved -0.24 -0.16 0.33 0.27 0.75 0.40 oxygen (mg L-1) Drainage area -0.24 0.73 0.46 -0.12 0.12 0.00 (km2) Froude 0.61 -0.72 0.10 0.19 0.11 0.16 number Relative 0.49 0.22 -0.23 -0.61 -0.03 0.52 roughness Max depth 0.54 0.71 -0.35 -0.08 -0.13 0.10 (m) Mean depth 0.60 0.74 -0.25 -0.14 -0.01 0.06 (m) pH 0.14 0.35 -0.26 0.69 -0.34 0.34 Shear stress 0.76 -0.51 -0.13 0.35 -0.03 0.52 (kg m-2) Slope (%) 0.34 -0.90 0.01 0.26 0.03 -0.04 Total 0.70 -0.49 0.36 -0.20 -0.22 0.02 nitrogen (mg L-1) Total 0.70 -0.49 0.36 -0.20 -0.22 0.02 phosphorus (mg L-1) Total 0.18 -0.24 0.44 -0.65 -0.20 -0.29 dissolved solids (mg L- 1) Velocity 0.86 -0.40 -0.00 0.19 0.07 0.13 (m s-1) Width-to- 0.08 0.27 0.89 0.25 0.15 -0.05 depth ratio X-sec area 0.59 0.76 0.27 0.02 -0.05 -0.01 (m2) Eigenvalue 6.02 5.50 2.60 2.44 1.38 1.08 Variance 30.1 27.5 13.0 12.18 6.88 5.42

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Figure 3.1. Conceptual model of expected fluvial geomorphic influences on in-stream Hg storage and aquatic-to-terrestrial export (via effects of physical habitat on aquatic emergent insects). Direction of wedges indicate change (wide to narrow = greater to less, and vice versa; rectangular shape indicates no change) in (a) Hg storage and (b) emergent insect export with increasing (I) particle size, (II) cross sectional area and width-to-depth ratio, and (III) stream power. In (I), as substrate particle size becomes smaller, in-stream Hg is expected to increase, due to binding of Hg on sediment (Inoue 1980), ultimately making Hg available to in-stream biota and subject to increased export across the aquatic- terrestrial boundary. However, more fine sediments reduce benthic insect assemblages through homogenization of habitat (Lemly 1982), which ultimately impacts aquatic insect emergence. In (II), as channel geometry shifts from narrow, deeper streams to wider, shallow streams, Hg might be expected to increase due to increased sediment deposition via aggradation following widening (Schumm et al. 1984, Simon 1995). Aquatic insect communities shift from larger-bodied but sensitive taxa (e.g., Trichoptera, Plecoptera) to highly abundant, smaller-bodied, intolerant taxa (e.g., Chironomidae) (Freeman and Schorr 2004, Walters et al. 2009), leading to no net change in the magnitude of emergent insect Hg export and an overall increase in Hg transport. In (III), increasing stream power (velocity, discharge, Froude number, etc.) may decrease in-stream Hg availability by decreasing retention time of sediment and water, but may increase potential for Hg export due to more heterogeneous fluvial conditions for in-stream biota [for example, more consistent flow during drought conditions, relative to reduced baseflows typical of modified urban streams (Burkhead et al. 1997)]. In spite of these broad conceptual relationships, we expect that reach-level variability in the interactions between Hg storage and insect abundance and community composition will determine the magnitude of aquatic-to-terrestrial Hg transport at the site level.

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Figure 3.1 Hg Transport Implications

Net effect: no change Substrate Size a. Hg Storage Increasing magnitude of emergent (I) insect export due to an increase in favorable habitat for the benthic life stage with larger and less embedded substrate particles is offset by less Emergent Insect b. Hg bound to larger substrate Export particles.

Net effect: increase Channel Geometry a. Hg Storage Shifts from larger-bodied but (II) sensitive taxa (e.g., Trichoptera, Plecoptera) to highly abundant, smaller-bodied, intolerant taxa (e.g., Chironomidae) leads to no net b. Emergent Insect change in the magnitude of emergent Export insect Hg export with channel widening. However, greater potential to store Hg in deposited sediment and a longer residence time of water and sediment leads to increased Hg storage.

Net effect: no change Stream Power a. Hg Storage Increasing magnitude of emergent (III) insect Hg export due to more favorable fluvial conditions (i.e., diversity of flood and depth regimes) b. Emergent Insect for the benthic life stage is Export attenuated by reduced retention time of Hg in water and sediment.

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Figure 3.2. Map of the 12 study reaches distributed in 1st- and 2nd-order streams of Columbus, Ohio, USA.

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)

ng ng g Tetragnathidae Hg ( Hg Tetragnathidae

Proportion of filter-feeding trichopterans

Figure 3.3. Relationship between tetragnathid spider Hg levels (ng g-1) and proportion of filter feeding trichopterans (y = 314.1 -151.5 x, R2 = 0.301, p = 0.081). Dashed lines represent confidence curves at α = 0.05.

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(log, # individuals) # (log,

Emergent insect abundance insect Emergent

PC1 “Stream Power and Nutrients”

)

ng ng g Tetragnathidae Hg ( Hg Tetragnathidae

PC1 “Stream Power and Nutrients”

Figure 3.4. Relationships between PC1 “Stream Power and Nutrients” and (a) emergent insect abundance (y = 4.20 + 0.27x, R2 = 0.389, p = 0.040) and (b) tetragnathid spider Hg (y = 5.45 + 0.074x, R2 = 0.328, p = 0.066). Dashed lines represent confidence curves at α = 0.05.

141

)

ng ng g Pisauridae Hg (log, (log, Hg Pisauridae

PC3 “Width-to-Depth Ratio”

Figure 3.5. Relationship between PC3 “Width-to-Depth Ratio” and pisaurid spider Hg levels (y = 5.17 + 0.084 x, R2 = 0.322, p = 0.068). Dashed lines represent confidence curves at α = 0.05.

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Tetragnathidae trophic position trophic Tetragnathidae

PC5 “Dissolved Oxygen”

Figure 3.6. Relationship between PC5 “Dissolved Oxygen” and tetragnathid TP (y = 2.70 + 0.30 x, R2 = 0.417, p = 0.032). Dashed lines represent confidence curves at α = 0.05.

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(# individuals) (#

Tetragnathidae abundance Tetragnathidae

Abundance of benthic insects (log, # individuals)

Figure 3.7. Relationship between tetragnathid spider abundance and benthic insect abundance (y = 10.73 + 8.79 x, R2 = 0.282, p = 0.093). Dashed lines represent confidence curves at α = 0.05.

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Figure 3.8. Non-metric Multidimensional Scaling (NMS) ordination plots (a) benthic macroinvertebrate community composition data (scaled by variance) and (b) emergent insect community composition data (scaled by variance). Reach scores are indicated by shaded circles, with larger marker size indicating a higher score along the PC for which the community exhibited the closest relationship [(a) for the benthic community, “Channel Geometry”, PERMANOVA: F1,9 = 1.204, p = 0.183; and (b) for the emergent community, “Dissolved Oxygen”, PERMANOVA: F1,9 = 1.421, p = 0.083]. The stress level was 10% in (a) and 17% in (b). The blue, dashed lines indicate variable loadings (i.e., family level abundance; only the most significant families are shown).

NMS2

NMS 1

continued

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Figure 3.8 continued

NMS2

NMS 1

146

feeding feeding

filter

trichopterans trichopterans

Proportion of Proportion

D95 (log, mm)

Figure 3.9. Relationship between proportion of filter-feeding trichopterans and D95 (y = - 2.44 + 0.63 x, R2 = 0.390, p = 0.040). Dashed lines represent confidence curves at α = 0.05.

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