1 Aerial Insectivorous Birds Linked to Water Quality and Climate In

Aerial Insectivorous Birds Linked to Water Quality and Climate in Urbanizing Landscapes

Thesis

Presented in Partial Fulfillment of the Requirements for the Degree Master of Science in the

Graduate School of The Ohio State University

Joseph William Corra

Graduate Program in Environment and Natural Resources

The Ohio State University

2019

Thesis Committee

Dr. S. Mažeika P. Sullivan, Advisor

Dr. Stephen N. Matthews

Dr. Rachel S. Gabor

Copyrighted by

Joseph William Corra

2019

Abstract

Aerial insectivorous birds – a guild comprising swallows, nightjars, swifts, and flycatchers – have experienced alarming population declines in eastern North America in recent decades.

Although declines in individual bird species may be linked to other causes, including loss or fragmentation of suitable breeding habitat and habitat degradation in tropical wintering grounds, similar declines across multiple, taxonomically diverse species in the guild indicate that changes in flying insect prey is likely a common factor. Aerial insectivores breeding in urban areas – comprising 69.4 million acres (3.6% of total) in the contiguous United States and continuing to expand – are affected by multiple environmental changes, including alterations to local climate, habitat structure, and water quality, as well as potential shifts in both terrestrial and emergent aquatic insect prey. Emergent aquatic insects have recently been shown to provide energetic advantages to aerial insectivorous birds relative to terrestrial insects, yet they are highly sensitive to changes in water quality.

Here, I used the Tree Swallow (Tachycineta bicolor) to investigate potential associations between aerial insectivorous birds and urbanization, local climate, and water quality.

Specifically, I evaluated Tree Swallow reproductive success, body condition, and trophic dynamics at seven river-riparian sites representing urban and natural/protected land use in greater

Columbus, Ohio over four consecutive breeding seasons (2014-2017). Study sites with impervious surface in the watershed >25% were classified as urban. Urban nests were associated with higher fledging success (linear mixed-effects model [LMM]: p = 0.009) and earlier clutch initiation (LMM: p = 0.060). Nestling mass was not related to land use (LMM: p = 0.930) but

ii exhibited high interannual variability (LMM: p = 0.006), as did body condition in adult males

(LMM: p = 0.010), and mercury (Hg) in both adults (LMM: p = 0.080), and nestlings (LMM: p <

0.001). The interaction of year × land use also had a significant influence on nestling Hg (LMM: p < 0.001). I also used an Urban Stream Index (USI) to explore potential relationships between continuous measures of stream urbanization (e.g., nutrient concentrations, riparian canopy cover, water temperature) and swallow reproductive success and body condition and found that the USI was related to greater fledging success (R2 = 0.16, p < 0.001). Multiple characteristics of urban sites appeared to drive patterns between swallow responses and urbanization, including differences in mean and extreme air temperatures and measures of water quality (e.g., water temperature, nutrient concentrations, turbidity). For example, higher mean air temperature was associated with earlier clutch initiation (R2 = 0.06, p = 0.039), while the frequency of extremely cold days was related to diminished fledging success (R2 = 0.14, p = 0.003).

Relative to trophic position, I investigated the relationships between urbanization and Tree

Swallow (Tachycinenta bicolor) reliance on aquatically derived energy (i.e., originating from aquatic primary production) and trophic position. Bayesian mixing models using 13C and 15N isotopes showed that nutritional reliance on both aquatic primary production and aquatic insects had significant interannual variability. Reliance on aquatic insects by nestling swallows exhibited a significant interaction of year × land use (LMM: p < 0.001), suggesting a possible relationship between elevated total N in the water column and aquatic insect consumption. Trophic position of adult swallows was 8.3% higher at urban than at natural/protected sites (LMM: p = 0.020), whereas nestlings exhibited high interannual variability in trophic position (LMM: p < 0.001), but were not related to land use (LMM: p = 0.720). My results for reliance on aquatic insects,

iii aquatically derived energy, and trophic position all revealed a strong random effect of site, suggesting that local-scale water chemistry and land-use/land-cover characteristics may play a prominent role in shaping flying insect assemblages and driving aquatic-terrestrial energetic pathways. Supporting this finding, I also observed strong relationships between the USI and swallow trophic dynamics: USI was related to greater reliance on aquatic insects among both adult and nestling swallows (adults: R2 = 0.10, p = 0.036; nestlings: R2 = 0.23, p < 0.001), greater reliance on aquatically derived energy (adults: R2 = 0.14, p = 0.019; nestlings: R2 = 0.20, p <

0.001), and higher trophic position (adults: R2 = 0.23, p = 0.001; nestlings: R2 = 0.22, p < 0.001).

Overall, despite the loss of environmental quality generally attributed to cities, Tree

Swallows exhibited greater reproductive success in urban settings where aquatic insects were larger, local climate conditions favored egg and nestling survival, and the breeding season was longer. However, the chronic effect of elevated Hg body burdens in urban areas represents a potentially adverse impact for urban-breeding aerial insectivores. For trophic measures, gradients of urbanization appeared to mediate nutritional reliance on aquatic resources and trophic position more strongly than did a coarser, categorical classification of urbanization. Aerial insectivore energetics may have important implications for long-term population trends as related to fitness, migration survival, exposure to contaminants, and reproductive success. However, owing to differences in foraging strategy, nesting strategy, and dietary preferences, the responses of other aerial insectivorous species to urbanization may vary. Determining the composite effects of urbanization on aerial insectivores represents an important research agenda as we continue to address declining aerial insectivore populations.

Acknowledgments

I would like to thank Kira Edic, Kaitlin Carr, Jenny Sanderson, Kate Gorman, and

Maggie Woodworth for their tireless work with field data collection and lab processing.

In addition, I would like to extend a special thanks to Jeffry Hayes, Reina Tyl, and

Danielle Vent for their role in directing data collection, developing field and lab work protocols, and managing the field teams. Thanks to Lars Meyer for logistical support in the field and in the lab. To Dr. David Manning and Kristen Diesburg – your help with R saved countless hours. To my committee members Dr. Stephen Matthews, Dr. Rachel

Gabor, and especially my advisor Dr. Mažeika Sullivan: I would like to express my gratitude for sharing your time, expertise, and knowledge throughout this process.

Vita

2014 …………………………………………. B.S., Science, Mathematics, and

Technology, SUNY Empire State College

2016 …………………………………………. Graduate Fellow, School of Environment

and Natural Resources, The Ohio State

University

2018 to present ………………………………. Graduate Research Assistant, School of

Environment and Natural Resources, The

Ohio State University

Fields of Study

Major Field: Environment and Natural Resources

Table of Contents

Abstract...... ii

Acknowledgments...... v

Vita...... vi

List of Tables...... ix

List of Figures...... xi

Chapter 1. Introduction...... 1

Chapter 2: Urbanization mediates the effects of water quality and climate on

Tree Swallow (Tachycineta bicolor) body condition and reproductive success.... 51

Abstract...... 52

Introduction...... 53

Methods...... 58

Results...... 67

Discussion...... 70

Conclusion...... 77

Chapter 3: Riparian aerial insectivorous bird trophic dynamics linked to urbanization of streams ...... 116

Abstract...... 117

Introduction...... 118

Methods...... 121

vii

Results...... 129

Discussion...... 131

Conclusion...... 138

Complete References...... 173

Appendix A. Chapter 2: Supplemental Material...... 196

Appendix B. Chapter 3: Supplemental Material...... 225

Appendix C. Permits...... 226

viii

List of Tables

Table 2.1 Study sites with impervious surface cover (% of total within 500-m on each side of the stream channel) and land-use designation based on impervious surface cover (Urban or Natural/Protected). Study reaches were categorized by adapting the thresholds developed by Schueler (1994): those < 25% impervious surface were categorized as Natural/Protected, and those reaches > 25% impervious surface were designated as Urban...... 103

Table 2.2 Water-chemistry variables for natural/protected vs. urban study sites, including means and standard deviations across the 4 years of the study...... 104

Table 2.3 Eigenvalues and the percent variance captured by the principal components (eigenvalues > 1), along with each principal component’s loadings and the proportion of the variance R2) each variable shared with the PCA axes. Only the first axis (PC1) was used for analyses...... 105

Table 2.4 Results from linear mixed-effects models with fixed (Year, Land Use [urban or protected], and Year × Land Use) and random (site, nestbox, site × nestbox). ** indicates a significant (p < 0.05) effect. * indicates evidence of a trend; i.e., 0.5 ≥ p < 0.10. Marginal R² = variation explained by fixed effects alone, while conditional R² = variation explained by both fixed and random effects...... 106

Table 3.1 Results from linear mixed-effects models with fixed (year, land use [urban or protected], and year × land use) and random (site, nestbox, site × nestbox). ** indicates a significant (p < 0.05) effect. * indicates evidence of a trend; i.e., 0.5 ≥ p < 0.10. Marginal R² = variation explained by fixed effects alone, while conditional R² = variation explained by both fixed and random effects. TP = trophic position. ADE = aquatically derived energy (i.e., nutritional subsidies originating from periphyton)...... 164

Table A.1 Study site coordinates, stream order, land use category (defined by % impervious surface), and % impervious surface coverage...... 197

Table A.2 Invertebrates collected 2014-2017, by year, site, season (early/late), transect (top/bottom), family, 10-day count, and mean mass (total family dry mass / count). Note that samples from both transects were pooled in 2015, so no transect is listed (entries represent totals for each family from both transects)...... 198

Table B.1 Carbon (ẟ13C) and nitrogen (ẟ15N) isotopic signatures for Tree Swallows. 226

Table B.2 Carbon (ẟ13C) and nitrogen (ẟ15N) isotopic signatures for dominant insect families...... 236 ix

Table B.3 Carbon (ẟ13C) and nitrogen (ẟ15N) isotopic signatures for primary production sources (detritus and periphtyton)...... 241

List of Figures

Figure 1.1 Percent population change among selected avian groups/guilds in Canada, 1970-2010. Source: Environment Canada, 2013...... 2

Figure 1.2 Tree Swallow range. Source: Cornell Lab of Ornithology, 2015...... 5

Figure 1.3 (a) 4-day-old Tree Swallow nestlings. Photo credit: Reina Tyl; (b) Adult male (left) and female (right) Tree Swallows. Photo credit. L. Villablanca...... 6

Figure 1.4 Food-web linkages between aquatic and riparian systems. Solid lines indicate strong pathways, while dashed lines indicate weaker pathways. From Sullivan & Rodewald (2012)...... 8

Figure 1.6 Female Tree Swallows in both the control and experimental groups (the latter with clipped wings to handicap foraging ability) produced, on average, smaller clutches as the breeding season advanced. From Nooker, Dunn, & Whittingham (2005)...... 20

Figure 1.7 Monthly means for day-to-day temperature variability for two paired North American urban-rural sites. Urban sites had consistently greater variability of the maximum daily temperature, while the reverse was true for rural sites. From Tam, Gough, & Mohsin (2015)...... 22

Figure 1.8 Example of a dual isotope (mean ± 1 SE) plot of 15N and 13C for aquatic and terrestrial food web components in the Scioto River basin (Ohio, USA) for (a) rural and (b) urban river reaches. Four-letter codes represent riparian swallow species. From Alberts, Sullivan, and Kautza (2013)...... 24

Figure 1.9 Map of study sites in the Scioto River system in metropolitan Columbus (OH, USA). Dark green-shaded areas indicate forested land cover. Source: Homer, et al., 2015 and QGIS Dev Team, 2017...... 25

Figure 2.1 Conceptual model showing the hypothesized relationships among urbanization, climate, water quality, flying insects, and aerial insectivorous birds. Local climate and land-use characteristics are expected to exert both direct effects (e.g., habitat availability) and indirect effects (via impact on invertebrate assemblages) on bird condition and reproductive success...... 107

Figure 2.2 Study sites and land cover in the greater Columbus, Ohio area. Source: Homer, et al., 2015 and QGIS Dev Team, 2017...... 108

Figure 2.3 Annual means from 2014-2017 by land use (i.e., protected or urban) for local climate variables across the breeding season (30 April-28 June): (a) air temperature °C, (b) humidity (%), (c) no. of days of extreme cold, and (d) number of days of extreme heat. Error bars indicate +/- 1 SE...... 109

Figure 2.4 Annual means from 2014-2017 by land use (i.e., natural or urban) for flying insect (a) family richness, (b) abundance (i.e., capture rate of individuals / m2 10-day-1, and (c) median body size (mg, dry mass). Emergent aquatic insects (left) and terrestrial insects (right) are shown separately. Error bars indicate +/- 1 SE...... 110

Figure 2.5 Annual means from 2014-2017 by land use (i.e., protected or urban) for reproductive response: (a) clutch size (no. eggs) (LMM: p = 0.710), (b) clutch initiation date (Julian date, calendar days) (LMM: p = 0.060) (c) no. successfully fledged (LMM: p = 0.009), and (d) nestling mass (g) (LMM: p = 0.150). Error bars indicate +/- 1 SE. ... 111

Figure 2.6 Relationships between the number of successfully fledged nestlings and (a) the Urban Stream Index (R2 = .16, F = 5.30, p < 0.001) and (b) the number of days of extreme cold between 30 April-28 June (R2 = .14, F = 4.31, p = 0.003). For the regression lines, red = 2014, green = 2015, blue = 2016, and purple = 2017...... 112

Figure 2.7 Relationships between (a) nestling mass (g) and the number of days of extreme heat between 30 April-28 June (R2 = 0.15 , F = 4.36, p < 0.003); and (b) clutch initiation date (Julian date, no. calendar days) and mean air temperature (°C) between 30 April and 29 May (R2 = 0.06 , F = 2.62, p = 0.039). For the regression lines, red = 2014, green = 2015, blue = 2016, and purple = 2017...... 113

Figure 2.8 Annual means from 2014-2017 by land use (i.e., protected or urban) for individual body condition: (a) Scaled Mass Index (g) for with females (F) (LMM: p = 0.680) and males (M) (LMM: p = 0.650), (b) blood mercury (Hg) concentration (ppb) for adults (LMM: p = 0.420) and nestlings (LMM: p = 0.920), and (c) blood glucose concentration (mg/dL) for adults and nestlings (LMM: p = 0.310). Error bars are +/- 1 SE. Note that there are only three years of data for blood mercury concentration (2014- 2016) and one year for glucose (2017). Glucose levels for adults were not tested due to small sample size, but their results are included for reference...... 114

Figure 2.9 Relationship between blood Hg concentration (ppb) in adult Tree Swallows and the Urban Stream Index (R2 = 0.53, F = 9.212, p < 0.001). For the regression lines, red = 2014, green = 2015, and blue = 2016. Hg was log10-transformed; however, the figure above shows the raw values...... 115

Figure 3.1 Tree Swallow study sites in urban and natural/protected areas in the greater Columbus, Ohio area. Source: Homer, et al., 2015 and QGIS Dev Team, 2017...... 165 xii

Figure 3.2 Annual means from 2015-2017 by land use (i.e., natural/protected or urban) for Tree Swallow trophic position of: (a) adults (LMM: p = 0.020) and (b) nestlings at ~13 days (LMM: p = 0.430). Error bars indicate +/- 1 SE. Different letters A, B indicate significant pairwise differences...... 166

Figure 3.3 Annual means from 2015-2017 by land use (i.e., natural/protected or urban) for Tree Swallow nutritional reliance on aquatically derived energy (e.g., originating from algae/periphyton) for (a) adults (LMM: p = 0.490) and (b) nestlings at ~13 days (LMM: p = 0.570). Error bars indicate +/- 1 SE. Different letters A, B indicate significant pairwise differences...... 167

Figure 3.4 Relationships between the Urban Stream Index (USI) for study sites and Tree Swallow trophic position for (a) 2015, (b) 2016, and (c) 2017. Separate multiple regression models with year as a categorical variable were developed for adults (R2 = 0.23, F = 6.04, p = 0.001) and nestlings (R2 = 0.22, F = 24.75, p < 0.001). For the regression lines, gold = adult swallows and green = nestlings at ~13 days...... 168

Figure 3.5 Relationships between Urban Stream Index and Tree Swallow nutritional reliance on emergent aquatic insects (vs. terrestrial flying insects) for (a) 2014, (b) 2015, and (c) 2016, and (d) 2017. Separate multiple regression models with year as a categorical variable were developed for adults (R2 = 0.10, F = 2.77, p = 0.036) and nestlings (R2 = 0.23, F = 23.40, p < 0.001). For the regression lines, pink = adult swallows and blue = nestlings at ~13 days...... 169

Figure 3.6 Composition of emergent aquatic and terrestrial flying insects (by percentage) in Tree Swallows diets across all study sites, 2014-2017. Sites are organized according to their Urban Stream Index (USI), with the lowest (i.e., least urbanized) on the left, and highest on the right. Note that positions of each plot are laterally equidistant from one another for display purposes, and do not necessarily reflect the magnitude of the USI. Blue = emergent aquatic insects, brown = terrestrial insects. Results for adult swallows are shown on the left side of each plot, nestlings on the right...... 170

Figure 3.7 Tree Swallow nutritional reliance on aquatically derived energy (e.g., originating from algae/periphyton) across all study sites, 2015-2017. Sites are organized according to their Urban Stream Index (USI), with the lowest (i.e., least urbanized) on the left, and highest on the right. Note that positions of each plot are laterally equidistant from one another for display purposes, and do not necessarily reflect the magnitude of the USI. Results for adult swallows are shown on the left side of each plot, nestlings on the right...... 171

Figure 3.8 Relationships between Tree Swallow nutritional reliance on emergent aquatic insects and reliance on aquatically derived energy (e.g., originating from algae/periphyton) for (a) early-season insects, (b) late-season insects. Separate linear regression models were developed for early-season insects and adult swallows (R2 = 0.01, F = 0.76, p = 0.390) early-season insects and nestlings (R2 = 0.01, F = 3.87, p = 0.050), xiii late-season insects and adult swallows (R2 = 0.01, F = 1.03, p = 0.318), and late-season insects and nestlings (R2 = 0.01, F = 3.41, p = 0.066). For the regression lines, pink = adult swallows and blue = nestlings at ~13 days...... 172

Figure B.1 Annual means from 2014-2017 by land use (i.e., natural/protected or urban) and period (May or July) for Tree Swallow nutritional reliance on emergent aquatic insects (vs. terrestrial flying insects). Flying insects prey are represented by the two most numerically abundant aquatic and terrestrial families in each season per study site. (a) Early season insects consumed by adults (LMM: p = 0.380), (b) early season insects consumed by nestlings at ~13 days (LMM: p = 0.310), (c) late season insects consumed by adults (LMM: p = 0.770), (d) late season insects consumed by nestling at ~13 days (LMM: p = 0.640). Error bars indicate +/- 1 SE. Different letters A, B indicate significant pairwise differences...... 245

xiv

Chapter 1. Introduction

Aerial insectivorous birds – a guild comprising swallows, nightjars, swifts, and flycatchers – have experienced alarming population declines in eastern North America in recent decades (Environment Canada, 2012; Nebel et al., 2010) (Figure 1.1). Although the cause of these population declines is not yet fully understood, they are likely anthropogenic in origin.

Possible environmental attributes connected to the decline of aerial insectivorous bird populations include the availability and quality of flying insect prey populations which, in turn, are shaped by environmental conditions (Nebel et al., 2010). Although declines in individual bird species may be linked to other causes, including loss or fragmentation of suitable breeding habitat (MacHunter et al., 2006) and degradation or loss of habitat in tropical wintering grounds

(Fraser et al., 2012), similar declines across multiple, taxonomically diverse species in the guild indicate that changes in insect prey are likely a common factor.

Figure 1.1 Percentage population change among selected avian groups/guilds in Canada, 1970-2010. Source: Environment Canada, 2013.

Riparian zones can be hotspots of aerial insectivorous bird diversity (Naiman & Decamps, 1997).

Although aerial insectivorous birds are terrestrial organisms, they are often highly dependent on aquatic ecosystems (Uesugi & Murakami, 2007). Therefore, changes in the condition of streams, rivers, lakes, and wetlands might be expected to influence many facets of aerial insectivorous bird ecology, including nutrition and energetics (e.g., Kautza & Sullivan, 2016), nearshore and riparian habitat conditions (Naiman & Decamps, 1997), health (Smits & Fernie, 2013), and reproductive success (McCarty & Secord, 1999). In particular, predator-prey relationships between birds and aquatic insects may be shaped by changes in chemical water quality, vegetation structure and cover, and environmental contaminants (Alberts, Sullivan, & Kautza,

2013).

Additional lines of evidence suggest that urbanization of the landscape may contribute to aerial insectivorous bird population declines (Nebel et al., 2010; Schlesinger, Manley, &

Holyoak, 2008). Urban encroachment in the riparian zone is associated with profound changes in avian community composition, population density (Blair, 1996; Miller et al., 2003; Rodewald &

Bakermans, 2006), and reproductive success (Rodewald, Kearns, & Shustack, 2013). These changes have been linked to the consequences of urban land use change, such as the removal of riparian vegetation (Lussier et al., 2006), habitat fragmentation and altered vegetation structure

(Crooks, Suarez, & Bolger, 2004), and the increased presence of invasive biota (Rottenborn,

1999). Further, intensification of urbanization is associated with decreased invertebrate density and species richness (Macivor & Lundholm, 2011; Urban et al., 2006) and this relative paucity of available prey may be implicated in reduced reproductive success observed among aerial insectivores breeding in urban areas (Teglhøj, 2017). Environmental contaminants can also have

3 a considerable impact on wild bird populations. Mercury, for instance, has been related to diminished reproductive success, even at relatively low blood concentrations (Rowse, Rodewald,

& Sullivan, 2014), and urbanization of watersheds is associated with increased mercury concentrations in stream sediment (Chalmers et al., 2014; Lutz et al., 2009).

Finally, the effects of climate change may be exerting quantitatively important effects on avian reproductive success and survival (Both et al., 2010; Hussell, 2003). Long-term data on

Tree Swallows (Tachycineta bicolor) in Canada revealed a negative impact on fledging success and brood size associated with both increased air temperatures and higher rainfall during the breeding season (McArthur et al., 2017). Changes in climate and local weather conditions may also exacerbate the effects of other environmental variables. For instance, evidence suggests an interactive effect between increased air temperatures and the aforementioned reproductive impacts associated with mercury exposure (Hallinger & Cristol, 2011). Considered together, the complex relationships between climate, water quality, and land use change demand a broad perspective to address their impacts on aerial insectivorous birds.

Background

Tree Swallow – a model aerial insectivore

The Tree Swallow is a secondary cavity-nesting passerine associated with riparian areas

(Rendell & Robertson, 1989). The species is migratory, wintering along the Gulf Coast, in

Mexico, the Caribbean, and Central America before returning north in the early spring (Butler,

1988; Robertson, Stutchbury, & Cohen, 2011). The Tree Swallow’s breeding range extends coast to coast through most of North America, from Alaska to northernmost California in the west, and

Newfoundland south to northern Georgia in the east (Robertson, Stutchbury, & Cohen, 2011)

(Figure 1.2).

Figure 1.2 Tree Swallow range. Source: Cornell Lab of Ornithology, 2015.

Tree Swallows are primarily aerial insectivores, catching flying insects on the wing.

Although adult Tree Swallows can supplement their diets with vegetable matter such as bayberries during winter or early spring migration (Piland & Winkler, 2015), adults and young are dependent on flying insects during the breeding season (Mengelkoch, Niemi, & Regal, 2004;

Nooker, Dunn, & Whittingham, 2005). Both male and female parents (Fig. 1.3a) forage for flying insects to feed nestlings (Fig. 1.3b), with increased rates of feeding as nestlings age

(Leffelaar & Robertson, 1986). Adult swallows provision nestlings frequently throughout the day, delivering a bolus of captured insects to nestlings on each return trip to the nest (McCarty,

2002). Tree Swallows prey on a wide variety of invertebrates, primarily flies, true bugs, and dragonflies and damselflies, but also many mayflies, caddisflies, beetles, wasps, stoneflies, other flying insects and occasionally spiders and other nonflying terrestrial invertebrates (Beck,

Hopkins, & Jackson, 2014; Mengelkoch, Niemi, & Regal, 2004; Quinney & Ankney, 1985).

When foraging, invertebrate body size may be a more important factor in prey selection than taxonomy (Hespenheide, 1971), as swallows preferentially select larger-bodied prey (McCarty &

Winkler, 1999a). Nevertheless, small-bodied insects (<10mm in length) still often constitute the larger proportion of their diets (Quinney & Ankney, 1985).

Figure 1.3 (a) Adult male (left) and female (right) Tree Swallows. Photo credit. L. Villablanca. (b) 4-day-old Tree Swallow nestlings. Photo credit: Reina Tyl.

Tree Swallows are cavity nesters, and reproduction may therefore be site-limited by the number of available cavities (Stutchbury & Robertson, 1985). Clutch initiation varies depending on latitude. In Ohio, laying usually commences in May, though the timing of reproduction has been linked to air temperature, food abundance, intraspecific competition, and geography (Dunn

& Winkler, 1999; Nooker, Dunn, & Whittingham, 2005). A mated pair of Tree Swallows will typically have a single brood in a season, though two broods are not uncommon in the southern parts of their range (Monroe et al., 2008). Female swallows lay one egg per day, with a typical clutch size of 5 or 6 eggs (Rendell & Robertson, 1993); however, clutch size varies and increased food availability has been linked to larger clutch size (Hussell & Quinney, 1987). Hatching occurs 14-16 days after laying, and nestlings fledge 18-21 days after hatch (Hussell & Quinney,

1987). Ambient temperature and food availability have been associated with nestling growth rates and nestling mass (McCarty & Winkler, 1999b). Lifespan is about three years, on average, though individuals may live up to eight years (Butler, 1988).

Energetic linkages between aquatic and terrestrial ecosystems

Many riparian terrestrial consumers, including Tree Swallows and other aerial insectivores, are nutritionally linked to aquatic ecosystems via their dietary reliance on adult, emergent aquatic insects (Figure 1.4). Cross-system subsidies are a key component of the relationship between streams and the adjacent riparian zones (Baxter, Fausch, & Saunders, 2005;

Power & Dietrich, 2002). The food webs of streams and riparian areas are connected via the transfer of nutrients, prey, and detritus (Polis, Anderson, & Holt, 1997). Fluxes of emergent aquatic insects constitute a critical resource subsidy to adjacent terrestrial food webs, including

7 insectivorous avian consumers (Nakano & Murakami, 2001). In the other direction, allochthonous subsidies (i.e., those originating outside the stream system), such as leaf litter and terrestrial invertebrates (Nakano & Murakami, 2001; Wallace et al., 1997) provide important resource inputs to the recipient aquatic systems (Richardson & Sato, 2015).

Figure 1.4 Food-web linkages between aquatic and riparian systems. Solid lines

indicate strong pathways, while dashed lines indicate weaker pathways. From The importance of insect prey to avian consumers is evidenced by the latter’s Sullivan & Rodewald (2012).

distribution, and evidence strongly suggests that insect prey availability can shape the spatial distribution of insectivorous birds. In fact, distributions of insects have been found to exert substantial influence over the distribution of many terrestrial-consumer taxa (Burdon & Harding,

2008; Kautza & Sullivan, 2016b; Muehlbauer et al., 2014). Prey availability has emerged as a strong predictor of avian insectivore distribution, with fluctuations in avian abundance tracking emergent aquatic insect abundance within season for some species (Murakami & Nakano, 2002;

Razeng & Watson, 2015).

Even insectivorous birds nesting in areas spatially distant from riparian zones may rely on aquatic systems for food. Uesugi and Murakami (Uesugi & Murakami, 2007) found that emergent aquatic insects can constitute an important food source for insectivorous upland forest birds as well, which congregate and forage in riparian areas in the early spring prior to bud break. Insect abundance, including that of terrestrial insect taxa, may be more strongly associated with riparian vegetation rather than upland vegetation in forested landscapes, driving many bird species to forage near streams (Hagar et al., 2012). Evidence suggests that insectivorous birds will alter the geographic scope of their foraging habits to take advantage of seasonal fluctuations in prey subsidies from both aquatic and terrestrial sources (Uesugi & Murakami, 2007). This effect is especially pronounced among many terrestrial consumers that move toward aquatic zones to take advantage of insect subsidies during peak emergence (Baxter, Fausch, & Saunders,

2005). For instance, seasonal changes in insect emergence have been linked to local densities of insectivorous birds across different habitat types (Gray, 1993). Insectivorous bird abundance also may be strongly influenced by insect assemblage composition and distribution (Hagar et al.,

2012; Hussell & Quinney, 1987).

Among aquatic insects, the orders Ephemeroptera (mayflies), Plecoptera (stoneflies), and

Trichoptera (caddisflies) (collectively referred to as EPT) are generally more sensitive to changes in water quality and therefore the richness and relative abundance of these taxa are commonly

9 used as indicators of stream condition (Compin & Céréghino, 2003; Lenat, 1988). The relatively large-bodied EPT taxa insects are typically associated with less-disturbed streams and serve as an important food source to many aerial insectivores, including Tree Swallows (Hussell &

Quinney, 1987; McCarty & Winkler, 1999a). Research on the prey preferences of other guilds of insectivorous birds has produced similar findings. In temperate forests, gleaners and ground foragers were observed to prefer prey taxa with higher body fat and micronutrient content than low-frequency prey taxa (Razeng & Watson, 2015). Similarly, Heinrich, Whiles, and Roy observed that the abundance of larger-bodied insect taxa was correlated with avian abundance and species richness (2014). Experimental evidence also suggests that aquatic insects confer an energetic advantage over terrestrial insects. In a study that underscores that importance of aquatic insects for insectivorous birds, Twining et al. (2016) showed that Tree Swallow nestlings exhibited faster growth rates and better body condition when provisioned with aquatic insects rich in long-chain omega-3 polyunsaturated fatty acids. Twining et al. (2018) (Fig. 1.5a) further demonstrated that the availability of high-quality aquatic insect prey (Fig. 1.5b) was related to increased fledging success among Tree Swallows.

Figure 1.5 (a) Experimental treatment reaction norms for Tree Swallows for six body condition metrics. Black represents diets low in long-chain omega-3 polyunsaturated fatty acids; gray represents diets low in these fatty acids. From Twining, et al. (2017). (b) Magnified contents of a single Tree Swallow bolus collected in June 2017 in Columbus, OH (USA). Insects in bolus comprise 23 caddiesflies (Family Hydropsychidae) and 2 long-legged flies (Family Dolichopodidae, not shown).

Just as aerial insectivorous birds are influenced by distributions of emergent insect prey, the composition and distribution of aquatic insect assemblages can be shaped by their environment. For instance, stream discharge, morphology, and other stream characteristics regulate aquatic insect production and, consequently, are strongly connected to the distribution of

11 a variety of insectivorous birds (Gray, 1993). Water chemistry can also strongly impact aquatic insects (Allan & Flecker, 1993; Johnson et al., 2013). Human alteration and management of the landscape can have profound and far-reaching effects on both terrestrial and aquatic ecosystems.

Anthropogenic change to watersheds, including deforestation, agricultural use, impoundments, stream channelization, and the introduction of invasive organisms can have strong effects on populations of both riparian birds and their insect prey (Fausch, Baxter, & Murakami, 2010;

Kautza & Sullivan, 2015; Rioux-Paquette et al., 2014). In a study in the Scioto River system of

Ohio, Kautza and Sullivan (2015) found that higher densities larger-bodied taxa were associated with rural river reaches compared to river reaches in developed landscapes. Strasevicius et al.

(2013) found that emergent aquatic insect subsidies to adjacent forests were reduced by a full order of magnitude along river reaches in Sweden modified for hydropower production.

Availability of insect prey is associated with egg quality (Ardia, Wasson, & Winkler,

2006), nestling survival and body condition (Quinney et al., 1986), brood size and body mass

(Teglhøj, 2017), nestling immune response (Lifjeld, Dunn, & Whittingham, 2002), and oxidative damage in adult individuals (Stanton, Lark, & Morrissey, 2017) among various aerial insectivorous bird species. Overall, evidence suggests aerial insectivores suffer from the diminished abundance and quality of their insect prey. For example, Poulin, Levebre and Paz

(2010) demonstrated a link between the suppression of flying invertebrates via pesticide applications and reduced clutch sizes and fledging success in European House Martins (Delichon urbicum).

Reductions in emergent aquatic insect prey subsidies are reflected in the reproductive success of avian consumers. Research on insectivorous Pied Flycatchers (Ficedula hypoleuca)

12 has shown that reduced aquatic insect subsidies are linked to lower hatchling survival rates

(Strasevicius et al., 2013). The quality of prey is important as well. As songbird fledgling mass has been strongly linked to post-fledging avoidance of predation (Naef-Daenzer & Grüebler,

2016), the diminished condition of smaller swallows could impair their odds of survival.

Environmental variability

Land-use and land-cover change

As rivers and streams are inexorably connected to their surrounding landscapes, the character of those landscapes in instrumental in determining the conditions of the lotic systems that flow through them (Allan, 2004). Consequently, shifts in land use and land cover can have significant impacts on both insectivorous birds and their invertebrate prey. Reduced taxonomic diversity of birds and invertebrates is associated with the built environments of urban areas

(McKinney, 2002). Diversity of invertebrates in urbanized landscapes may be significantly reduced due to habitat loss and fragmentation, pollution of soil and water, and the proliferation of introduced exotic organisms (McIntyre, 2000). Linked aquatic-riparian ecosystems are affected by land use change in multiple ways, as changes in upstream riparian zones can have sizeable impacts on aquatic systems downstream by altering flow regimes, stormwater discharge, and sediment loads (Booth & Jackson, 1998). Engineering of stream channels to accommodate development usually results in reduced channel complexity and a concomitant homogenization of in-stream habitat (Walsh et al., 2005), with negative impacts on emergent insect density

(Iwata, Nakano, & Murakami, 2003). In contrast, less-developed riparian zones are typically characterized by substantial differences in the taxonomic composition of aquatic insect

13 assemblages, with more sensitive taxa more strongly represented. For example, evidence indicates that lower densities of residential land use are associated with increases in

Ephemeroptera (Freeman & Schorr, 2004; Lussier et al., 2008).

Deforestation is associated with many kinds of land-use changes and may substantially alter aquatic insect assemblages in both urban and rural stream reaches. Besides providing energy inputs, terrestrial leaf debris provides habitat for many aquatic insects, enhancing in- stream invertebrate diversity (Compson et al., 2013). Larger materials from trees, like large wood, also provide nutrients and habitat to aquatic biota, and the magnitude and composition of these inputs can be altered by landscape changes (Richardson & Sato, 2015). Alteration or removal of riparian vegetation may influence birds’ spatial foraging patterns; for instance, Tree

Swallows living in urban riparian areas may be more reliant on aquatic subsidies for food, compared to their rural counterparts (Alberts, Sullivan, & Kautza, 2013).

Among land-use changes, urbanization may produce the most pervasive and long-lasting impacts (Ren et al., 2003). Commercial, industrial, and residential land use has been associated with nutrient pollution, increased sediment loads, and the loss of sensitive benthic insect taxa from streams (Klein, 1980). Erosion and deposition of sediment increase substantially in urban stream reaches due to the presence of impervious surfaces, increased flow, and higher localized stream gradients (Meyer, Paul, & Taulbee, 2005; Walsh et al., 2005). Channelized streams in developed landscapes, subject to increased runoff, are often characterized by a more homogenous in-stream morphology and altered substrate size (Arnold, Boison, & Patton, 1982;

Grable & Harden, 2006). Impaired streams in more developed areas were found to exhibit features of Urban Stream Syndrome; i.e., elevated levels of nutrients and toxic contaminants,

14 altered flow regimes with a flashier hydrograph, wider channels, higher water temperatures, fewer sensitive aquatic taxa, and simplified stream morphology (Vietz, Walsh, & Fletcher, 2015;

Walsh et al., 2005). Nitrogen and phosphorus enrichment is typically higher in streams in developed landscapes (Paul & Meyer, 2001; Tromboni & Dodds, 2017). Eutrophication is associated with increased algal biomass, with negative impacts on aquatic systems including elevated pH, reduced light penetration, and depletion of dissolved oxygen (Chislock et al., 2013).

Together, these changes have well-documented harmful consequences for aquatic biota and ecosystem function (Dodds & Smith, 2016; Smith, Tilman, & Nekola, 1998). Other adverse effects on aquatic systems associated with urbanized land use changes associated include increased salinity, reduced dissolved oxygen content, increased sedimentation, and elevated levels of coliform bacteria (Peters, 2009). Although stream restoration projects have become increasingly popular endeavors in recent decades, the long-term impacts of urbanization have proved challenging to reverse (Kondolf, 1995). Even restored urban streams may exhibit diminished aquatic invertebrate richness, underscoring the pervasive character the urban environment’s influence at the watershed scale (Violin et al., 2011).

Habitat loss and impairment strongly influence aquatic invertebrate communities and often lead to substantial declines in species richness of benthic macroinvertebrates on a gradient toward the urban center (McKinney, 2002; Walsh et al., 2001). Overall aquatic insect diversity has been shown to be negatively associated land development in the watershed, particularly the increased land coverage by impervious surfaces (Gage, Spivak, & Paradise, 2004). In such environments, the diversity of aquatic insect assemblages is diminished and the relatively small- bodied, tolerant Chironomidae (midges) are frequently the dominant insect family (Heinrich,

Whiles, & Roy, 2014). In contrast, insects of the EPT taxa are often sensitive to poor water quality and are suppressed by the poor stream conditions brought on by urbanization (Roy et al.,

2003).

Environmental contaminants

Environmental contaminants in aquatic systems can affect survival and reproductive success of both aquatic and terrestrial organisms, as linked aquatic-terrestrial food webs can facilitate the bioaccumulation and transfer of trace contaminants to terrestrial consumers. PCBs, for example, are readily transferred by aquatic insects to terrestrial consumers with associated negative impacts on the individual health and reproductive success of songbirds (McCarty &

Secord, 1999; Walters, Fritz, & Otter, 2008) including Tree Swallows (Custer et al., 2003).

Similar effects on Tree Swallow populations have been shown for other contaminants found in aquatic systems, including selenium (Ohlendorf et al., 1988), various contaminants associated with mine tailings (Gentes et al., 2006) and wastewater effluent (Dods et al., 2005), including pharmaceuticals (Richmond et al., 2018).

Among the environmental contaminants found in aquatic systems, mercury is one of the most pervasive and hazardous, and consequently one of the best-studied (Morel, Kraepiel, &

Amyot, 1998). Concentrations of environmental mercury are relatively high in eastern North

America (Driscoll et al., 2007). Elevated levels of mercury in the environment are a byproduct of human activities, particularly combustion of fossil fuels, which release mercury into the atmosphere (Pirrone et al., 2010). Because mercury is both persistent and mobile in the atmosphere, it can be distributed far beyond its point sources (Li & Cai, 2013). As a pollutant,

16 mercury can be hazardous to organisms in its inorganic form, but it becomes a far more potent neurotoxin when converted by microbial processes to organic methylmercury (Boening, 2000;

Munthe et al., 2007). Methylmercury biomagnifies in food webs, with the greatest impact on consumers at the highest trophic levels (Driscoll et al., 2007).

A number of studies have shown a relationship between blood-mercury levels and impaired reproduction in various aquatic and marine birds, including reduced hatchability, lower chick survival rates, and lower chick weights (Burger & Gochfeld, 1997; Fimreite, 1974). Long- term exposure to dietary methylmercury in passerine birds is associated with reduced fledging success (Varian-Ramos, Swaddle, & Cristol, 2014). Similar responses to elevated blood- mercury levels have been observed in insectivorous passerines that nest in riparian habitats, including Tree Swallows (Brasso & Cristol, 2008) and Carolina Wrens (Jackson et al., 2011).

Hallinger and Cristol (Hallinger & Cristol, 2011) demonstrated that elevated levels of blood mercury in adult Tree Swallows were associated with diminished hatching and fledging success.

At higher concentrations, elevated mercury levels can impair immune system response in adult

Tree Swallows (Hawley, Hallinger, & Cristol, 2009). The negative impact of methylmercury may manifest at concentrations too low to produce health impacts in adult birds. For instance, a study of Acadian flycatchers by Rowse, Rodewald, and Sullivan (2014) revealed a strong relationship between elevated blood mercury levels and lower reproductive success, even when mercury concentrations were below the threshold of a predicted negative health response.

The transfer of mercury from aquatic systems to terrestrial consumers can be strongly mediated by aquatic insects (Cristol et al., 2008; Powell, 1983, Sullivan & Rodewald 2012), which suggests that biomagnification of mercury will be high in insectivorous birds, due to their

17 trophic position (Evers et al., 2005). Differences in environment and life history of aerial insectivores play a role in exposure to mercury. Prey selection plays an important role in the transfer of contaminants, suggesting that biomagnification will vary across species of insectivorous birds (Beck, Hopkins, & Jackson, 2014). In addition, land-use and land-cover changes may play a regulatory role in the flux of these contaminants from aquatic systems to terrestrial consumers. Aquatic systems in areas experiencing elevated deposition of inorganic mercury will not necessarily exhibit higher concentrations of methylmercury in aquatic organisms (Chalmers et al., 2014). Lastly, land-use changes can alter the relationship between aquatic systems and terrestrial consumers by altering insect food assemblages and influencing the spatial extent of avian foraging efforts (Alberts, Sullivan, & Kautza, 2013).

Climate change

Climate change has the potential to exert substantial impacts on bird populations and their phenology, including the timing of breeding and migration, reproductive success, and the ranges of individual species (Crick, 2004; McCarty, 2001). In Ohio, 55% of bird species are projected to be climate threatened or climate endangered (i.e., in danger of losing more than 50% of their current range by 2080 or 2050, respectively) if present warming trends continue (Langham et al.,

2014). Climate change may impact aerial insectivorous birds through both direct and indirect means. Warming climates are likely driving phenological changes in Tree Swallows (Winkler,

Dunn, & McCulloch, 2002). Specifically, emerging evidence shows that increasing spring temperatures are related to earlier clutch initiation dates (Bourret et al., 2015; Hussell, 2003); in

18 fact, long-term data from 1959-1991 show a 9-day advance in Tree Swallow clutch initiation

(Dunn & Winkler, 1999).

Experimental treatments suggest that Tree Swallows commence breeding based on immediate food availability, rather than aligning breeding to take advantage of peak food availability for hatchlings (Nooker, Dunn, & Whittingham, 2005), highlighting the influence of early-spring insect emergence as a driver of clutch initiation. Climate change is predicted to exert a similar influence on emergent aquatic flying insects, with potentially adverse effects on birds from the temporal decoupling of hatching and food availability (Visser et al., 1998). Both climate change and encroaching urbanization tend to warm streamwater temperature (Nelson &

Palmer, 2007). Streamwater temperatures have shown a strong correlation with changes in air temperature (Pilgrim, Fang, & Stefan, 1999), with substantial impacts projected for aquatic invertebrates (Durance & Ormerod, 2007). Rising temperatures throughout the year may have an impact on aquatic insect populations, as water temperature serves as a critical cue for insect emergence (Learner & Potter, 1974). Warmer winter temperatures are associated with earlier emergence (Nebeker, 1971). Experimental evidence has shown that changes in water temperature can influence both the timing of aquatic insect emergence (Harper & Peckarsky,

2006; Nordlie & Arthur, 1981) and the composition of emergent insect assemblages in streams

(Jonsson et al., 2015). Extended periods of warm temperatures may spur early incubation of eggs as well, possibly advancing hatch dates in Tree Swallows and promoting asynchronous hatching

(Ardia, Cooper, & Dhondt, 2006). Owing to the importance of aquatic insect emergence for breeding swallows, it is possible that shifts to the timing of clutch initiation, hatching, and fledging may become increasingly asynchronous with aquatic insect emergence, limiting Tree

Swallows from fully exploiting a critical food source (Thomas et al., 2001). Experimental treatments conducted by Nooker, Dunn, and Whittingham (2005) (Fig. 1.6) provided evidence to support such links among insect emergence, clutch initiation, and Tree Swallow reproductive success: earlier clutch initiation was related to insect abundance and, in turn, larger clutches and heavier nestlings.

Apart from the direct impact of climate on the abundance of composition of insect prey assemblages, a changing climate may have other detrimental effects on insectivorous bird populations. Meteorological conditions may affect the foraging success of birds in complex ways, via impacts on the foraging capabilities of adult birds, or by altering flying insect activity.

For instance, bird broods hatched early may succumb to cold snaps, either from direct thermal

20 stress or due to a paralysis of the flying insect food supply (Winkler, Luo, & Rakhimberdiev,

2013) as the ability of insects to fly is suppressed by low temperatures (Taylor, 1963). Similarly, unusually warm weather may limit nestling growth. Pipoly et. al (2013) showed a negative relationship between periods of extreme heat and body condition in House Sparrows (Passer domesticus), another cavity-nesting songbird. Precipitation during the breeding season may also play a role, as sustained high rainfall has been associated with diminished reproductive success in Tree Swallows, either due to the suppression of foraging by adult birds or, more directly, by increased mortality of nestlings due to nest saturation (McArthur et al., 2017).

Conversely, increasing temperatures may also have a beneficial impact on aerial insectivorous birds under certain conditions. Experimental treatments by Perez et al. (2008) observed that Tree Swallow nestlings in artificially-heated nests exhibited better body condition, likely due to increased provisioning activity by adult females. The work of Pipoly et al. (2013) also demonstrated that, despite the unfavorable consequences of extreme heat on nestling body condition, overall higher mean temperatures during the breeding season were associated with increased nestling body mass. Similarly, McCarty and Winkler (1999b) revealed that the positive effect of temperature was strongly evident in the growth of Tree Swallow nestlings.

Global climate change is not the only driver behind variation in climate conditions.

Urbanization also has dramatic local impacts on temperature variability, particularly in large cities, where mean temperatures may be several degrees higher than in adjacent rural areas (Oke,

1982). Built environments with extensive impervious surfaces are characterized by diminished evapotranspiration and limited shade due to reduced vegetation, lower surface albedo, reflective vertical surfaces, suspended airborne particulate matter, and anthropogenic heat sources, all of

21 which contribute to the urban heat island effect (Oke, 1982; US EPA, 2008). Notably, among the

60 largest cities in the United States, Columbus was identified as the 8th most intense of its urban heat island from 2004-2014 (Kenward et al., 2014). Patterns of daily temperature variation differ as well. Tam, Gough, and Mohsin (2015) (Fig. 1.7) observed that urban areas had greater day-to- day temperature variability than surrounding rural areas, driven by greater variability in maximum daily temperatures and, in contrast, more stable minimum temperatures. Conversely, rural sites had greater heat loss and therefore greater temperature variability at night. The differences in temperature regimes illustrate the outsized role played by urbanization of the landscape.

Food webs and stable isotope analysis

The nature of food webs can reveal the relationships between terrestrial secondary consumers (e.g. aerial insectivorous birds), their diets, and trophically linked aquatic-riparian ecosystems (e.g., Collier, Bury, & Gibbs, 2002; Kautza & Sullivan, 2016b). Analysis of stable isotopes is a key tool to ascertain these pathways (Layman et al., 2012; Peterson & Fry, 1987).

Isotopes of carbon (12C and 13C) and nitrogen (14N and 15N) are commonly used in such analyses

(Post, 2002). These stable (i.e., not radioactive) isotopes occur naturally and each pair is identical in terms of most chemical reactions; however, they can be differentiated by their nuclear masses, an attribute which allows the ratio of isotopes to be measured (Fry, 2006).

Enrichment of 15N relative to 14N occurs in consumers with increases in trophic level (Minagawa

& Wada, 1984). In contrast, ratios of carbon isotopes are more stable as carbon moves through food webs (Post, 2002), and they reflect the photosynthetic pathways of primary production

(Rounick & Winterbourn, 1986). By examining the ratios of these isotopes, the dietary contributions of terrestrial and aquatic sources – as well as of terrestrial vs. aquatic insects – to aerial insectivorous birds can be estimated (e.g., Kautza & Sullivan, 2016), as can trophic position and food chain-length (e.g., Kautza & Sullivan, 2016a; Post, Pace, & Hairston, 2002) .

In broad strokes, ratios of 15N in tissues is useful for estimating a consumer’s trophic position

(Cabana & Rasmussen, 1996; Minagawa & Wada, 1984), while 13C ratios yield information about dietary sources (Bearhop et al., 1999, 2002; Caquet, 2006; DeNiro & Epstein, 1978;

Harvey & Kitchell, 2000) (Fig. 1.8).

Study system

My study area consists of seven river-riparian sites on an urban-forested gradient in the

Scioto River system (Ohio, USA). The Scioto River watershed drains 16868-km2 as it flows south into the Ohio River (Ohio Environmental Protection Agency, 2014). The central portion of the Scioto River catchment lies within the Columbus Metropolitan Area. The study sites are 24 located on the Scioto, a 6th-order river, plus two major tributaries, the Olentangy River and Big

Darby Creek (Fig. 1.9). The regional landscape is composed overwhelmingly of developed

(45%) or cultivated (40%) lands, with forests comprising only 6% of the total land cover (Ohio

Environmental Protection Agency, 2014).

Figure 1.9 Map of study sites in the Scioto River system in metropolitan Columbus (OH, USA). Red circles indicate the urban study sites. Dark green- shaded areas indicate forested land cover. Source: Homer, et al., 2015 and QGIS Dev Team, 2017.

Objectives

The overarching objective of this research was to quantify the relationships between urbanization, water quality, climate, and aerial insectivores across an urban-rural landscape gradient in central Ohio’s Scioto River basin. This research builds on several years of data

collected by Dr. Sullivan’s Stream and River Ecology (STRIVE) Lab in the greater Columbus,

Ohio area since 2011. I addressed this objective through the following specific goals:

1. Investigate linkages among landscape urbanization, local climate, water quality, and reproductive

success and body condition of Tree Swallows.

a. Quantify land-use and land-cover (LULC) at the local (e.g., 500-m buffer) and landscape

(e.g., catchment draining to each study site) scale at seven river reaches in and around

Columbus, Ohio.

b. Assess chemical water-quality conditions (e.g., turbidity, pH, contaminants) at study

reaches.

c. Assess composition, abundance, and availability of aquatic and terrestrial insect

assemblages at study reaches.

d. Monitor air temperature and humidity at Tree Swallow nest boxes.

e. Evaluate potential relationships between Tree Swallow reproductive metrics and

temperature and humidity throughout the breeding season.

f. Relate landscape urbanization and water quality directly and indirectly (via emergent

insects) to Tree Swallow reproductive success (e.g., clutch initiation date, clutch size,

nestling mass, fledging success) and body-condition/health metrics (e.g., mass,

contaminant loads).

2. Investigate influence of urbanization on trophic dynamics of Tree Swallows.

a. Assess composition, abundance, and availability of aquatic and terrestrial insect prey

assemblages at study reaches.

b. Quantify isotopic signature of insect prey as well as aquatic and terrestrial primary

producers at study reaches.

c. Investigate, via stable isotope analysis, proportions of Tree Swallow energetic sources

derived from pathways originating from aquatic primary production relative to terrestrial

primary production across the study reaches.

d. Evaluate, via stable isotope analysis, Tree Swallow trophic position across the study

reaches.

The Tree Swallow (Tachycineta bicolor) serves as an ideal model North American aerial insectivore for this study. Tree Swallows are common in the eastern United States, easily observed, and readily make use of artificial nest boxes (Robertson, Stutchbury, & Cohen, 2011).

The latter characteristic offers a great benefit for field research, as investigators can limit external perturbations (e.g., predation, interspecific competition, and nest parasitism) and capture both juveniles and adults. Tree Swallows are also useful models for the effects of environmental contamination, particularly in aquatic systems, since they commonly feed on emergent aquatic insects (Jones, 2003). I anticipate that my investigation of the linkages among climate, water quality, and aerial insectivores will help to reveal the causes of aerial insectivore population declines, and thereby inform conservation and wildlife management efforts in Ohio and throughout North America.

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Winkler, D. W., Luo, M. K., & Rakhimberdiev, E. (2013). Temperature effects on food supply and chick mortality in tree swallows (Tachycineta bicolor). Oecologia, 173(1), 129–138. https://doi.org/10.1007/s00442-013-2605-z

Chapter 2. Urbanization mediates the effects of water quality and climate on Tree Swallow (Tachycineta bicolor) body condition and reproductive success

Authors: Joseph W. Corra and S. Mažeika P. Sullivan

Abstract

Aerial insectivorous birds – swifts, swallows, flycatchers, and nightjars – have experienced alarming population declines in eastern North America. Meanwhile, urbanization continues to increase rapidly, with urban land use comprising 69.4 million acres (3.6% of total) in the contiguous United States as of 2018. Multiple environmental changes are associated with urbanization, including alterations to local climate, changes in habitat structure, and potential shifts in both terrestrial and emergent aquatic flying insects on which aerial insectivorous birds depend. In particular, emergent aquatic insects have recently been shown to provide energetic advantages to aerial insectivorous birds as compared to terrestrial insects, yet are highly sensitive to losses in water quality. Here, we investigated the linkages between urbanization, water quality, and Tree Swallow (Tachycineta bicolor) reproductive success and body condition at seven river-riparian sites representing urban and natural/protected land use in Columbus, Ohio over four consecutive springs (2014-2017). Urban nests were associated with higher fledging success and earlier laying dates. Nestling mass was not related to land use, but exhibited high interannual variability, as did body condition in adult males and Hg in both adult birds and nestlings. The interaction of year × land use also had a significant influence on Hg in nestlings, suggesting that contaminant transfers to nestlings at urban sites are related to shifts in the composition of insect prey assemblages. Multiple characteristics of urban sites appeared to drive these patterns including differences in mean and extreme air temperatures, and measures of water quality (e.g., water temperature, nutrient concentrations, turbidity). Overall, despite the loss of environmental quality generally attributed to cities, Tree Swallows exhibited greater reproductive success in urban settings where aquatic insects were larger, climate more amenable

52 to egg and nestling survival, and the breeding season longer. However, chronic effects of elevated Hg body burdens in urban areas may disadvantage individuals in other ways. Further, characteristics of urban landscapes that benefit Tree Swallows may not advantage other aerial insectivorous birds owing to differences in life-history and foraging strategies. These findings implicate urbanization, local climate, and water quality as important considerations in the conservation of riparian aerial insectivorous birds.

Introduction

Aerial insectivorous birds – a guild comprising swallows, nightjars, swifts, and flycatchers – have experienced alarming population declines in North America, with population losses outpacing declines observed in any other avian group (Environment Canada, 2012; Nebel et al., 2010). Population trends among species in the guild exhibit spatial and temporal variation, but there is broad evidence for intensified declines for most species beginning in the 1980s (A.

C. Smith et al., 2015). Population declines in individual bird species are likely in response to a complex set of conditions, including loss of suitable breeding habitat (MacHunter et al., 2006), predation (Bohning-Gaese, Taper, & Brown, 1993), and environmental degradation or habitat loss on wintering grounds (Fraser et al., 2012; Michel et al., 2016) and migration stopover sites

(Rioux-Paquette et al., 2014). However, comparable declines observed across multiple, taxonomically-diverse species in the guild implicate changes in the availability and quality of flying insect prey (Nebel et al., 2010).

Riparian zones are often hotspots of aerial insectivorous bird diversity (Naiman,

Decamps, & Pollock, 1993), where condition of aquatic-riparian ecosystems can influence many aspects of aerial insectivorous bird ecology, including individual health (Smits & Fernie, 2013) and reproductive success (McCarty & Secord, 1999; Sullivan, Watzin, & Hession, 2006). In particular, aquatic insects that emerge from the water as winged adults (i.e., emergent aquatic insects) provide a critical nutritional subsidy to many riparian aerial insectivorous birds (Jackson

& Fisher, 1986; Nakano & Murakami, 2001). For instance, Kautza and Sullivan (2016b) showed that 41% of riparian swallows’ energy was derived from aquatic primary production via emergent insects. For terrestrial consumers including birds, aquatic insects may provide

54 energetic advantages over terrestrial insects; Twining et al. (2016) experimentally showed that

Tree Swallow (Tachycineta bicolor) nestlings fed a diet high in long-chain omega-3 polyunsaturated fatty acids – found in greater concentrations in aquatic insects – exhibited faster growth rates and better body condition than nestlings fed non-enriched diets.

Anthropogenic landscape changes in a watershed, such as urbanization, can have strong impacts on aquatic ecosystems (Allan, 2004; Meyer, Paul, & Taulbee, 2005; Roth, Allan, &

Erickson, 1996). Urban land cover accounted for 69.4 million acres of land (3.6% of total) in the contiguous United States as of 2018, an increase of 470% since 1945 (Merrill & Leatherby,

2018). Further, urban land use is expected to triple worldwide between 2000 and 2030, outpacing human population growth (Seto, Guneralp, & Hutyra, 2012). Urban encroachment into riparian zones is associated with marked changes in avian community composition, distribution, and reproductive success (Blair, 1996; Rodewald & Bakermans, 2006; Rodewald, Kearns, &

Shustack, 2013), commonly associated with the removal of riparian vegetation (Lussier et al.,

2006), habitat fragmentation and altered vegetation structure (Crooks, Suarez, & Bolger, 2004), and the increased presence of invasive biota (Rottenborn, 1999). The Urban Stream Syndrome

(Meyer, Paul, & Taulbee, 2005) synthesizes the collective impacts of urbanization on streams flowing through urbanized catchments including channelization of streams, increased runoff, homogenized stream geomorphology, flashier hydrographs, elevated concentrations of nutrients and contaminants, and fragmentation of in-stream habitat (Vietz, Walsh, & Fletcher, 2015;

Walsh et al., 2005). These alterations commonly lead to reduced diversity of aquatic insects

(Gage, Spivak, & Paradise, 2004; Urban et al., 2006), with potential implications for aerial insectivorous birds (Stenroth et al., 2015; Twining et al., 2016). For example, Alberts, Sullivan,

55 and Kautza (2013) found that riparian swallows breeding in urban habitats were more dependent on emergent insects and fed at higher trophic levels than rural swallows, thereby increasing their exposure to aquatic contaminants. Stenroth et al. (2015) demonstrated that agricultural development in riparian corridors was related to smaller average body size in flying insect assemblages, which in turn were related to the distribution of insectivorous birds in riparian habitats, which were associated with forested streams and large-bodied insects.

Climatic conditions may also influence avian populations and phenology, including the timing of breeding and migration, reproductive success, and the ranges of individual species

(McCarty, 2001). For instance, 55% of Ohio’s bird species are projected to be climate-threatened or climate-endangered (i.e., in danger of losing more than 50% of their current range by 2080 or

2050, respectively) if present warming trends continue (Langham et al., 2014). Climate change is predicted to exert a similar influence on the phenology and range of flying insects (Hughes,

2000), with potentially adverse effects on insectivorous birds (Visser et al., 1998). Altered temperature regimes may influence birds directly (Becker & Weisberg, 2015; Pipoly et al., 2013) or indirectly via changes to invertebrate assemblages (Jonsson et al., 2015) and insect activity

(Nooker, Dunn, & Whittingham, 2005). At local scales, the urban heat-island effect is a chief driver of localized climate variability, as urban environments typically experience elevated temperatures owing to reduced vegetation cover, diminished evapotranspiration, lower surface albedo, numerous reflective vertical surfaces, higher concentrations of suspended airborne particulate matter, and the presence of anthropogenic heat sources (Oke, 1982; US EPA, 2008).

At the regional scale, climate conditions may have an additional impact on aerial insectivores at the population level. For instance, climate change may facilitate multiple broods per season, a

56 phenomenon which has been observed with increased frequency at the southern end of the Tree

Swallow’s breeding range (Monroe et al., 2008).

Considered together, the complex relationships among climate, water quality, and urbanization demand a broad perspective to adequately address their impacts on aerial insectivorous birds. Here, we investigated the potential impact of these factors on reproductive success and individual body condition of Tree Swallows over four years at a suite of urban and protected (i.e., forested or other natural vegetation) riparian study sites in and around Columbus,

Ohio, USA. We hypothesized that water quality would mediate the effects of urbanization via changes in nutritional subsidies to Tree Swallows (i.e., the availability and composition of invertebrate prey assemblages) (Fig. 2.1). Specifically, we predicted that increased landscape urbanization would be related to impaired water quality and reduced diversity (McKinney, 2002;

Walsh et al., 2005) and quality (Roy et al., 2003) of both aquatic and terrestrial insects, resulting in lower reproductive success and body condition in Tree Swallows. We also anticipated that emergent aquatic insects would serve to transfer contaminants from aquatic ecosystems to Tree

Swallows (Walters, Fritz, & Otter, 2008), with potential implications for swallow survival and reproduction (Rowse, Rodewald, & Sullivan, 2014; Sullivan & Rodewald, 2012). In particular, we predicted that mercury (Hg) blood concentrations would be elevated in urbanized landscapes in both adult and nestling swallows (Driscoll et al., 2007). Further, we predicted that higher spring temperatures at urban sites would be related to earlier clutch initiation (Dunn & Winkler,

1999; Hussell, 2003) and that greater variability in air temperatures (i.e., greater extremes) would be negatively associated with fledging success, clutch size, and nestling mass owing to both direct effects on individual condition (e.g., increased stress) and indirect effects via prey

57 availability (Andrew et al., 2017; Ardia, Cooper, & Dhondt, 2006; Winkler, Luo, &

Rakhimberdiev, 2013).

Methods and Analyses

Study area and experimental design

The Tree Swallow was selected as the model aerial insectivore as this species readily makes use of artificial nest boxes, limiting external perturbations (e.g., predation, interspecific competition, and nest parasitism) and enabling capture of both juveniles and adults (Jones,

2003). At the outset of breeding season in late March each year (2014-2017), nest boxes were deployed at seven riparian study sites distributed across urban and natural/protected land cover in the Scioto River system (Ohio, USA) (Table 2.1, Fig. 2.2). The Scioto River watershed drains

16,868 km2 as it flows south into the Ohio River (Ohio Environmental Protection Agency, 2014).

The middle Scioto River catchment lies within the Columbus Metropolitan Area. Our study sites were located on the Scioto River mainstem (6th order), as well as on two major tributaries, the

Olentangy River (5th order) and Big Darby Creek (4th order). The middle Scioto River basin is composed of developed (45%) and cultivated (40%) lands, with forests comprising only 6% of the total land cover (Ohio Environmental Protection Agency, 2014). Each study site consisted of a 500-m river reach and the adjacent riparian zone, extending 500-m from each bank perpendicular to the river channel. This spatial extent was selected because (1) Tree Swallows generally forage within 500 m of their nest boxes (Dunn & Hannon, 2016; Quinney & Ankney,

1985) and (2) 500 m is considered sufficient to capture the influence of land use and land cover on stream water chemistry at the local scale (Strayer et al., 2003). Nest boxes were constructed

58 according to Golondrinas de las Américas (2011), with identical interior dimensions (12.7-cm ×

12.7-cm × 27.9-cm) to control for possible differences in clutch size due to cavity size (Rendell

& Robertson, 1993). Nest boxes were mounted on a combination of steel rebar and electrical conduit to deter predators; predator guards constructed from PVC pipe were installed at sites where predation was observed or suspected although such instances were very rare. Five to six nest boxes were deployed at each study site, spaced at moderate distance from one another (~20 m) to avoid territorial overlap (Muldal, Gibbs, & Robertson, 1985). All nest boxes were monitored at 2-3 day intervals from shortly before to the end of the breeding season each year.

Nestlings were monitored and accessed via the swinging side-door of the nestbox, while adult birds were captured using the nestbox’s ‘wig-wag’ (i.e., trapdoor at the entrance hole) which can be closed from a distance with a length of monofilament fishing line, thus trapping the adult bird inside and facilitating capture via the side-door (Golondrinas de las Américas, 2011). All captured birds were banded in accordance with North American Bird Banding Program’s banding protocols (U.S. Geological Survey, 2018).

Land use and land cover

Using Quantum GIS (QGIS Development Team, 2017), the percentage of forested or wetland land cover and developed land cover with impervious surfaces at each stream reach were quantified. Land-cover data were obtained using the 2011 National Land Cover Database, which classifies land cover for the continental United States at a 30-m spatial resolution (Homer et al.,

2015). A 500-m buffer was delineated on each side of the stream channel as described above.

Land-cover percentage for each land-cover class (forest, developed, et al) was then calculated for

59 each delineated buffer. Additional GIS layers were added to calculate the percent impervious surface (U.S. Geological Survey, 2014a), percentage canopy cover (U.S. Geological Survey,

2014b), and mean human population density (U.S. Census Bureau, 2010).

Tree Swallow reproductive success and body condition

Nestboxes were observed 2-3 times per week to determine if and when breeding pairs of

Tree Swallows established nests. Tree Swallow reproductive success was measured using several metrics: clutch initiation date (Julian date, no. calendar days), clutch size (no. eggs), no. of eggs successfully hatched, no. of nestlings successfully fledged, and mean nestling weight (mg). The latter measure, nestling mass, is considered an important metric of reproductive success – songbird fledgling mass has been strongly linked to post-fledging avoidance of predation (Naef-

Daenzer & Grüebler, 2016), so the diminished condition of smaller swallows may impair their odds of post-fledging survival. Tree Swallow females lay one egg per day (Hussell & Quinney,

1987; Nooker, Dunn, & Whittingham, 2005); clutch initiation date was determined based on this rate (e.g., observing a nest with two eggs indicated that laying began the day prior). To measure growth rates, each brood was weighed at four, seven, and ten days using an OHAUS Scout Pro

SP601 portable balance (Parsippany, New Jersey, USA). Any disturbance of nesting or laying, such as predation or destruction of eggs by competing cavity-nesters (typically the House

Sparrow [Passer domesticus] or the House Wren [Troglodytes aedon]) was recorded, though such events were very infrequent. Furthermore, nesting material deposited by House Sparrows or

House Wrens was removed in an effort to discourage competition. Tree Swallows occasionally produce second clutches if their first broods fledge sufficiently early (Monroe et al., 2008).

However, second clutches as well as replacement clutches (i.e., those laid when the first clutch is destroyed or fails to hatch) are often smaller than the first (Karagicheva et al., 2016), as are late- season clutches of any sort (Winkler & Allen, 1996) so these clutches were identified and not considered in our analyses. Clutches that were destroyed or reduced due to factors outside of the scope of this project (e.g., predation, interspecific competition, flooding) were also identified and such clutches were excluded from subsequent analyses. Similarly, if such factors were implicated in nestling deaths, the affected broods were excluded from subsequent analyses.

Morphological and individual health measurements were performed on all captured swallows. We weighed all adults and measured tarsus length (mm) using handheld calipers.

These measurements, in conjunction with body mass measurements, may be used to estimate individual body fat and general condition (Labocha & Hayes, 2012). Scaled mass index (SMI) was used as an estimate of body fat percentage in passerine birds and served as the primary metric of individual body condition for adult birds. SMI was calculated using the method developed by Peig and Green (2017) as follows:

푏푆푀퐴 퐿0 푀̂푖 = 푀푖 [ ] , 퐿푖 where Mi is the body mass of the individual bird, and Li is the appropriate linear body measurement (in the case of passerine birds, tarsus length). L0 is the mean linear body measurement of the study sample, and bSMA is the scaling exponent determined by dividing the slope from an ordinary least squares regression by the Pearson’s correlation coefficient. Due to sex differences in morphology and mass, male and female swallows were modeled separately in subsequent analysis. Nestling birds were weighed at ~13 days, but morphological measurements were not taken; consequently, SMI was not calculated for nestlings. 61

Blood samples were drawn from the jugular vein of adults and nestlings (e.g., Sullivan &

Vierling, 2012) on day 13 after hatching. Concentrations of Hg were estimated from these blood samples by applying a small amount of blood was applied to a dried blood spot (DBS) card, considered a suitable method for measuring contaminant concentrations in wild birds (Shlosberg et al., 2011). DBS cards were sent to Michigan State University’s Diagnostic Center for

Population and Animal Health (Lansing, MI, USA) to determine concentrations of Hg, measured in parts per billion (ppb). Blood plasma glucose was also estimated from blood samples, as blood glucose levels are a commonly used indicator of avian body condition. For instance, hyperglycemia or hypoglycemia in birds may indicate septicemia or stress, respectively (Harris,

2009; Lill, 2011). A droplet of blood was collected on a plastic cuvette and placed inside a

HemoCue Glucose 201 Analyzer (Brea, CA, USA) glucose meter for analysis. Since we could safely draw only a limited quantity of blood from any individual bird, 2-3 replicates of each of the above measurements were made per brood.

Climate, chemical water quality, and nutrients

Climate variables included temperature and humidity measurements, collected using deployable ThermocronTM and HygrochronTM Ibutton (Baulkham Hills, NSW, AU) passive temperature and humidity sensors, respectively, installed inside the nest boxes. One temperature sensor was installed inside each nest box at each study site. In addition, one humidity sensor was installed per site. All sensors were set to record temperature and/or humidity at 6-hour intervals to obtain a daily average. These data were downloaded from the sensors at close of each field season.

A suite of chemical water-quality variables was measured at each study site twice per year (mid-late May and mid-late July). Water samples were collected at the thalweg and both stream edges at upstream, middle, and downstream sections of each reach, for a total of nine

250-mg water samples per site. These samples were then sent to The Ohio State University’s

Service Testing and Research Laboratory (Wooster, OH, USA) for analysis of total phosphorus

(mg L-1), total nitrogen (mg L-1), phosphate (mg L-1), nitrate (mg L-1), ammonia (mg L-1), total dissolved solids (mg L-1), and mercury (Hg; ppt). At the same nine locations per site, water temperature, pH, conductivity, and dissolved oxygen content were also measured using a handheld Hach sensION+ Portable Meter (Loveland, CO, USA). In addition, water samples were collected in 60-mg plastic bottles in the stream thalweg at the upstream, middle, and downstream of each reach (three samples total per reach) and analyzed in a Hach 2100N Turbidimeter

(Loveland, CO, USA) to measure the Nephelometric Turbidity Units (NTUs) of each sample. All water-chemistry data was collected in the early season (mid-late May) and again in the late season (mid-late July), yearly.

For temperature and humidity data, means were calculated for each nest box. Since temperature and humidity varied by nest box within each study site (e.g., owing to the presence of shading vegetation near some boxes), a reach-wide temperature mean was also calculated for each reach and year. Separate means were calculated for 30-d periods: one for the earlier part of the breeding season from 30 April to 29 May (inclusive), and one for the latter part of the season,

30 May to 29 June (inclusive). Most clutch initiations were expected to fall within the designated early 30-d period from late April through May, while most nestling growth and fledging was expected to occur within the 30-d period from late May through June.

Extreme temperatures were evaluated by counting the number of days during the May-

June breeding season in which the temperature passed a designated limit, following methodology developed by Pipoly et al. (2013). Extreme heat days were those in which the highest recorded daily temperature exceeded the 90th percentile of all high temperatures measured for the time period (May-June) in all next boxes across all 4 years of data. Similarly, extreme cold days were those in which the lowest recorded daily temperature fell below the 10th percentile of all low temperatures in the same timeframe.

Aquatic and terrestrial insects

Tree Swallow food resources, in the form of flying insect prey, were sampled for 10 d in mid-late May and again for 10 d in mid-late July at each of the study sites, aligning with climatic and water-chemistry measurements. For emergent aquatic insects, two floating, 1-m2 pyramid- style emergent traps (Kautza & Sullivan, 2015) were deployed at each study site (upstream and downstream sections) for two 10-d periods. Likewise, two cloth mesh 1-m × 1-m × 0.6-m

Malaise traps (MegaView Science Co., Taichung, Taiwan; Townes, 1972) were deployed in nearshore vegetation at upstream and downstream locations at each site, suspended from trees at a height of 1-m for two 10-d periods per breeding season. Any insects from families of aquatic origin (e.g., Chironomidae) were excluded so that only terrestrial families of flying insects were considered from Malaise traps. All invertebrates collected from boluses and traps were stored in

70% ethanol solution before being enumerated and identified to family using Johnson and

Triplehorn (2005) and Merritt, Cummins, and Berg (2008). Insects were dried in a 60°C oven and weighed by family, reach, and collection period. Insect capture rate (i.e., no. of insects m-2

10 d-1), family richness, and average body size (g, dry mass) were calculated. Due to the presence of a few very large insects in some of the samples (in some instances, 2-3 orders of magnitude larger than the typical insect), median insect body size (i.e., dry mass) was used for subsequent analysis.

Statistical analyses

All statistical analyses were performed using the R statistical package version 3.4.4 (R

Core Team, 2018) and the R Studio package version 1.1.453 (RStudio Team, 2016). Reach-wide means were calculated for each water-quality/nutrient parameter for each year and collection period (e.g., May or July). Principal component analysis (PCA) was then run on the water- chemistry/nutrient data and LULC data. We retained axes with eigenvalues >1 (Rencher, 1995), from which we selected PC1 to represent an Urban Stream Index (USI) similar to Alberts et al.

(2013) (see Results for details), which was used as a predictor variable in subsequent linear regressions (see below).

We used linear-mixed effects models (LMMs) as our main statistical tool in order to test for the potential effects of urbanization on Tree Swallow reproductive responses and measures of individual body condition. The following variables were modeled: clutch initiation date, clutch size, no. successfully fledged, nestling mass, male SMI, female SMI, adult Hg blood concentration, nestling Hg blood concentration, and nestling blood glucose. For individual body condition models, it was necessary to log10-transform the response variables to meet assumptions of normality. Year, urbanization (i.e., a categorical variable indicating the site’s land use: urban or natural/protected) and an urbanization × year interaction term were included as fixed effects,

65 with study site included as a random effect. Nestbox nested within study site was also included a random effect in the models in which nestling responses were included. The model for blood glucose did not include the year or urbanization-year interaction as fixed effects, since there was only one year (2017) of data available. Hg blood concentration data were available for the years

2014-2016 only. LMMs were developed using lmerTest version 3.0 (Kuznetsova et al., 2017) in

R. LmerTest was used to evaluate random and fixed effects from the models, including the coefficients and p-values for fixed effects. The overall effect of each variable (i.e., the sum-to- zero contrast) was also assessed via p-values and F-statistics calculated with the afex package version 0.22-1 (Singmann et al., 2018), which employs the Kenward-Roger approximation for degrees-of-freedom values. R2 values, both marginal (fixed effects) and conditional (fixed + random effects), were determined for each model with the MuMIn package (Bartoń, 2018), which calculates a pseudo-R2 for mixed-effects models. In addition, we explored potential mechanisms (i.e., USI as well as individual measures of water chemistry, nutrient concentrations, air temperature, flying insect prey) driving Tree Swallow reproductive success and body condition with post-hoc multiple regression. Multiple regression models incorporated year as a categorical variable, as many of our predictor variables (particularly the measures of local climate) had considerable interannual variability.

An alpha level of α ≤ 0.05 was used as the threshold of statistical significance, while α ≤

0.10 was employed as evidence of a trend. Note that the afex package reports p-values to only two decimal places, except in cases where p < 0.01.

Results

Physicochemical and climatic variability between urban and natural/protected sites

We observed variability in water-chemistry and nutrient concentrations between urban and natural/protected sites (Table 2.2). Concentrations of many nutrients and contaminants were higher, on average, at the urban sites. Hg concentrations, for example, were ~1.8x greater at urban sites than at natural/protected sites. Mean turbidity of the water column at urban sites was

~30% lower than their natural/protected counterparts, while water temperature at urban sites was

0.9 °C higher than at natural/protected sites.

Principal component analysis was performed using three land-use/land-cover variables and nine water-chemistry and nutrient variables (Table 2.3). PC1 – with an eigenvalue of 7.48 and accounting for 62.3% of the variance among the selected variables – served as our USI. This axis was primarily influenced by human population density, total nitrogen, Hg, canopy cover, nitrate, impervious surface, dissolved oxygen, and water temperature. The results of the USI closely correspond with the urban/natural dichotomy established by classifying sites based on impervious surface coverage (Table 2.1), with some variability given the inclusion of other urban characteristics (e.g., “Mussel” USI was driven mostly by human population density and water- quality chemistry conditions).

We observed sizeable and consistent differences in air temperature between urban and natural/protected reaches (Fig. 2.3). On average, urban reaches were warmer than protected reaches through the breeding season from 30 Apr – 28 Jun. This disparity was similar in both the early period from 30 Apr – 29 May (urban: 18.9 °C; protected: 17.1 °C) (Fig. 2.3a) and the late period from 30 May – 28 June. (urban: 24.0 °C; protected: 22.5 °C) (Fig. 2.3b). Moreover,

67 urban reaches experienced fewer days of extreme cold (Fig. 2.3c) and more days of extreme heat

(Fig. 2.3d). Urban sites experienced an average of five days of extreme cold and 11 days of extreme heat between 30 Apr – 28 June, compared to nine days of extreme cold and seven days of extreme heat at natural/protected sites.

Invertebrate prey assemblages between urban and natural/protected reaches

Our measures of flying insect family richness, abundance, and average body size revealed variability between the urban and natural/protected sites and across the four years of the study

(Fig. 2.4). For emergent aquatic insects, there was little distinction in family richness (Fig. 2.4a) or abundance (Fig. 2.4b; estimated via rate-of-capture) between urban and natural/protected sites.

However, median emergent aquatic insect body size was over 4x larger at urban reaches over the course of the study (Fig. 2.4c). For terrestrial flying insects, the opposite held true: median insect body size at the natural/protected sites was over twice that of urban reaches (Fig. 2.4c). The natural/protected sites also exhibited higher terrestrial insect family richness and abundance.

Tree Swallow reproductive success

We used linear mixed models (LMMs) to evaluate differences in reproductive response variables by land use and over time. We observed little variability in clutch size, either across years (LMM: p = 0.440; Table 2.4, Fig. 2.5a) or between urban and protected sites (LMM: p =

0.710), with some evidence of an interaction effect for year × land use (LMM: p = 0.090); nests at urban sites produced about one less egg in 2015 than in the previous year (LMM: p = 0.045).

There was also a trend toward earlier clutch initiation at urban sites (LMM: p = 0.060; Table 2.4,

Fig. 2.5b). Land use was a significant factor for number of successfully fledged young, which was significantly higher at the urban nesting sites than at protected sites (LMM: p = 0.009; Table

2.4, Fig. 2.5c), with urban nests yielding, on average, ~1.5 > fledglings each year than their protected counterparts. Nestling mass at ~13 d was significantly different by year (LMM: p =

0.006; Table 2.4, Fig. 2.5d); average nestling mass ranged from 22.5 g in 2014 to 19.2 g in 2015

(p = 0.001), and from 21.9 g in 2016 (p = 0.594) to 20.4 g in 2017 (p = 0.037). In addition, urban nests produced the highest average nestling mass in 2015 (LMM: p = 0.038; Fig. 2.5d). Site, included as a random effect in our models, emerged as an influential factor for nestling mass only, where the inclusion of random effects increased the R2 from 0.13 to 0.35.

We investigated, via regression models, some possible mechanisms driving variability in reproductive responses. For no. fledged, the USI was positively related to fledging success (p <

0.001; Fig. 2.6a). The no. of days of extreme cold, on the other hand, was negatively associated with fledgling success (p = 0.003; Fig. 2.6b). Similarly, nestling mass was negatively related to the no. of extreme heat days during the breeding season (p = 0.003; Fig. 2.7a). Clutch initiation date was also related to temperature, with earlier laying associated with higher average temperatures in the early part of the season (p = 0.039; Fig. 2.7b).

Tree Swallow individual body condition

For several body condition indices, variability was fairly limited. Neither male (LMM: p

= 0.650) nor female (LMM: p = 0.680) SMI varied by land use alone (Table 2.4; Fig. 2.8a).

However, male swallows exhibited significant annual variability (LMM: p = 0.010; Table 2.4), with the highest SMI in 2016 (LMM: p = 0.027; but note that no adult male Tree Swallows were

69 captured in 2015). There was no significant difference in Hg blood concentrations between swallows at urban and protected sites for adults (LMM: p = 0.440) or nestlings (LMM: p =

0.920; Table 2.4, Fig. 2.8b), though the latter had an interaction effect with year (LMM: p <

0.001; Table 2.4), which was related to lower Hg among nestlings at urban sites in 2015 and

2016. Nestlings with the lowest average blood concentrations of Hg were observed in 2016 (p <

0.001), a year that coincided with very low observed Hg concentrations in the water column.

Both adults and nestling Hg concentrations were heavily influenced by site and/or site × nestbox, suggesting that local-scale variability among sites was important. For instance, the USI emerged as a strong predictor of elevated Hg (p = 0.001; Fig. 2.9). For 2017, nestling blood glucose showed no significant differences between site types (LMM: p = 0.31; Table 2.4, Fig. 2.8c), although random effects (site, site × nestbox) were very influential (R2 marginal = 0.05, R2 conditional = 0.41; Table 2.4).

Discussion

Urbanization can have a dramatic impact on avian communities and ecosystem function

(Schlesinger, Manley, & Holyoak, 2008). For instance, meta-analysis by Chamberlain et al.

(2009) of ten species of urban-breeding North American and European passerine birds, though revealing some variability among species, showed a general trend toward smaller clutches, fewer successfully fledged young, earlier clutch initiation dates, and lower nestling mass in among urban breeders. In the Columbus-area study system of our current study, Rodewald et al. (2013) observed that the aerial insectivorous Acadian Flycatcher (Empidonax virescens) exhibited diminished reproductive output from 2001-2011. Given the reliance of riparian swallows and

70 other aerial insectivores on aquatic invertebrate subsidies both in our study system and elsewhere

(Kautza & Sullivan, 2016b; Nakano & Murakami, 2001), we expected that losses in water quality and altered temperature regimes in urban areas would be reflected in urban-breeding Tree

Swallows, both through impaired reproductive success and compromised body condition in both nestlings and adults.

Our results challenged some of our expectations. We observed limited differences in nestling mass or clutch size between urban Tree Swallows and their counterparts nesting in natural areas. There was also little variability in clutch size by land-use, although nestling mass exhibited strong interannual differences: we observed significantly lower mean nestling mass over all study sites in 2015, a result we also observed across the study system in 2017. Of all reproductive responses, nestling mass also had the most variability among sites, suggesting local, site-specific effects such as microclimatic differences (e.g. Fig. 2.7a) may have exerted a greater influence on nestling mass than urbanization alone. Extreme heat, for example, which was more pronounced at our urban sites (Fig. 2.3d), has been related thermal stress, inhibited growth, and even increased mortality in nestling passerine birds of various species (e.g., Andrew et al., 2017;

Cunningham et al., 2013; Rodríguez & Barba, 2016). Water quality, via its influences emergent insects (e.g., body size, Hg contamination) was also implicated as an urban-related factor linked to swallow responses. For example, urbanization in the reach (estimated via the USI) was strongly related blood Hg concentrations in adult Tree Swallows (Fig. 2.9).

In a notable divergence from our hypothesis, the number of successfully fledged young was higher for urban broods than at the natural/protected sites. Although there was a trend toward interannual variability in fledging rates, the effect of urban land use had, by far, the

71 strongest relationship (Table 2.4, Fig. 2.5c). As clutch size did not vary substantially with land use (Table 2.4, Fig. 2.5a), lower nest productivity at natural/protected sites must be attributed to higher mortality of eggs and/or nestlings. Since nests experiencing losses from predation, interspecific competition, floods, and similar events were excluded from our analyses, this mortality is due to starvation, exposure, disease, or other indeterminate causes, some of which can be related to the environmental conditions measured, including local climate and food availability. Fledging success, more than other factors, may play a leading role in driving population growth. Population models developed by Cox et al. (2018), based on 42-year study of a Tree Swallow population in Ontario, Canada, indicated that fledging success and overwinter survival were the two most influential drivers of population change, while clutch size was of limited importance.

Urbanization, as described by the USI, was related to increased fledging success (Fig.

2.6a). Variability in the quantity and quality (Fig. 2.4) of insect prey among sites and land-use types was likely a factor related to the influence of land use on Tree Swallow fledging success.

Previous studies of insectivorous birds including for Barn Swallows (Teglhøj, 2017) and House

Wrens (Newhouse, Marra, & Johnson, 2008) have reported lower nestling mass among birds breeding in developed areas, results which are thought to be related to reduced prey quality or lower prey abundance. However, we found that the median body size of emergent insects was greater at our urban sites relative to the natural/protected sites (Fig. 2.4c). Although terrestrial insect body size was greater at the natural/protected sites, experimental work by Twining et al.

(2016) suggests that the benefit of aquatic insects, which are relatively rich in omega-3 long- chain polyunsaturated fatty acids, may be more important than terrestrial insect prey availability.

Other evidence indicates that Tree Swallows selectively catch larger-bodied prey when provisioning nestlings (McCarty & Winkler, 1999a; Quinney & Ankney, 1985), underscoring the importance of prey body size as a determinant of fledging success and post-fledging survival.

Among our four reproductive response metrics, only responses of clutch initiation date conformed to our hypotheses, with urban birds laying eggs significantly earlier than their natural/protected counterparts (Fig. 2.5b). Evidence implicates climate as the likely driver behind the observed differences in clutch initiation. There were considerable differences in both air temperatures means (Fig. 2.3a) and frequency of extreme temperatures (Fig. 2.3c, d) between urban sites and protected sites, implicating the urban heat island effect as the likeliest driver of air temperature variability among sites (Oke, 1982). The early temperature period (30 April – 29

May) captured most of the clutch initiations (91.5%). In our investigation of potential drivers of reproductive response, air temperature during this period emerged at the strongest predictor of clutch initiation date, a finding that was consistent across all years of data (Fig. 2.7b). Urban sites were both markedly warmer and associated with significantly earlier clutch initiation (Fig. 2.5b), strongly suggesting local-scale climate conditions are the mechanism driving these differences.

The availability of insect prey has been identified as a breeding cue in Tree Swallows (Nooker,

Dunn, & Whittingham, 2005), suggesting that temperature-driven insect activity and/or emergence as the link between temperature and clutch initiation (Eeva, Veistola, & Lehikoinen,

2000). Our results echo findings by Hussell (2003), who observed strong and consistent correlations between laying date and local-scale air temperature in early May among Tree

Swallows from 1969-2001. More recently, Bourret et al. (2015) not only revealed similar correlations between regional-scale temperatures and Tree Swallow clutch initiation, but strong

73 evidence that the rise in interannual spring temperature was associated with advances in Tree

Swallow clutch initiation dates from 2004-2013.

Although our results did not show consistent increases in spring air temperature year-to- year 2014-2017, the strong relationship between clutch initiation and early spring temperatures suggest that elevated temperatures may be linked to advanced laying dates among Tree Swallows across our study system. These results are consistent with the analysis of Tree Swallow breeding data from 1959-1991 performed by Dunn and Winkler (1999), which showed advances in laying dates across the breeding range linked to climate change, including a more pronounced rate of change at the southern edge of the range. Air temperature was identified as potentially influential driver of the number fledged. The frequency of extreme cold days was related to fewer fledged young (Fig. 2.6b). Again, the role of land use was prominent – natural/protected sites were associated with more days of extreme cold (Fig. 2.3c). Cold snaps are strongly associated with elevated nestling mortality, as they suppress flying insect activity while increasing Tree Swallow energy demands (Winkler, Luo, & Rakhimberdiev, 2013). Higher spring temperatures have been identified as potentially beneficial for reproductive success among some species of breeding birds (Becker & Weisberg, 2015). The warmer conditions associated with urban heat islands – that is, overall higher temperatures, fewer days of extreme cold, and more stable overnight temperatures (Tam, Gough, & Mohsin, 2015) – may both mitigate the impact of periods of cold weather, while intensifying flying insect activity early in the morning relative to cooler non- urban areas. Tree Swallow nests observed by Ardia (2013) yielded results in line with these expectations, as higher nighttime temperatures in nestboxes were related to higher fledge rates.

The benefit of increased nest temperature on Tree Swallow nestling mass was also evidenced in

74 experimental treatments by Perez et al. (2008). As noted above, however, higher temperature regimes may have negative implications for aerial insectivore body condition and reproductive success. For instance Ardia (2013) found that extreme heat (> 35°C) in Tree Swallow nests was related to both impaired body condition (including reduced mass) and reduced fledging success.

As Ohio lies near the southern end of the Tree Swallow’s breeding range (Robertson, Stutchbury,

& Cohen, 2011), it is likely that extreme heat events will have a greater impact on Ohio Tree

Swallow populations. However, higher spring temperatures may provide an additional benefit to

Tree Swallows by extending the breeding season, a phenomenon which has been linked to greater fecundity in some other passerine species (e.g. Townsend et al., 2013). Monroe et al.

(2008) observed regular double-brooding among a population of Tree Swallows in Virginia, a practice which dramatically increases swallow reproductive output and may have a positive influence on population growth. High-resolution data that capture the effects of temperature across both local and broader spatial scales, including potential impacts on nestling and egg mortality (Ardia, 2013) or nestling growth (Pipoly et al., 2013), will be critical in understanding impacts of both urbanization and climate change on aerial insectivores.

We did not observe our significant variability by land use among individual body condition metrics. Regarding blood concentrations of Hg, variability among sites was far more influential than land use for both adults and nestlings, highlighting the potential importance of local-scale variability (Table 2.4, Fig. 2.8b). We observed similar results among blood glucose levels in nestlings (Fig. 2.8c). In addition, we observed extreme interannual variability among nestlings for blood Hg concentrations (Table 2.4, Fig. 2.8b). In another finding inconsistent with our expectations, urban nests in 2015 and 2016 were associated with reduced blood Hg.

Although land use by itself was not identified as a strong predictor of Hg concentrations, the results of our mixed-effects model (Table 2.4) indicate substantial variability among study sites beyond the urban/protected land-use dichotomy. For example, the USI, which reflects a suite of continuous variable of urbanization, strongly predicted Hg blood concentrations in adult swallows, supporting the premise that emergent insects serve as vectors for aquatic contaminants to terrestrial consumers (Walters, Fritz, & Otter, 2008; Sullivan & Rodewald, 2012). The elevated concentrations found in the adult swallows (t = 2.814, p = 0.010) align with previous findings in the same region by Alberts, Sullivan and Kautza (2013), who speculated that higher concentrations among adult swallows may be a due to nestlings’ rapid growth or a consequence of prey selection; the latter would also account for the strong positive relationship between urbanization and Hg in adult Tree Swallows, compared to the negative relationship we observed in nestlings in 2015 and 2016. Further, Tree Swallows tend to forage over open fields or water

(Ghilain & Bélisle, 2008; McCarty & Winkler, 1999a), therefore stream channel width or the distribution of riparian vegetation may cause swallows to prey more or less on aquatic insects, thereby increasing their exposure to Hg (Alberts, Sullivan, & Kautza, 2013). Hg concentrations in both adult and nestling swallows also appeared to track interannual variation in Hg concentrations measured in the water column – for instance, adult swallows across the study system and urban nestlings exhibited significantly lower Hg in 2016 (Table 2.4), a year with the lowest observed Hg concentrations in the river reaches – supporting a link for the transfer of contaminants between the aquatic and terrestrial systems.

Hg has an array of well-documented harmful health impacts on birds (Boening, 2000); among Tree Swallows, weakened immune response has been identified as a major adverse effect

76 of Hg exposure, with potential long-term consequences for fitness (Hawley, Hallinger, & Cristol,

2009). Further, Hg has been linked to impaired reproductive success in Tree Swallows (Brasso

& Cristol, 2008). In fact, the adverse effects of Hg on reproductive response may manifest even when there are the effect on individual health is undetected: in the same Columbus-area study system, Rowse, Rodewald, and Sullivan (2014) reported that Hg concentrations in aerial insectivorous Acadian Flycatchers were related to fewer fledged young, even though there was no observed impact on body condition.

Conclusions

Urban land use has transformed landscapes across North America, with far-reaching and persistent ecological consequences. However, our results indicate that these changes may not prove uniformly harmful to aerial insectivorous birds. On the contrary, some characteristics of the urban environment may advantage aerial insectivores in terms of reproductive success. For instance, in our Columbus-area study system, we observed greater fledging success at urban nests, which may be linked to both greater body size of energetically profitable emergent insects and to microclimatic differences. Temperature regimes in urban areas may mitigate the impact of early spring cold snaps, affording a longer breeding season, and spurring insect activity. If climate change prompts phenological changes in the timing of spring migration as predicted

(Crick, 2004), the risks associated with early nesting could be mitigated by the availability of warmer urban breeding habitats. However, climate shifts may also have unfavorable consequences. Global climate change is expected increase frequency of extreme heat events

(Gerald & Tebaldi, 2004), the impacts of which will be intensified by the urban heat island effect

(Stone, Hess, & Frumkin, 2010). Cities like Columbus – home to the 8th-most intense summer urban heat island among the 60 largest U.S. cities (Kenward et al., 2014) – may experience the most intense impact. Nestling mass been identified as a strong indicator of post-fledging survival, itself a key predictor of population growth (Cox et al., 2018). Post-fledging survival was not addressed by our research, but it represents a key line of investigation for understanding of aerial insectivore population declines.

Other hazards associated with urbanization are also evident. Among these, perhaps the most notable in our results are the differences in Hg body burdens. The relationship between Hg in the aquatic environment and bioaccumulation of Hg in Tree Swallows indicates not only a danger from Hg toxicity, but also a vulnerability to the transfer of other contaminants from aquatic systems, such as selenium (Ohlendorf et al., 1988) and PCBs.

It is important to note that Tree Swallows may be unique in their responses to urbanization, and other aerial insectivorous species might fare differently. Foraging strategy, nesting strategy, and dietary preferences may all figure in individual species’ responses to urbanization. For instance, altered vegetation structure and composition in urban habitats may be detrimental to the foraging efforts of salliers like flycatchers (Schneider & Miller, 2014). Cavity- nesters like Tree Swallows may benefit from the proliferation of artificial nestboxes (Norris et al., 2018) and the open-water environments of impounded urban rivers, while nesting sites for otherwise urban-tolerant species, like Chimney Swifts (Environment Canada, 2007) and

Common Nighthawks (Brigham, 1989) have declined due to changes in building construction.

Overall, the joint effects of urbanization, water quality, and climate change on Tree Swallows