How Ant Communities Are Shaped by Vacant Land Management Strategies, Landscape Context

How ant communities are shaped by vacant land management strategies, landscape context,

and a legacy of industrialization

Thesis

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

Graduate School of The Ohio State University

Alexander Marcus Tyrpak

Graduate Program in Entomology

The Ohio State University

2020

Thesis Committee

Dr. Mary Gardiner, Advisor

Dr. Rachelle Adams

Dr. Larry Phelan

Dr. Joseph Raczkowski

Copyrighted by

Alexander Marcus Tyrpak

2020

Abstract

Many cities around the world have lost population due to economic decline and deindustrialization which has led to an overabundance of infrastructure. These abandoned structures are eventually torn down, resulting in the formation of vacant lots which are planted with turfgrass. Vacant lots require regular mowing, which is a significant expense for cities with a shrinking tax base. Vacant land has the potential to support many arthropod groups yet mowing could be limiting its conservation value. The first goal of this research was to determine if reducing the mowing frequency could improve the conservation value of vacant land by creating urban meadows. Second, I examined if converting vacant lots into pocket prairies planted with native wildflowers would improve their conservation value, in comparison to

Metropark forests which are the principal conservation habitat example in Cleveland, Ohio.

Ants were chosen as the study organism due to their abundance and importance as biological indicators. Landscape context was evaluated in both goals, and soil contamination and vegetation factors were evaluated between vacant lots and pocket prairies. Reducing the mowing frequency and establishing pocket prairies did not significantly alter ant species and functional richness, or composition, and Metropark forests had a lower ant species and function richness and a different species composition than pocket prairies. However, ant species richness was negatively correlated with increasing impervious surface area, and local soil contamination was negatively correlated with ant body size. These results illustrate that local soil contamination and impervious surface area in the surrounding landscape can potentially negatively affect ant colonization and functional diversity. Urban meadows, pocket

ii prairies, and vacant lots all harbored relatively similar communities of ants, however, soil contamination and impervious surfaces in the surrounding landscape could be detrimental for vacant land conservation practices.

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Acknowledgements

Dr. Kayla Perry who mentored me and acted as a co-advisor and helped me along the way with learning about ants, field work, statistics, and writing up my thesis. My advisor Dr. Mary

Gardiner for always providing encouragement and guidance throughout my masters and giving me the opportunity to study one of my favorite urban insects, ants. Dr. Joe Raczkowski for helping me to identify the difficult ants and for his expertise on ant biology. Dr. Larry Phelan for asking all the tough questions and allowing me to use the soil data collected by the Phelan Lab in my research. Dr. Rachelle Adams for providing a lot of advice and feedback on my project, giving out guidance on ant biology and behavior, especially pertaining to my collection methods and interpretation of data. Emily Sypolt and Denisha Parker for their support and guidance throughout my degree, especially with field work. Chris Riley for helping with statistics. Yvan

Delgado de la flor for helping with statistics and for providing the ants that he collected in 2015 and 2016. The Cleveland Land Bank for allowing us to lease and do research on the parcels of vacant land used in this study, as well as the Cleveland Metroparks for allowing us to sample at several of their parks. Ellen Dunkle, Anthony Ursetti, Michael Rogers, Caleb Whitney, Jena

Copley, Sierra Weir, Amanda Han, Kelly Luebbering, Ava Wilson, Meagan Carey, Emily Goodwin,

Mary Roth, and Molly Frabotta for their assistance with fieldwork and laboratory tasks related to my research. This research was supported in part by the Division of Environmental Biology

CAREER Grant (CAREER-1253197) to Mary Gardiner, the USDA AFRI Agroecosystem

Management Grant (20166701925146), and the USDA Agricultural Research Program Initiative

Foundational Programs Grant (2017-67013-26595).

Vita

2010………………………………………………………………….Brunswick High School

2014-2017…………………………………………………………Undergraduate research assistant in Susan

Jones’ lab

2014………………………………………………………..………..Ground beetle ID with Sarah Bowman PhD

candidate

2016…………………………………………………………………..A.S. Environment and Natural Resources, The

Agricultural Technical Institute of The Ohio

State University

2016……………………………………………………………………B.S. Entomology, The Ohio State University

2017-present……………………………………………………..M.S. Entomology, The Ohio State University

2017…………………………………………………………………..Ohio Pest Management Association

scholarship

2018…………………………………………………………………..1st Place Delong 10 minute presentation

competition

2018……………………………………………………………………2nd place PIE student competition 10 minute

presentation competition

2019……………………………………………………………………NCB travel scholarship

2019……………………………………………………………………1st place PIE student competition 10 minute

presentation

2019…………………………………………………………………….ROOT award recipient 10 minute

presentation

2020…………………………………………………………….………NCB virtual 1st place PIE student

competition 10 minute presentation

Fields of study

Major Field: Entomology

Environmental and Natural Resources

Table of Contents

Abstract……………………………………………………………………………………………………………………………………..ii Acknowledgements…………………………………………………………………………………………………………………..iv Vita…………………………………………………………………………………………………………………………………………….v List of Tables……………………………………………………………………………………………………………………..………ix List of Figures……………………………………………………………………………………………………………………………..x Chapter 1. Mowing frequency does not influence ant communities living in urban vacant lots…………………………………………………………………………………………………………………………………….………1 Abstract…………………………………………………………………………………………………………………………..……..1 Introduction……………………………………………………………………………………………………………………………2 Methods…………………………………………………………………………………………………………………………………5 Site Selection……………………………………………………………………………………………………………….……..5 Ant collection and identification…………………………………………………………………………………………8 Vegetation data………………………………………………………………………………………………………………….9 Landscape data…………………………………………………………………………………………………………………10 Data Analysis……………………………………………………………………………………………………………….……11 Results………………………………………………………………………………………………………………………………….12 Discussion………………………………………………………………………………………………………………………..…..19 Conclusions……………………………………………………………………………………………………………………..…..22 Chapter 2. Using ant bioindicators to assess how vacant land conservation compares to existing suburban Metropark preserves……………………………..…………………………………………………………….….24 Abstract…………………………………………………………………………………………………………………..…………..24 Introduction…………………………………………………………………………………………………………………..…….25 Methods………………………………………………………………………………………………………………………….……29 Site Selection……………………………………………………………………………………………………………….……29 Ant collection and identification…………………………………………………………………………………….…32 Soil data……………………………………………………………………………………………………………………………36

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Vegetation data………………………………………………………………………………………………………………..38 Landscape data…………………………………………………………………………………………………………………39 Data Analysis…………………………………………………………………………………………………………………….40 Results………………………………………………………………………………………………………………………………….42 Discussion…………………………………………………………………………………………………………………………….54 Conclusions………………………………………………………………………………………………………………………….57 References……………………………………………………………………………………………………………………………….59

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List of Tables

Table 1: Ants collected from unbaited pitfall traps in control and urban meadow vacant lots in 2015-2016 in Cleveland, Ohio, USA. …………………………..……………………………………………………………13

Table 2: Ant community weighted means for individual f Ant community weighted means for individual functional traits, functional richness, and functional dispersion of dietary niches across treatments in 2015 and 2016.unctional traits, functional richness, and functional dispersion of dietary niches across treatments in 2015 and 2016. ………….……………..………...….…19

Table 3: List of all plant species seeded in the pocket prairie sites during the establishment phase in 2014 and 2016. The seed mix consisted of 3 native grasses and 16 native forb species which were planted on 3-12 November 2014. The overseed mix consisted of the six forb species which were added on 28-29 January 2016. ……………………………………………………………………………..32

Table 4: Ants collected from unbaited pitfall traps in vacant lots, pocket prairies, and Metropark forests in 2018 and 2019 in Cleveland, Ohio, USA. ………………………………………….………………………44

Table 5: Ants collected from baits in vacant lots, pocket prairies, and Metropark forests in 2018 in and near Cleveland, Ohio, USA. …….………………………………………………………………………………….…45

Table 6: Observed species richness and first-order and second-order jackknife estimates for individual-based rarefaction curves for vacant lot, pocket prairie, and Metropark forest treatments in 2018 and 2019. …………………………………………………………………………………………..…….45

Table 7: Differences in ant species richness, functional dispersion of ant dietary niches, and body length CWMs. Averages (±SEs) are provided. Results are from GLMMs. ……………….…….…49

Table 8: Results of PLSR analyses examining the influence of local soil and vegetation variables, and landscape variables on ant species richness, ant dietary dispersion, and CWM body length. …………………….………………………………………………………………………………………………………………………….52

List of Figures

Figure 1: Locations of vacant lot and urban meadow treatments established across eight neighborhoods within inner city Cleveland, Ohio, USA. ……………………………………………………….……7

Figure 2: Vacant lot and urban meadow treatments established across eight inner-city neighborhoods in Cleveland, Ohio, USA. A: Vacant lot consisting of turf grass species planted by the city post-demolition of the pre-existing structure, and mown monthly to a height of 10 cm. B: Urban meadow consisting of turf grass species planted by the city post-demolition of the pre- existing structure and then mown annually in October to a height of 10 cm. ……………………………7

Figure 3: Effects of mowing frequency on ant species richness in 2015 (A), and 2016 (B) in vacant lot and urban meadow treatments in Cleveland, Ohio, USA. …………………………………….…14

Figure 4: Ants species richness across the June, July, and August sampling intervals in 2015 (A), and 2016 (B). Ants were sampled in vacant lot and urban meadow treatments across eight inner-city neighborhoods in Cleveland, Ohio, USA. …………………………………………………………………14

Figure 5: Species accumulation curves for ant communities in vacant lots and urban meadows in 2015. ……………………………………………………………………………..……………………………………………………….15

Figure 6: Species accumulation curves for ant communities in vacant lots and urban meadows in 2016. ……………………………………………………………………………………………………………………………………….15

Figure 7: Nonmetric multidimensional scaling (NMDS) ordination of ant assemblages for the vacant lot and urban meadow treatments in 2015 using the Bray-Curtis distance measure. Ordination stress value: 0.1749. Plot in 2 dimensions. ……………………………………………………….……16

Figure 8: Nonmetric multidimensional scaling (NMDS) ordination of ant assemblages for the vacant lot and urban meadow treatments in 2016 using the Bray-Curtis distance measure. Ordination stress value: 0.117. Plot in 2 dimensions. ………………………………………………………………16

Figure 9: Spearman rank correlations for ant species richness and proportion of impervious surface area in 2015 at 500 m (A), 1,000 m (B), 1,500 m (C), 2,000 m (D) landscape buffers. ….17

Figure 10: Spearman rank correlations for ant species richness and proportion of impervious surface area in 2016 at 500 m (A), 1,000 m (B), 1,500 (C), 2,000 m (D) landscape buffers. ………18

Figure 11: Locations of vacant lot and pocket prairie treatments established across eight neighborhoods within inner city Cleveland, and Metropark forests near Cleveland, Ohio, USA. …………………………………………………………………………………………………………………………………………………30

Figure 12: Vacant lot and pocket prairie treatments established across eight inner-city neighborhoods in Cleveland, Ohio, and Metropark forest sites located in and around Cleveland, Ohio, USA. A: Vacant lot consisting of turf grass species planted by the city post-demolition of the pre-existing structure, and mown monthly to a height of 10 cm. B: Pocket prairie consisting of native flowering plants originally established in 2014, and mown annually. C: Metropark forest consisting of forested land unmanaged since at least 1930. ………………………..…………….…31

Figure 13: The general pitfall and bait station plot used for all field sites. Pitfall traps were roughly 3 meters from the center and 6 meters apart from the opposite pitfall. Baits were also roughly 3.5 meters from the center and 6 meters apart from the opposite bait. ……………….……33

Figure 14: Locations of vacant lot, pocket prairie, and Metropark forest sites used in the baiting subset in and near Cleveland, Ohio, USA. ……………………………………………………………………………….35

Figure 15: Individual-based species accumulation curves for ant communities in vacant lots, pocket prairies, and Metropark forests in 2018 from pitfall traps. ……………………………………….…46

Figure 16: Individual-based species accumulation curves for ant communities in vacant lots, pocket prairies, and Metropark forests in 2019 from pitfall traps. ………………………………..………..46

Figure 17: Species accumulation curves for ant communities in vacant lots, pocket prairies, and Metropark forests in 2018 from baits. …………………………………………………………………………………….47

Figure 18: Ants species richness across the June, July, and August sampling intervals in 2018 (A), and 2019 (B) from pitfalls. Ants were sampled in vacant lot and pocket prairie treatments across eight inner-city neighborhoods in Cleveland, Ohio, USA, and across eight Cleveland Metropark system forests located within and near Cleveland. Raw data is displayed. ………..…..48

Figure 19: Ants species richness across the June, July, and August sampling intervals in 2018 from the baits. Ants were sampled in vacant lot and pocket prairie treatments across four inner-city neighborhoods in Cleveland, Ohio, USA, and across four Cleveland Metropark system forests located within and near Cleveland. Raw data is displayed. ..………………………………………..48

Figure 20: Nonmetric multidimensional scaling (NMDS) ordination of ant assemblages collected from pitfalls for the vacant lot (VL), pocket prairie (PP), and Metropark forest (MP) treatments in 2018 using the Bray-Curtis distance measure. F = 10.9; P < 0.001, ordination stress value: 0.06. Plot in 2 dimensions. …………………………………………………..…………………………………………….……50

Figure 21: Nonmetric multidimensional scaling (NMDS) ordination of ant assemblages collected from pitfalls for the vacant lot (VL), pocket prairie (PP), and Metropark forest (MP) treatments in 2019 using the Bray-Curtis distance measure. F = 8.4; P < 0.001, ordination stress value: 0.09. Plot in 2 dimensions. ………....………………………………………………………………………………………………..…50

Figure 22: Nonmetric multidimensional scaling (NMDS) ordination of ant assemblages collected from baits for the vacant lot (VL), pocket prairie (PP), and Metropark (MP) forest treatments in 2018 using the Bray-Curtis distance measure. F = 6.7; P = 0.002, ordination stress value: 0.01. Plot in 2 dimensions. ………………………………………………………………………………………………………….……51

xii cover (Litter), percentage bare ground (Bare Ground), impervious surface area at 500 m (Imp 500), impervious surface area at 1,000 m (Imp 1000), impervious surface area at 1,500 m (Imp 1500), impervious surface area at 2,000 m (Imp 2000), grass and shrub cover at 500 m (Grass 500), grass and shrub cover at 1,000 m (Grass 1000), grass and shrub cover at 1,500 m (Grass 1500), grass and shrub cover at 2,000 m (Grass 2000), Tree Canopy at 500 m (Canopy 500), Tree Canopy at 1,000 m (Canopy 1000), Tree Canopy at 1,500 m (Canopy 1500), and Tree Canopy at 2,000 m (Canopy 2000). …………………………………………………………………………………………………..……..53

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Chapter 1. Mowing frequency does not influence ant communities living in urban vacant lots

Abstract

Economic and demographic changes within many Midwestern cities has resulted in an overabundance of infrastructure, which is eventually demolished, creating vacant lots. These greenspaces are managed with monthly mowing, a significant expense for cities with a shrinking tax base. Vacant land has the potential to contribute to the retention of biodiversity and provision of ecosystem services. However, monthly mowing could diminish its conservation value. I investigated if reducing mowing frequency from monthly to annually altered the richness of arthropods found within a vacant lot. Furthermore, I examined how the landscape context of these patches influenced their arthropod conservation value. I focused on ants

(Formicidae), which are important biological indicators of habitat quality. Ants were collected monthly (June-August) using pitfall traps from 16 vacant lots within 8 inner-city neighborhoods of Cleveland, Ohio, USA in 2015 and 2016. Within each neighborhood, one vacant lot was mown monthly (vacant lot treatment) and one vacant lot was mown annually in October (urban meadow treatment). I hypothesized that 1) reducing mowing disturbances would positively affect ant species richness, 2) greater surrounding landscape fragmentation (measured as proportion impervious surface at a radius of 500, 1,000, 1,500, and 2,000 m) would negatively affect ant species richness, and (3) ant functional traits will differ among vacant lots and urban meadows because of differences in vegetation and other environmental factors that would attract different ant species. I found no effect of mowing frequency or of vegetation composition on ant species richness, community composition, or functional traits. However,

1 impervious surface area negatively influenced ant species richness. My results illustrated that regardless of mowing regime, vacant land has the potential to support a similar community of urban-adapted ants. However, when selecting land for the purpose of conservation, its landscape context is an important driver of the ant species assemblage.

INTRODUCTION

Urbanization has dramatically influenced the structure and function of ecosystems. As a significant contributor to habitat loss, fragmentation and degradation from the construction of cities is widely recognized as a major contributor to species declines (McIntyre 2000, McKinney

2002, McKinney 2008). Yet while there is an overall growth in urban development and urban populations, some cities have experienced population loss. Many U.S. cities in the “rust belt” such as Cleveland, OH and Detroit, MI have experienced significant population declines (Oswalt and Rienets 2006, Silva 2010). Cleveland alone has lost over 530,000 residents as of 2010 (U.S

Census 2010). These cities are defined as “shrinking cities” (Martinez-Fernandez et al. 2012), whose population loss has resulted from a number of contributors including deindustrialization, economic decline, urban sprawl, and demographic changes (Reckien &

Martinez-Fernandez 2011). This population decline is unlikely to reverse course, and across shrinking cities, an overabundance of abandoned and foreclosed structures are eventually torn down, creating vacant lots (Keating 2010). Cleveland, OH currently manages > 27,000 vacant lots (Western Reserve Land Conservancy 2015) covering over 1,600 hectares (Cleveland Land

Lab 2008) of city property. These lots are expensive to maintain via mowing, commonly have perceived or actual associations with increased crime, and are widely regarded as blighted

2 areas, which can lower the market values of nearby houses (Accordino & Johnson 2000, Garvin et al. 2013, Goldstein et al. 2001).

Despite their negative perception, vacant lots have been found to host a high richness and abundance of many arthropods and provide important ecosystem services (Gardiner et al.

2013, Gardiner et al. 2014, Philpott et al. 2014). For instance, 131 species of bees

(Sivakoff et al. 2018, Turo et al. in review), 35 species of ground beetles (Delgado de la Flor et al,

2017, Perry et al. in press), and 19 genera of rove beetles (Delgado de la Flor et al. 2017) have been documented within vacant lots. Vacant land also supports a higher richness of woody plants and supplies the majority of forest-derived ecosystem services such as improving air quality and reducing the urban heat island effect for inner-city residents (Riley et al. 2018a).

Despite its potential to conserve arthropods, the regular mowing of vacant land may be diminishing the richness and abundance of species able to survive in this habitat. Within the city of Cleveland, OH, vacant lots are mown on a monthly basis, typically to a height of 10 cm.

Regular mowing has been found to be detrimental to arthropod communities in grasslands, and in urban residential mowed lawns (Helden et al. 2018, Giuliano et al. 2018, Watson et al. 2019).

Mowing has been demonstrated to kill up to 50% of the arthropods (Hemmann et al. 1987) and is correlated with reduced abundance and richness within managed grasslands (Buri et al. 2014,

Horton et al. 2003, Wastian et al. 2016). Other organisms such as birds can be adversely affected by mowing in grassland habitats (Simons et al. 2016, Tyler et al. 1998). Even mowing as infrequently as once or twice a year can have an impact on insects, including pollinators who

3 would have fewer flowering plants available immediately after a mowing (Noordijk et al. 2009).

However, for many species marginal reductions in mowing frequency can increase the abundance of arthropods in a landscape (Helden et al. 2018). When grassy habitats are cut less frequently, plant structural complexity increases, providing a greater range of niches for arthropods (Morris 2000). Currently, it is not known how monthly mowing conducted by the city of Cleveland might be affecting arthropods occupying vacant land.

Beyond local management practices, the landscape context of vacant lots could drive the community of arthropods able to colonize and persist in these habitats. Landscape fragmentation is generally a negative driver of arthropod richness and abundance (Liu et al.

2016, McKinney 2002, McKinney 2008). Urban greenspaces are embedded within a matrix of impervious surfaces such as roads, sidewalks, and buildings that make it difficult for arthropods to disperse across these landscapes (Egerer et al. 2017, McKinney 2002). Arthropod species richness and composition has been found to be significantly lower in areas with a greater impervious surface area (Lagucki et al. 2017, McKinney 2002, Su et al. 2015). Larger amounts of impervious surfaces lead to higher temperatures in urban areas compared to rural areas

(Imhoff et al. 2010). Warmer areas in urban centers could act as a filter for arthropods, limiting their ability to colonize highly fragmented urban areas (Diamond et al. 2018, Diamond et al.

2015, Menke et al. 2011). Therefore, vacant lots with greater amounts of surrounding impervious surfaces may have a lower conservation value, which could pose additional challenges for vacant land conservation.

My aim was to determine how reducing the mowing frequency of vacant land from monthly to annually would influence ant species richness and abundance (Objective 1). I also examined if the surrounding landscape influenced the ant communities in these treatments (Objective 2).

Ants were chosen as the study organism because they are biological indicators, which are sensitive to habitat quality (Andersen 1997, Andersen et al. 2002), and can play a large ecological role in their environment (Agosti et al. 2000). Urban development in general can adversely affect ant populations (Sanford et al. 2009, Vepsäläinen et al. 2008). There are numerous ant species that occur in the Cleveland area (Ivanov 2019), they are relatively easy to collect (Agosti et al. 2000), and many species forage on the ground. Urban greenspaces have also been shown to support a large diversity of ant species (Uno et al. 2010). Most ants are generalists and omnivores and will eat a broad range of food sources (Petry et al. 2012,

Symondson et al. 2002, Way & Khoo 1992). My hypotheses were that: (1) reducing mowing disturbances would positively affect ant species richness, (2) greater surrounding landscape fragmentation (measured as proportion impervious surface at a radius of 500, 1,000, 1,500, and

2,000 m) would negatively affect ant species richness, and (3) ant functional traits will be different between vacant lots and urban meadows due to differences in vegetation and other factors that would attract different ant species.

METHODS

Site selection

This study was conducted in Cleveland, Ohio, USA (Figure 1). Cleveland has been declining in population since the 1950’s due to deindustrialization and economic decline. Population decline

5 has resulted in an overabundance of infrastructure, which is eventually torn down creating vacant lots (Reckien & Martinez-Fernandez 2011, Keating 2010). Currently, there are over

27,000 vacant lots covering 1,600 ha (Western Reserve Land Conservancy 2015). These vacant lots are seeded with a fescue grass mixture, and then are mown monthly by the City of

Cleveland.

This research was conducted in eight inner city Cleveland neighborhoods. Data were collected from 16 vacant lot sites, which were approximately 12 x 30 m in size. Within each neighborhood, established: 1) vacant lots mowed monthly to a height of 10 cm (control), per the standard management practice employed by the city of Cleveland, and 2) vacant lots mowed annually in October to a height of 10 cm (urban meadow) (Figure 2).

The control vacant lots consisted of a turf grass, as well as weedy plant species such as white clover (Trifolium repens L.), red clover (Trifolium pretense L.), and chicory (Chicorum intybus L.).

The urban meadow treatment was cut once annually in October. However, as a cue-of-care, a 1 m boundary around each meadow was maintained with monthly mowing. The meadow treatment consisted of turf grass species as well as red clover, Kentucky bluegrass (Poa pratensis L.), and tall fescue (Festuca arundinacea Schreb). Trash at each site was collected twice a week at each site to account for varying amounts of littering, which could potentially affect the movement and foraging of ants at each site, as well as reduce this study’s perception by the general public, and nearby neighbors.

Figure 1: Locations of vacant lot and urban meadow treatments established across eight neighborhoods within inner city Cleveland, Ohio, USA.

A B

Ant Collection and Identification

Ants (Formicidae) were collected three times in 2015 (18-19 June, 15 July, and 17-18 August), and three times in 2016 (8 June, 13 July, and 11 August). Within each vacant lot site, a 105 1 m2 grid (7 m by 15 m) was installed in the center, and pitfall traps were installed in four randomly selected 1 m2 quadrats. Each pitfall trap consisted of a 1 L plastic cup (11.5 cm diameter, 14.0 cm deep) filled halfway with tap water mixed with Blue Dawn (Procter and Gamble

Corporation, Cincinnati, Ohio, USA) dish soap. Pitfall traps were open for seven days. Upon collection, trap contents were poured into a metal strainer and preserved in sample cups containing 70% ethanol.

Ants were identified to species using The Ants of Ohio (Coovert 2005), Ants of North America: A

Guide to the Genera (Fisher and Cover 2007), AntWiki.org (AntWiki 2020), and AntWeb.org

(AntWeb 2020). Nomenclature for ants follows AntWeb.org and AntWiki.org. Only ant workers were identified, with the exception of Tetramorium atratulum (Schenck), a workerless parasite of Tetramorium immigrans (Santschi). Ants that lacked body parts such as the abdomen or the head, or were in poor condition, were not identified.

Ant functional traits that have important ecological significance (Schofield et al. 2016, Fichaux et al. 2019, Parr et al. 2016) were chosen for measurement on five representatives from each collected species. These traits included quantitative measurements of morphological characters and included Weber’s length, head width, mandible length, and clypeus length. The four quantitative functional traits were measured using a compound microscope with an eyepiece

8 reticle (mm). Up to five individuals per species were measured for each trait, with each trait being measured three times per individual. These measurements were then averaged, and the averages of these traits were relativized by the average Weber’s length of each species. I also chose to look at ant dietary niches for each species as a qualitative functional trait, and I obtained dietary niche information from Coovert (2005), Antwiki.org, and Antweb.org. Ant dietary niches included: 1) generalist predator, 2) generalist, 3) sugar feeder, and generalist, 4) seed harvester and generalist, and 5) specialist predator (Parr et al. 2017).

Vegetation Data

Vegetation was measured in each site twice in 2015 (16 June – 3 July and 22 July – 13 August) and three times in 2016 (13-24 June, 11-22 July, and 4-16 August). Within the sampling grid, 20

1 m2 quadrats were randomly selected, and vegetation biomass and dominant plant species diversity were quantified within a centrally placed 0.5 m2 PVC pipe square. The Comparative

Yield Method (Haydock & Shaw 1975) was used to estimate biomass at each field site. Five

“standards” were identified to represent the range of vegetation biomass at each site, with 1 representing the lowest biomass, and 5 representing the highest biomass. Using the five standards, comparative yield scores were estimated for 20 randomly selected quadrats. Once the comparative yield scores were estimated, the vegetation within the five standards was collected and dried (at 75 °C for 36-48 hours) to get the dry weights, which were then used to make a linear regression equation. This equation was used to plug in all 20 estimated yield scores to calculate the biomass in each quadrat. These estimates were averaged to calculate the mean biomass (g/m2) at each field site per year.

Dominant plant species diversity was measured in each site from 20 randomly selected quadrats, and the three most abundant plant species were recorded. The total species counts were made at each field site, and the Shannon-Wiener Index was used to calculate the dominant plant species diversity per site. The total bloom abundance was measured at each field site using six randomly chosen quadrats (1 m2), using a 0.5 m2 PVC square placed in the center. All flowering plant species were recorded within a quadrat, and all blooms per species were counted to get the bloom abundance. Vegetation data was not obtained from one replicate of the Meadow treatment during the 22 July-13 August sampling period in 2015, and the 13-24 June and 4-16 August sampling periods in 2016 due to a citizen mowing of the site.

Landscape Data

The City of Cleveland Planning Commission provided aerial landscape imagery from 2010 for use in this study. Land cover was delineated at a 1 m2 resolution with the following land cover classes: buildings, impervious surfaces (sidewalks, roads, railroads, other paved surfaces), bare soil, grass and shrubs, tree canopy cover over impervious surfaces, and tree canopy cover over vegetation. Water was not taken into account due to the fact that water accounts for less than

2% of the landscape. The landscape composition of impervious surface area was quantified at the 500 m, 1,000 m, 1,500 m, and 2,000 m radii around each field site by combining the surface area (m2) of buildings, roads, and other impervious paved surfaces from the provided landscape data.

Data Analysis

The four pitfall samples at each site were pooled during each sampling month. All statistical analyses were conducted in R Version 3.6.1 (R Core Team 2019). All data were checked for normality and homogeneity of variance. Ant species richness (number of ant species per site) did not meet these assumptions and transformations did not improve normality. Therefore,

Kruskal-Wallace tests were used to compare ant species richness across vacant lot treatments for each year separately. To compare ant species composition across vacant lot treatments, an

Analysis of Similarity (ANOSIM) was conducted, with each year examined separately. These analyses were visualized using a nonmetric multidimensional scaling (NMDS) plots within the R package ‘vegan’ (Oksanen et al. 2011). NMDS ordinations used Bray-Curtis distances in three dimensions. Bray-Curtis distances were at 999 permutations.

Rarefaction analyses were also performed using the packages ‘vegan’ (Oksanen et al. 2011) and

‘BiodiversityR’ (Kindt & Coe 2005) to account for missing traps. Rarefaction analysis was also done to measure if species richness varies between vacant lots and urban meadows. The 2016 rarefaction analysis included the Detroit Shoreway urban meadow. I created species accumulation curves for both years looking at each treatment separately.

Spearman rank correlations were used to assess the relationship between ant species richness and the amount of impervious surface area in the landscape surrounding each site at all four landscape buffers (500, 1,000, 1,500, and 2,000 m). Rank correlations were also used for looking at the relationship between ant species richness and vegetation height, biomass, and

11 bloom abundance data collected separately in 2015 and 2016. Correlations were conducted using the package ‘stats’ (R Core Team 2019).

I calculated functional richness for all functional traits and the community weighted means for each trait using the package ‘FD’ (Laliberte et al. 2014). I then calculated the functional dispersion of ant dietary niches across treatments (Parr et al. 2016, Laliberte & Legendre 2010

2010), also using the ‘FD’ package. I then compared the four traits (Weber’s length, head width, mandible length, and clypeus length), functional richness, and functional dispersion of ant dietary niches using Welch two sample t-tests when normality assumptions were met by the

Anderson-Darling normality test. In cases where data did not meet normality assumptions, such as with functional richness in 2015 and head width in 2016, I ran Kruskal-Wallace tests.

RESULTS

In 2015 and 2016, I found 10,104 total ants, with 20 species being represented in the subfamilies Dolichoderinae, Formicinae, Myrmicinae, and Ponerinae (Table 1). Lasius neoniger

(Emery) (69.3% of the total individuals collected), Tetramorium immigrans (Santschi) (23.0%),

Tapinoma sessile (Say) (3.9%), Solenopsis molesta (Say) (0.9%), and Camponotus pennsylvanicus

(De Geer) (0.9%) were the most abundant species collected in both years. I found two exotic species, Tetramorium immigrans and T. atratulum (Schenck), which made up 10% of the total collected species and 23.0% of the total collected individuals.

Table 1: Ants collected from unbaited pitfall traps in control and urban meadow vacant lots in 2015-2016 in Cleveland, Ohio, USA. 2015 2016 Subfamily Vacant Lot Urban Meadow Vacant Lot Urban Meadow Dietary Species (8) (8) (8) (7) niche

Dolichoderinae 67 91 57 182 Tapinoma sessile (Say) 67 91 57 182 3 Formicinae 1329 786 3098 2022 Brachymyrmex depilis Emery 8 12 10 2 3 Camponotus subbarbatus Emery 1 0 1 1 3 Camponotus nearcticus Emery 0 1 0 1 3 Camponotus pennsylvanicus (De Geer) 23 10 15 44 3 Formica pallidefulva Latreille 34 3 12 3 4 Lasius nearcticus Wheeler 0 0 1 0 3 Lasius neoniger Emery 1244 733 3031 1955 3 Nylanderia faisonensis (Forel) 8 11 1 13 2 Prenolepis imparis (Say) 11 16 27 3 3 Myrmicinae 391 364 766 932 Crematogaster cerasi (Fitch) 4 6 2 9 3 Myrmecina americana Emery 0 1 0 0 4 Myrmica americana Weber 1 2 1 11 2 Myrmica pinetorum Wheeler 0 3 0 0 Solenopsis molesta (Say) 31 16 9 35 2 Stenamma brevicorne (Mayr) 5 4 0 1 5 Temnothorax curvispinosus (Mayr) 0 0 0 2 3 Tetramorium atratulum* (Schenck) 0 0 2 0 Tetramorium immigrans* Santschi 350 332 752 874 2 Ponerinae 6 3 5 5 Ponera pennsylvanica Buckley 6 3 5 5 1 * = exotic Diet Categories: 1 = generalist predator, 2 = generalist, 3 = sugar feeder + generalist, 4 = seed harvester + generalist, 5 = specialist predator

Ant species richness was similar between vacant lot and urban meadow treatments in 2015

(Figure 3A; χ2= 0.67, P = 0.41, Figure 5) and in 2016 (Figure 3B; χ2= 0.73, P = 0.31, Figure 6).

Moreover, ant species richness remained consistent across the growing season in 2015 (Figure

4A; χ2= 2.61, P = 0.27) and 2016 (Figure 4B; χ2 = 0.02; P = 0.63). Ant species composition also was similar between vacant lot and urban meadow treatments in 2015 (Figure 7; R = -0.12; P =

0.95, Stress = 0.17) and in 2016 (Figure 8; R = -0.07; P = 0.78, Stress = 0.17).

Figure 3: Effects of mowing frequency on ant species richness in 2015 (A), and 2016 (B) in vacant lot and urban meadow treatments in Cleveland, Ohio, USA.

Figure 5: Species accumulation curves for ant communities in vacant lots and urban meadows in 2015.

Figure 6: Species accumulation curves for ant communities in vacant lots and urban meadows in 2016.

Ant species richness decreased with increasing proportion of impervious surface in the landscape at 500 m (r = -0.57, P = 0.021), 1,000 m (r = -0.51, P = 0.042), 1,500 m (r =-0.60, P =

0.014), and 2,000 m (r = -0.65; P = 0.006) in 2015 (Figure 9). In 2016, ant species richness decreased with increasing proportion of impervious surface at 1,000 m (r = -0.53, P = 0.039),

1,500 m (r = -0.51, P = 0.049), and 2,000 m (r = -0.59, P = 0.018), but this pattern was not observed at 500 m (r = -0.33, P = 0.23) (Figure 10).

Figure 9: Spearman rank correlations for ant species richness and proportion of impervious surface area in 2015 at 500 m (A), 1,000 m (B), 1,500 m (C), 2,000 m (D) landscape buffers.

Figure 10: Spearman rank correlations for ant species richness and proportion of impervious surface area in 2016 at 500 m (A), 1,000 m (B), 1,500 (C), 2,000 m (D) landscape buffers.

Ant species richness was not significantly correlated with vegetation height (r = 0.26, P = 0.32), plant biomass (r = 0.08, P = 0.75), or bloom abundance (r = -0.36, P = 0.17) in 2015. I also found similar trends when looking at vegetation height (r = -0.05, P = 0.84), plant biomass (r = -0.06, P

= 0.82), and bloom abundance (r = 0.09, P = 0.73) in 2016. Overall, ant community weighted means, functional richness, and functional dispersion for dietary niches were similar among vacant lots and urban meadows in 2015 and 2016 (Table 2).

Table 2: Ant community weighted means for individual functional traits, functional richness, and functional dispersion of dietary niches across treatments in 2015 and 2016.

Year Trait Statistics

2015 CWM Weber’s length t = -0.89, P = 0.39

2015 CWM Head width t = 0.28, P = 0.78

2015 CWM Mandible length t = 0.87, P = 0.40

2015 CWM Clypeus length t = 0.01, P = 0.99

2015 Functional richness χ2 = 0.16, P = 0.68

2015 Function dispersion of dietary t = 0.03, P = 0.98 niches 2016 CWM Weber’s length t = -0.13, P = 0.89

2016 CWM Head width χ2 = 0.28, P = 0.60

2016 CWM Mandible length t = 0.66, P = 0.52

2016 CWM Clypeus length t = 0.16, P = 0.87

2016 Functional richness t = -0.29, P = 0.77

2016 Function dispersion of dietary t = 0.17, P = 0.87 niches

DISCUSSION

Urbanization is well known to reduce species richness and fragment habitats for arthropods

(McIntyre 2000, McKinney 2002). While urbanization is increasing overall on a global scale, many cities which have lost population from economic decline have a growing amount of urban vacant land which has potential for conservation purposes (Gardiner, et al. 2013, Gardiner et al.

2014, Philpott et al. 2014). Although mowing has been shown to negatively affect many different arthropods and vertebrates (Wastian et al. 2016, Tyler et al. 1998), I found that ant richness was not significantly affected by mowing frequency. One explanation could be that

19 many ant species are ground nesting, often have relatively large colonies, and therefore may not lose too many workers during a mowing event. When I looked at specific vegetation parameters across treatments, I did find it somewhat surprising that there was not a significant correlation between ant species richness and vegetation characteristics, such as height, biomass, and bloom abundance. Blooms would likely be more beneficial for pollinators such as bees (Lowenstein et al. 2015, Matteson & Langellotto 2010) and it would likely be expected that bees would be impacted by the monthly mowing in Cleveland vacant lots. The absence of correlations with biomass or vegetation height were particularly surprising considering that a greater amount of plant material would provide more nectar sources, as well as attract more prey-based food sources and/or honeydew sources from aphids for ants to consume. Most ant species collected in this study would be readily able to tend aphids. However, many ant species are omnivorous and often feed on multiple food sources and therefore are not as limited by the reduction of food sources that mowing may cause (Symondson et al. 2002, Way & Khoo 1992).

Also, it is possible that mowing may provide additional food sources to ants in the form of dead arthropods killed by a mowing event. On the other hand, it is also possible that regular mowing may still be killing a significant number of ants, leading to a potential reduction in colony sizes. I was not able to assess how mowing may affect ant abundances in particular. Many pitfall traps may have been very close to an ant colony or colonies and would therefore have given biased representations from a few species.

Ant species richness and composition as well as the diversity and divergence of functional traits were both similar in vacant lots and urban meadows, illustrating that mowing regime is not the

20 principal driver of ant communities. Many of the ants collected commonly in this study can forage long distances from their colonies to locate above ground for food sources (Agosti et al.

2000, Coovert 2005, Traniello 1989), thus resources removed by localized mowing might not influence colony survival and growth. Lasius neoniger was the most common ant found in both years, which is likely due to it preferring to nest and forage in grass-dominated meadows, lawns, cultivated fields, and is often the dominant species in these habitats (Coovert 2005,

Wang et al. 1995). Further, L. neoniger is a generalist omnivore; workers collect aphid honey dew, nectar, and predate other insects to feed their brood (Coovert 2005). Exotic species such as T. immigrans often dominate urban communities (Coovert, 2005, Chin & Bennett 2018, Uno et al. 2010), and this species was also common in this study; together T. immigrans and L. neoniger accounted for 23.0% of the total ant abundance. T. immigrans is highly tolerant of human activity and tends to nest in disturbed habitats, and therefore would likely be able to nest in sites less favorable to other ants (Coovert 2005, Hedges 1998). Uno et al. (2010) found ant species richness to be negatively correlated with T. immigrans, which could possibly be outcompeting many native species (King & Green 1995). Also, resources such as nearby trees, decaying wood, or other nesting substrates might be a larger driver of species distributions. For example, the acorn ant Temnothorax curvispinosus (Mayr), seeks oaks to establish a colony and

C. pennsylvanicus seeks hollow trees or walls of wooden buildings (Coovert 2005). Likewise, C. subbarbatus and C. nearcticus nest in rotting wood, which would need to be present in or near a site for these species (Coovert 2005).

I found that landscape fragmentation, resulting from impervious surface, was a consistent negative driver of ant species richness within both vacant lot treatments. Impervious surface area is known to limit the colonization of patches by many arthropod groups within urban landscapes (Egerer et al. 2017, Lagucki et al. 2017, McKinney 2002, Su et al. 2015). For instance, solitary bees (Bennett & Lovell 2019, Geslin et al. 2016), lady beetles (Rocha et al 2018, Rocha &

Fellowes 2018), and spiders (McKinney 2008) have been found to be negatively affected by impervious surfaces in urban environments. This pattern was also documented for ants in Lyon,

France (Gippet et al. 2017 and ant communities within urban parks and medians within

Manhattan, New York, USA (Savage et al. 2015). Urban heat island effect is also associated with impervious surface area and can play a role in filtering species in urban environments (Aronson et al. 2016, Diamond et al. 2015, Merckx et al. 2018). Some ant species such a T. curvispinosus have been found to tolerate warmer urban temperatures over time which further indicates that impervious surfaces could act as an environmental filter for ants (Diamond et al. 2018, Diamond et al. 2017, Diamond et al. 2015, Menke et al. 2011). Although all urban greenspaces are surrounded by impervious surface, this ranged from 46.2% to 67% at a 2,000 m buffer scale. At this scale, ant communities ranged from 2-12 species across all sites. Other conservation habitat designs on vacant land may be hindered by the challenges posed by a greater impervious surface area in the surrounding landscape.

CONCLUSIONS

Cleveland, OH and other “rust-belt” cities are likely to continue to experience population loss leading to a further increase in vacant land (Oswalt and Rienets 2006, Silva 2010). Urban

22 greenspaces, while being viewed as a blight to many, have the potential to support arthropods

(Gardiner et al. 2013) despite being mown monthly. While mowing did not impact ants in vacant lots, it still may affect pollinators, spiders, beetles, and other arthropods, and therefore more research is needed to understand how mowing frequency influences conservation goals.

Future studies should consider alternative conservation strategies on vacant land such as pocket prairies with native flowering plants may impact arthropods. Wildflower plantings would likely attract pollinators and other arthropod groups and may receive more public support than unmown urban spontaneous vegetation (Riley et al. 2018a). Urban meadows can look unmanaged and unsightly to neighbors and the public (Turo and Gardiner 2019). Greenspace design must be taken into account when looking at ways to improve the conservation value of urban greenspaces to balance the perception of a site to the public and how it can be a better conservation habitat (Haq 2011). Impervious surface area is known to be an impediment to conservation in urban areas (Lagucki et al. 2017, Egerer et al. 2017, McKinney 2002). My study reaffirmed that impervious surface area must be considered when selecting urban sites for conservation purposes, as it had an impact on ant species richness found in vacant lots and urban meadows. This research aims to ultimately influence city planners in their decisions regarding maintenance and conservation of urban greenspaces, including mowing practices to improve conservation but also consider community concern and ideals for neighborhood aesthetics. My hope is that other conservation practices could be used on vacant land to attract beneficial arthropods and reduce the costs of maintenance for cities with a shrinking tax base.

Chapter 2: Using ant bioindicators to assess how vacant land conservation compares to existing suburban Metropark preserves

ABSTRACT

Shrinking cities such as Cleveland, OH, have experienced population loss, resulting in an overabundance of infrastructure that is demolished, creating vacant land. Vacant land is viewed as a blight, but these greenspaces have the potential to promote biodiversity and ecosystem services. Nevertheless, local soil contamination and vegetation, as well as the urban landscape context could limit the conservation value of vacant land. My aim was to determine if creating

“pocket prairies” consisting of native wildflowers on vacant lots promoted ant species richness.

I chose ants because of their ecological value as biological indicators. Pocket prairie ant communities were compared with those in vacant lots and Metroparks. Vacant lots consisted of early successional weedy vegetation representing current inner-city land management practices. Metroparks are suburban forested preserves and are viewed as the principal urban conservation habitat for the greater Cleveland area. I hypothesized that establishing pocket prairies would promote species richness, but that Metroparks would be a superior habitat for ants. Further, I predicted that local variables such as soil contamination and vegetation, and landscape factors such as impervious surface area and greenspace would all have a significant association with ant richness, dietary niches, and average body length. Ants were collected monthly (June – August 2018, and 2019) using pitfalls and with baits (June – August 2018). My findings did not support my hypothesis, as species richness and ant dietary niches were equivalent within vacant lots and pocket prairies and reduced within Metroparks in 2018 and

2019. Interestingly, larger ant species were captured from Metropark forests, as indicated by

24 the negative relationship observed between community-weighted mean ant body length and local soil contamination, which was more common within the urban greenspaces. I also found that ant species richness was negatively influenced by both impervious surface and tree canopy cover. These results indicate that both local pollution and the surrounding landscape can alter the richness and functional trait composition of ants within a habitat, and therefore should be considered when choosing sites for conservation purposes.

INTRODUCTION

Urbanization can negatively impact the structure and function of ecosystems through habitat fragmentation and degradation. Consequently, city construction and growth are widely recognized as major contributors to species loss (McIntyre 2000, McKinney 2002, McKinney

2008). Yet as many cities are gaining population, over 350 worldwide have experienced long term substantial losses (Rieniets 2009). This includes several post-industrial cities such as

Cleveland, OH and Detroit, MI in the Midwestern USA (Oswalt and Rieniets 2006, Silva 2010), which have lost 530,000 and 1,098,000 residents, respectively (U.S Census 2010). These are considered “shrinking cities” (Martinez-Fernandez et al. 2012), wherein population loss results from multitude of contributors such as deindustrialization, economic decline, urban sprawl, and demographic changes (Reckien & Martinez-Fernandez 2011). This decline in population is unlikely to reverse course, and shrinking cities are often left with an overabundance of residential and commercial infrastructure, which is eventually torn down creating vacant land

(Keating 2010). Historically vacant land is often viewed as blight within a community, which can

25 potentially lower the market values of nearby houses (Accordino & Johnson 2000, Garvin et al.

2013, Goldstein et al. 2001).

Despite their negative perception, vacant lots have more recently been considered as an opportunity for shrinking cities to act towards improving their environmental quality and equity

(Burkholder 2012, Kremer et al 2013, Pallagst et al. 2019). As a conservation resource, vacant land has been found to host a high richness and abundance of many arthropods (Gardiner et al.

2013, Gardiner et al. 2014, Philpott et al. 2014). For instance, 131 species of bees (Sivakoff et al.

2018, Turo et al. in review), 35 species of ground beetles (Perry et al. in press), 19 genera of rove beetles (Delgado de la flor et al. 2017), 58 genera of spiders (Delgado de la flor et al. 2020), and 20 species of ants (see Chapter 1) have been documented within vacant lots. Vacant land also supports a higher richness of woody plants and supplies the bulk of forest-derived ecosystem services to inner-city residents such as a reduction in the urban heat island effect, providing shade cover, and improving the air quality for urban residents (Riley et al. 2018a).

When vacant land is managed in ways to promote resident use, increased recreation, and lower stress levels, and even reduced incidence of crime can result (Branas et al. 2011, Carrus et al.

2015, Garvin et al. 2013, Haq 2011, Maas et al. 2006).

Alternative land management strategies for vacant land seek to garner some of these human health and conservation benefits (Kattwinkel et al. 2011, Gardiner et al. 2013). However, it is unclear how urbanization and deurbanization processes such as construction and demolition as well as the industrial legacy shared by many shrinking cities might limit conservation gains. For

26 instance, vacant lots have particularly high levels of soil contamination and compaction (Kim et al. in review, Radford et al. 2001). Plant growth and soil burrowing organisms can be adversely affected by higher levels of soil degradation (Beniston 2013, Kozlowski 1999, Shuster et al.

2014). For instance, compaction reduced species richness within riparian neotropical and tropical forest ecosystems, and only small species were able to excavate and tunnel through these soils (Costa-Milanez et al. 2017, Schmidt et al. 2017). Pollutants can also negatively affect plants and soil dwelling organisms (Tangahu et al. 2011, Migliorini et al. 2004). For example, high levels of heavy metal pollution can adversely affect colony size and worker size in ants

(Eeva et al. 2004; Sorvari et al. 2007).

Further, surrounding landscape composition and configuration will drive community structure in greenspace patches and must be investigated. For instance, conservation-focused greening has been shown to harbor many desirable arthropod species such as bees and butterflies

(Matteson and Langellotto 2010, Lowenstein et al. 2015). However, the landscape context of these efforts, such as establishing native wildflower habitats, is likely to influence conservation gains. For instance, wildflower plantings within highly fragmented urban landscapes may not support the same community of insect species as similar habitat plantings established in rural landscapes. Greenspaces within rural and suburban areas tend to be larger and in closer proximity to natural habitat, which could increase the species pool able to colonize added habitat patches (Tscharntke et al. 2002, Tscharntke et al. 2012). Instead, urban habitat plantings aiming to attract native conservation targets may yield primarily generalist and exotic species

(Burkman & Gardiner 2014, Burkman & Gardiner 2015, Delgado de la Flor et al. 2017, Horváth et al. 2012, Liu et al. 2016, McKinney 2002).

Using ants as bioindicators (Andersen et al. 2002), the hypothesis that urban greening projects can enhance the species and functional richness of vacant land insect communities was tested.

To test this hypothesis, the Cleveland Pocket Prairie Project was established wherein a network of minimally managed vacant lots containing urban spontaneous vegetation (Riley et al. 2018b) were converted to “pocket prairies” seeded with native Ohio wildflowers. Pocket prairies were predicted to support higher ant species and functional trait richness because of a greater plant diversity, which would potentially also provide more food sources for ants. Certainly, a legacy of soil degradation is likely to influence the value of urban greening, therefore, ant species and functional trait richness was predicted to decline with increased soil compaction and contamination. Landscape fragmentation also likely influences the success of urban greening, and therefore, ant species and functional trait richness was predicted to decline with increasing impervious surface area in the surrounding landscape. Finally, to fully understand the value of conservation investment, the number of ant species utilizing pocket prairies must be compared to the regional species pool of potential colonists. To do this, both vacant lot and pocket prairie ant communities were compared with those occupying intact forests within the suburban

Cleveland Metroparks system, and it was predicted that pocket prairie ant communities would share a greater similarity in species composition to forests than vacant lots.

METHODS

Site selection

This study was conducted in Cleveland, Ohio, USA (Figure 11) and surrounding suburbs within

Cuyahoga and Medina counties. Cleveland, OH currently manages > 27,000 vacant lots which encompasses 1,600 hectares of city property (Western Reserve Land Conservancy 2015). These vacant lots are seeded with a fescue grass mixture, and then are mown monthly by the city of

Cleveland. Monthly mowing costs the City of Cleveland approximately $3M per year

(Community Research Partners & ReBuild Ohio 2008). Historically, Cleveland, OH has also been considered the forested city since the 19th century (Berry et al. 2017) located in a region that remains 19.6% forested as of 2014 (Flinn et al. 2018). The forested portions of the Cleveland

Metroparks system have been preserved since 1930 and are considered a valuable conservation resource within the region that supports many forms of wildlife (Berry et al. 2017,

Carreiro et al. 2008, Nesbitt et al. 2017).

In 2018-19, data were collected from 24 sites belonging to three treatments (n=8): 1) Vacant

Lots; 2) Pocket Prairies; and 3) Metropark forests (Figure 12). Vacant lots consisted of fescue grasses and weedy forb species, maintained following standard city protocols (vegetation cut monthly to a height of 10 cm). The pocket prairie habitats were established first by removing any existing vegetation with two herbicide applications (non-selective glyphosate-based) on 28-

30 May and 23-25 June of 2014. The Ohio Prairie Nursery (Hiram, Ohio) then seeded these sites with 3 native grasses and 16 native forb species on 3-12 November 2014 (Table 3). The pocket prairie sites were also overseeded with six forb species on 28-29 January 2016 to improve site

29 appearances (Table 3). The pocket prairie sites were cut annually in October. Trash was collected at the pocket prairies and vacant lots at least once a month to account for litter which could affect ant foraging at each site. Metropark forests consisted of eight forest stands selected from the Cleveland Metropark system in and surrounding Cleveland (Figure 12). The

Metropark forests were used in this study to compare how well pocket prairies can act as an early form of successional habitat and conserve arthropod communities. Within each site, a plot was established (10 m x 7 m) wherein all data collection occurred. In the vacant lots and pocket prairies, plots were established about one meter from the sidewalk. In the Metropark forests, plots were established about 50 m from the forest edge.

Figure 11: Locations of vacant lot and pocket prairie treatments established across eight neighborhoods within inner city Cleveland, and Metropark forests near Cleveland, Ohio, USA.

A B

Seed Mix Overseed Mix Aster novae-angliae Bidens aristosa Coreopisis lanceolata Chamaecrista fasciculata Elymus canadensis Coreopsis tinctoria Eryngium yuccifolium Gaillardia pulchella Eupatorium purpureum Monarda citriodora Liatris spicata Rudbeckia hirta Lobelia siphilitica Monarda fistulosa Penstemon digitalis Ratibida pinnata Schizachyrium scoparium Silphium perfoliatum Silphium terebinthinaceum Solidago riddellii Sorghastrum nutans Tradescantia ohiensis Verbena hastata Vernonia fasciculata Zizia aurea

Ant Collection and Identification

Ants (Formicidae) were collected once a month in June, July, and August in 2018 and 2019 with pitfall traps and ant baits. Pitfall traps were typically set for seven days, but in some cases were out for as little as five, and at most ten days before their contents were collected. At each field site, four pitfall traps were installed (Figure 13) within the research plot, with each pitfall trap consisting of a 1 L plastic cup (11.5 cm diameter, 14.0 cm deep). Pitfall traps contained tap water mixed with a few drops of Blue Dawn (Procter and Gamble Corporation, Cincinnati, Ohio,

USA) dish soap. Pitfall trap contents were strained, and then placed into a sample cup with 70%

32 ethanol for temporary storage until sorting and identification. Between sampling intervals, each pitfall trap was capped with a lid.

Figure 13: The general pitfall and bait station plot used for all field sites. Pitfall traps were ca. 3 m from the center and 6 m from the opposite pitfall. Baits were also ca. 3.5 m from the center and 6 m from the opposite bait.

For the ant baiting, a subset of 12 sites (four from each treatment) were randomly selected for sample collection in 2018 (Figure 14). Four ant bait stations were set once a month during 11-

13 June, 16-20 July, and 13-20 August (Figure 13). Ant baiting was performed from 7:00 am to

6:00 pm, with baits were set out 2-7 hr before collection. The bait stations were placed in between pitfall traps and each bait station consisted of proteinaceous and sugar-based food sources, following Agosti et al. (2000). The proteinaceous bait consisted of one teaspoon of lightly drained Kroger® brand chunk light tuna in oil. The sugar-based bait consisted of one

33 teaspoon ground Sandies® pecan shortbread (Keebler Corporation, Elmhurst, Illinois, USA).

Each bait station consisted of two halves of an index card (7.62 by 12.7 cm), which were placed on the ground secured by an 8D nail driven through each card and into the soil. Tuna bait was placed in the center on one card and cookie crumbs were placed in the center on the other card, 2.5-5.0 cm apart. At the collection time, both bait cards from each bait station were placed in a Kroger® deep dish 1.89 L container with a lid to minimize escaping ants before aspiration. A mouth-operated vial aspirator (1135A Aspirator, BioQuip Products Inc., Compton,

California, USA) with a HEPA filter (1135Y HEPA filter, BioQuip Products Inc., Compton,

California, USA) was used to aspirate the ants off the cards in the plastic containers. These ants were placed in 20 mL scintillation vials with 80% ethanol. Featherweight forceps (4750 Forceps,

BioQuip Products Inc., Compton, California, USA) were used to collect larger ants from the containers.

Figure 14: Locations of vacant lot, pocket prairie, and Metropark forest sites used in the baiting subset in and near Cleveland, Ohio, USA.

Ants were identified to species (Coovert 2005, Fisher and Cover 2007, AntWiki.org, and

AntWeb.org). Nomenclature for ants follows AntWeb.org and AntWiki.org. Of the ants that were collected, only ant workers were identified. The only ant males or queens that were identified were Tetramorium atratulum (Schenck), a workerless parasite of Tetramorium

35 immigrans (Santschi). Ants that lacked body parts such as the abdomen or the head, or were in poor condition, were not identified.

Ant dietary niche, a qualitative functional trait with important ecological significance (Schofield et al. 2016, Fichaux et al. 2019, Parr et al. 2017) was assessed for all collected ant species

(Coovert 2005, Antwiki.org, and Antweb.org). Ant dietary niche categories were: 1) generalist predator, 2) generalist, 3) sugar feeder and generalist, 4) seed harvester and generalist, and 5) specialist predator (Parr et al. 2017). Average body length for each ant species was calculated from Coovert (2005).

Soil Data

Soil collection and analyses followed Kim et al. (in review). Three soil cores measuring 2.5 cm dimeter by 15 cm deep were collected at each site in 2016 and 2017. To quantify soil moisture, soil was weighed following field collection, processed through a 2 mm sieve, dried at 110 °C for

24 hr, and then weighed again. Soil pH was determined using an AccupHastTM pH meter (Fisher

Scientific, USA) on a 1:1 deionized water:soil ratio (Thomas 1996). Soil organic matter (SOM) was calculated as percentage loss on ignition relative to dry soil after heating at 400 °C for 16 hr. Permanganate oxidizable carbon (POX-C) was determined on duplicate subsamples (Culman et al. 2012, Weil et al. 2003). Briefly, 20 mL 0.02 mol L-1 KMnO4 were added to a 50 mL centrifuge tube containing 2.5 g air-dried soil. After shaking for 2 min at 240 oscillations min-1 and allowed to settle for 10 min 0.5 mL supernatant was transferred to another centrifuge tube and diluted with 49.5 ml of deionized water. A Multiskan FC spectrophotometer (Thermo

Scientific, USA) was used to detect sample absorbance values at 550 nm. The POX-C concentration of each sample was calculated using the formula described by Hurisso et al.

(2016).

Soil texture, which included the percentage of clay, was classified using texture triangle and the

Miller and Miller micropipette method (Miller & Miller 1987). To determine the plant-available nitrogen (PAN), measured as ammonium (NH4) and nitrate (NO3) content, all collected soil samples were extracted with the 2M KCl Method (Maynard and Kalra 2007; Chibuike and

Obiora 2014) and then PAN was determined using a QuikChem 8500 automatic flow injection analyzer (Lachat, USA). Soluble elements were determined by extraction with Mehlich 3 reagent (Mehlich 1984) and quantitated by inductively coupled plasma-atomic emission spectroscopy (ICP-OES).

Polycyclic aromatic hydrocarbons (PAHs) and heterocyclic aromatic compounds (HACs) were evaluated using a modified method of the Bligh-Dyer extraction (Buyer & Sasser 2012). Two g sieved soil were placed in a 13 mm x 100 mm screw-cap test tube, and 4 ml Bligh-Dyer extractant (100 ml of 0.15M citric acid buffer 129 with pH 4.0, 250 ml of methanol, and 125 ml of chloroform) and 2 ml methanol/chloroform (2:1 V/V) was added. Several rounds of centrifuging and vortexing were done as described in Kim et al. (in review). The extracts were then analyzed using gas chromatography/mass spectrometry following Kim et al. in review.

Each soil sample had two extractions, and the compounds from each sample were quantified using the peak areas of their target ions, and from the average of the two extractions.

All sieved soil samples were also analyzed for the total elemental content of 13 heavy metals

(Al, As, Ba, Cd, Cr, Cu, Fe, Mn, Ni, Pb, Sb, V, and Zn) by X-ray fluorescence (XRF) on a NitonTM

FXL-959 (Thermo Scientific, USA) according to US EPA Method 6200 (US EPA, 2007). Quality control was verified by periodically running standard reference materials NIST 2709a, RCRA

180-661, NIST 2711a, and a preparation blank consisting of 99.995% SiO2. Any elemental content values that were lower than detected values were substituted for the background limit of each heavy metal based on the US EPA Method (USEPA 2007). To assess overall site-level heavy metal contamination, I quantified the Pollution Load index using the 13 heavy metals. I first calculated a Contamination Factor (CF) (Hakanson 1980, Weissmannová and Pavlovský

2017) for each heavy metal at each site as the ratio of observed concentrations (Csi) compared to the background levels (Cbi) found in the eastern US (US EPA 2007) using the equation

퐶퐹=퐶푠푖/퐶푏푖. Background levels (ug/g) of these elements were 71000.0 for Al, 5.0 for As, 350.0 for Ba, 0.23 for Cd, 45.0 for Cr, 18.0 for Cu, 21000.0 for Fe, 430.0 for Mn, 15.0 for Ni, 19.0 for

Pb, 1.0 for Sb, 60.0 272 for V, and 45.0 for Zn (US EPA 2007). Next, I determined the Pollution

Load Index (PLI) (Tomlinson et al. 273 1980, Weissmannová and Pavlovský 2017), which is an integrated pollution level for each site. The PLI was calculated for each site using this equation:

푃퐿퐼= √(퐶퐹1×퐶퐹2×퐶퐹3...×퐶퐹푛).

Vegetation Data

Ground-level vegetation was sampled in each field site in June, July, and August of 2018 and

2019. Within each plot, 12 subplots (1 m2) were established and 6 of these subplots were randomly selected to measure vegetation. A 1 m2 PVC quadrat was used to sample vegetation

38 within each of the 6 subplots. Within each quadrat, plant richness, bloom abundance, bloom richness, and the percentage cover of grass, forbs, bare ground, woody debris, and leaf litter were quantified. Percentage cover data were visually estimated such that all variable percentages within a quadrat totaled 100%. Plant height (cm) was measured at three points within each vegetation subplot and then averaged.

Landscape data

The City of Cleveland Planning Commission provided aerial landscape imagery for the vacant lot and pocket prairie field sites in 2010. Land cover was delineated at a 1 m2 resolution with the following land cover classes: buildings, impervious surfaces (sidewalks, roads, railroads, etc.), bare soil, grass and shrubs, tree canopy cover over impervious surfaces, tree canopy cover over buildings, and tree canopy cover over vegetation. Water was excluded because it accounted for less than 2% of the landscape. The proportion of impervious surface area, grass and shrubs, and canopy was quantified at the 500 m, 1,000 m, 1,500 m, and 2,000 m radii around each vacant lot and pocket prairie site. Proportion of impervious surface was quantified by combining the surface area (m2) of buildings, roads, and other impervious paved surfaces from the provided landscape data. Proportion of canopy cover was quantified by combining tree canopy cover over impervious surfaces, tree canopy cover over buildings, and tree canopy cover over vegetation.

Data Analysis

Data from the four pitfall samples at each site were pooled for each sampling month. All statistical analyses were conducted in R Version 3.6.1 (R Core Team 2019). All data were checked for normality and homogeneity of variance. Ant species richness consisted of the number of ant species found per site each month from pitfalls and from baits. Functional dispersion of ant dietary niches was calculated using the ‘FD’ package (Laliberte et al. 2014).

Some species such as Myrmica pinetorum Wheeler and Tetramorium atratulum (Schenck) did not have sufficient information available for dietary niche and were therefore excluded from the functional dispersion of diet analysis. Community weighted means (CWM) for ant body length was calculated using the ‘FD’ package (Laliberte et al. 2014).

Generalized Linear Mixed Models (GLMMs) were developed using the package ‘GlmmTmB’

(Brooks et al. 2017) to assess the responses of ant species to the habitat treatment over the growing season. Response variables were ant species richness, functional dispersion of dietary niches, and CWMs for ant body length from ants collected from pitfall traps and from baits.

Predictor variables used in the GLMMs were treatment (vacant lot, pocket prairie, and

Metropark forests) and month (June, July, and August). Each year was analyzed separately. Two models were developed, one with only the simple effects and one with a treatment x time interaction term. For each response variable, these two models were compared using the anova function in the package ‘car’ (Fox & Weisberg 2019). Model assumptions were visually assessed to determine the appropriate distribution for each model. When looking at ant species richness, I used a negative binomial (negbinom2) distribution for 2018 pitfall and bait data. A

Poisson distribution was used for the 2019 pitfall data. A Gaussian distribution was used for functional dispersion of dietary niches and CWMs for both years and collection methods. To account for missing pitfall traps or unequal sampling intervals, the number of operational trap days was incorporated into the GLMMs as an offset term. Trap days were calculated by adding the total number of days each pitfall trap was operational, and then adding all pitfalls per site per sampling interval. A similar offset term was used for bait collected ants by adding the total number of minutes that baits were out at each treatment and each time.

Ant species composition among treatments was visualized using nonmetric multidimensional scaling (NMDS) and assessed with permutational multivariate analysis of variance

(PERMANOVA) using the adonis2 function in the package ‘vegan’ (Oksanen et al. 2011). Each year and sampling method were examined separately. Pairwise permutational multivariate analysis of variance was used to assess composition differences among treatments using the package ‘pairwiseAdonis’ (Arbizu 2017). NMDS ordinations used Bray-Curtis distances in two dimensions and 999 permutations. Ant occurrence was pooled across the three months for each site and analyses were conducted using a presence/absence species matrix.

Rarefaction analyses and jackknife estimates were calculated, which were also performed using the packages ‘vegan’ (Oksanen et al. 2011) and ‘BiodiversityR’ (Kindt and Coe 2005), to evaluate sampling efficiency and overall species richness. For the rarefaction analyses, a species accumulation curve was created for each treatment. First and second order jackknife estimates calculated the number of species likely to be found in a treatment to compare against the

41 actual number found. The first order jackknife richness estimate involves removing one sample containing a unique species, whereas the second order jackknife richness estimate involves removing two samples which contain the same unique species (Gotelli &Colwell 2011). First order jackknife estimates are considered to be more precise but have more bias because one sample is removed, whereas second order jackknife estimates have less bias, but more standard error because two samples are removed from the calculation (Gotelli & Colwell 2011).

Partial Least Squares Regression (PLSR) analyses were used with the package ‘plsdepot’

(Sanchez 2012) to assess how local and landscape factors affected ant species richness, functional dispersion of ant dietary niches, and CWMs for body size in 2018 and 2019 from pitfall traps. A PLSR analysis allows for many response variables and predictors which may be collinear (Carrascal et al. 2009). Local factors included vegetation and soil variables. Vegetation variables included in the analysis were plant richness, bloom richness, bloom abundance, percentage forbs, percentage grass, percentage leaf litter, percentage bare ground, percentage dead wood cover, and vegetation height. Soil variables were pH, PLI, PAN, PAH, HAC, POX-C, moisture content, and percentage clay content. Landscape factors were percentage impervious surface area, grass and shrubs, and tree canopy at four buffers (500, 1,000, 1,500, and 2,000 m), and were only analyzed across the vacant lot and pocket prairie treatments.

RESULTS

In 2018 and 2019, I collected 12,120 ants from pitfall traps belonging to 34 species and the subfamilies Dolichoderinae, Formicinae, Myrmicinae, and Ponerinae (Table 4). Lasius neoniger

(Emery) (59.3% of the total individuals collected), Tetramorium immigrans (Santschi) (19.4%),

Solenopsis molesta (Say) (6.6%), Camponotus pennsylvanicus (De Geer) (3.9%), and Tapinoma sessile (Say) (2.9%) were the most abundant species collected in both years. I found three exotic species, T. immigrans, Tetramorium atratulum (Schenck), and Nylanderia flavipes (Smith), which made up 19.5% of the total collected species and 8.8% of the total collected individuals from pitfall traps. In 2018, from baits, I found 7,761 total ants, with 17 species being represented in the subfamilies Dolichoderinae, Formicinae, and Myrmicinae (Table 5).

Tetramorium immigrans (Santschi) (52.3%), Solenopsis molesta (Say) (29.9%), Lasius neoniger

(Emery) (9.3% of the total individuals collected), Tapinoma sessile (Say) (3.8%), and Lasius americanus (Emery) (1.9%) were the most abundant species collected in 2018. I only found one exotic species, T. immigrans (Schenck), which made up 5.9% of the total collected species and

52.3% of the total individuals collected from baits. Rarefaction analyses and jackknife estimates illustrated that my sampling effort collected most species that likely occurred within vacant lots. However, pitfall trap catches from pocket prairies and especially Metropark forests represents a likely underestimate of species richness (Table 6, Figure 15, 16, and 17). This is reflected by my second order jackknife estimates which indicated that my sampling efforts collected only 57.5% of species in Metropark forests in 2018 and 61.7% of species in 2019.

Table 4: Ants collected from unbaited pitfall traps in vacant lots, pocket prairies, and Metropark forests in 2018 and 2019 in Cleveland, Ohio, USA.

2018 2019 Subfamily Vacant Pocket Metropark Vacant Pocket Metropark Diet Species Lot Prairie Forest Lot Prairie Forest category

Dolichoderinae 45 160 1 2 139 0 Tapinoma sessile (Say) 45 160 1 2 139 0 3 Formicinae 2263 1584 471 2415 1321 78 Brachymyrmex depilis Emery 97 11 0 119 36 0 3 Camponotus caryae (Fitch) 0 0 0 1 0 0 Camponotus chromaiodes Bolton 0 0 0 0 0 15 3 Camponotus subbarbatus Emery 0 1 0 0 0 3 3 Camponotus nearcticus Emery 1 0 1 1 0 1 3 Camponotus pennsylvanicus (De Geer) 4 0 424 2 5 33 3 Formica glacialis Wheeler 0 0 0 0 0 1 3 Formica pallidefulva Latreille 10 10 3 3 7 0 4 Formica subsericea Say 0 0 0 0 0 10 3 Lasius americanus Emery 0 0 12 0 0 6 3 Lasius nearcticus Wheeler 0 1 0 2 1 0 3 Lasius neoniger Emery 2129 1522 29 2269 1240 4 3 Lasius umbratus (Nylander) 0 0 0 0 0 1 3 Nylanderia faisonensis (Forel) 19 30 0 1 25 0 2 Nylanderia flavipes* (Smith) 0 6 0 0 0 0 2 Prenolepis imparis (Say) 3 3 2 17 7 4 3 Myrmicinae 989 630 197 1005 597 169 Aphaenogaster picea (Wheeler) 0 0 5 0 0 12 4 Aphaenogaster rudis (Enzmann) 0 0 1 0 0 0 4 Crematogaster cerasi (Fitch) 0 4 0 0 6 0 3 Myrmecina americana Emery 1 1 0 1 0 4 4 Myrmica americana Weber 17 4 163 15 6 123 2 Myrmica pinetorum Wheeler 0 1 2 0 0 0 Myrmica punctiventris Roger 0 0 13 0 3 17 4 Solenopsis molesta (Say) 101 228 1 233 242 1 2 Stenamma impar Forel 0 0 3 0 0 0 Stenamma brevicorne (Mayr) 2 0 1 1 1 4 5 Temnothorax ambiguous (Emery) 0 0 1 0 0 0 3 Temnothorax americanus (Emery) 0 0 1 0 0 0 Temnothorax curvispinosus (Mayr) 0 2 2 1 3 4 3 Temnothorax Schaumii (Roger) 0 1 0 0 0 0 3 Tetramorium atratulum* (Schenck) 0 0 0 3 0 0 Tetramorium immigrans* Santschi 868 389 4 751 336 4 2 Ponerinae 19 4 1 23 3 4 Ponera pennsylvanica Buckley 19 4 1 23 3 4 1 * = exotic Diet Categories: 1 = generalist predator, 2 = generalist, 3 = sugar feeder + generalist, 4 = seed harvester + generalist, 5 = specialist predator

Table 5: Ants collected from baits in vacant lots, pocket prairies, and Metropark forests in 2018 in and near Cleveland, Ohio, USA.

2018 Subfamily Vacant Lot Pocket Prairie Metropark Diet category Species Forest Dolichoderinae 1 297 1 Tapinoma sessile (Say) 1 297 1 3 Formicinae 435 374 200 Brachymyrmex depilis Emery 8 0 0 3 Camponotus nearcticus Emery 0 0 1 3 Camponotus pennsylvanicus (De Geer) 9 0 22 3 Lasius americanus Emery 0 0 149 3 Lasius neoniger Emery 417 308 0 3 Nylanderia faisonensis (Forel) 1 0 0 2 Nylanderia flavipes* 0 1 0 2 Prenolepis imparis (Say) 0 65 28 3 Myrmicinae 2072 4302 79 Aphaenogaster picea 0 0 47 4 Myrmica americana Weber 0 2 5 2 Myrmica pinetorum Wheeler 0 0 6 Myrmica punctiventris 0 0 11 4 Solenopsis molesta (Say) 220 2095 4 2 Stenamma brevicorne (Mayr) 0 0 1 5 Temnothorax curvispinosus (Mayr) 0 0 4 3 Tetramorium immigrans* Santschi 1852 2205 1 2

* = exotic Diet Categories: 1 = generalist predator, 2 = generalist, 3 = sugar feeder + generalist, 4 = seed harvester + generalist, 5 = specialist predator

Treatment Collection Year First-Order Second-Order Observed method Jackknife Jackknife species Vacant lot Pitfall traps 2018 16.0 16.0 14 Vacant lot Baits 2018 8.9 10.9 7 Vacant lot Pitfall traps 2019 23.9 25.9 18 Pocket prairie Pitfall traps 2018 22.9 26.9 18 Pocket prairie Baits 2018 8.9 10.9 7 Pocket prairie Pitfall traps 2019 18.0 18.0 16 Metropark forest Pitfall traps 2018 28.9 34.8 20 Metropark forest Baits 2018 18.9 23.7 13 Metropark forest Pitfall traps 2019 24.9 30.8 19

Figure 15: Individual-based species accumulation curves for ant communities in vacant lots, pocket prairies, and Metropark forests in 2018 from pitfall traps.

Figure 16: Individual-based species accumulation curves for ant communities in vacant lots, pocket prairies, and Metropark forests in 2019 from pitfall traps.

Figure 17: Species accumulation curves for ant communities in vacant lots, pocket prairies, and Metropark forests in 2018 from baits.

Pitfall trap ant species richness was lower in July in Metropark forests compared to other treatments in 2018 (Figure 18A; χ2 = 20.3, P < 0.001), and was lower in Metropark forests compared to pocket prairies and vacant lots across all months in 2019 (Figure 18B; χ2 = 15.7, P

< 0.001). Ant species richness collected from baits did not vary among treatments (Figure 19; χ2

= 0.81, P = 0.66). Ant functional dispersion of dietary niches was higher in vacant lots compared to Metropark forests (χ2 = 6.47, P = 0.039) in 2018 when collected from pitfalls, but in 2019, was similar among treatments (χ2 = 3.38, P = 0.18) (Table 7). From baits, ant functional dispersion of dietary niches was also similar among treatments (χ2 = 0.093, P = 0.95). Ant body length CWMs were larger on average in Metropark forests compared to vacant lots and pocket prairies in 2018 (χ2 = 107, P < 0.001) and in 2019 (χ2 = 94.4, P < 0.001) when collected from pitfalls (Table 7). Also, ant body length CWMs were larger in Metropark forests compared to vacant lots and pocket prairies when collected from baits (χ2 = 44.4, P < 0.001) (Table 7).

Table 7: Differences in ant species richness, functional dispersion of ant dietary niches, and body length CWMs. Averages (±SEs) are provided. Results are from GLMMs.

Variables Year Collection Vacant Pocket Metropark χ2 P method lots prairies forests Richness 2018 Pitfalls 6.4 (0.9) 7.6 (0.9) 5.5 (0.8) 20.3 < 0.001 Richness 2018 Baits 4 (0.4) 4.2 (0.6) 5 (1.8) 0.81 0.66 Richness 2019 Pitfalls 7.5 (0.7) 7.9 (0.8) 5.9 (0.4) 15.7 < 0.001 Fdisp Diet 2018 Pitfalls 2.1 (0.2) 1.8 (0.1) 1.5 (0.2) 6.47 0.039 Fdisp Diet 2018 Baits 1.4 (0.1) 0.8 (0.04) 1.5 (0.6) 0.093 0.95 Fdisp Diet 2019 Pitfalls 2.2 (0.1) 1.9 (0.1) 1.9 (0.2) 3.38 0.18 CWM Body 2018 Pitfalls 3.2 (0.2) 2.8 (0.1) 5.0 (0.2) 107 < 0.001 length CWM Body 2018 Baits 3.0 (0.3) 3.0 (0.1) 5.2 (0.4) 44.4 < 0.001 length CWM Body 2019 Pitfalls 3.2 (0.2) 3.4 (0.1) 5.3 (0.2) 94.4 < 0.001 length

Ant species composition estimated by pitfall traps was distinct in Metropark forests compared to vacant lots (F = 12.0, P = 0.001) and pocket prairies (F = 14, P = 0.001) in 2018 (Figure 20). Ant community composition was similar among vacant lots and pocket prairies (F = 1.7, P = 0.18). A similar pattern was observed in 2019 with distinct communities of ant species found in

Metropark forests compared to vacant lots ( F = 10.2, P = 0.001) and pocket prairies (F = 8.7, P =

0.001) (Figure 21). However, in 2019, ant species composition also differed among vacant lots and pocket prairies ( F = 3.7, P = 0.011). Ant species composition estimated by baits was significantly different in Metropark forests compared to vacant lots (F = 8.5, P = 0.027) and pocket prairies (F = 7.0, P = 0.039) in 2018 (Figure 22). Ant species composition was similar among vacant lots and pocket prairies (F = 1.2, P = 0.37).

CWM body length was strongly correlated with multiple soil and vegetation variables in both study years. In 2018, the local PLSR model had significant Q2 scores for t1 or t2 (Table 8). The

PLSR axis 1 explained 35% of the variation in the response variables and the PLSR axis 2 explained an additional 15% (Table 8). Sixteen variables had a VIP score ≥ 0.8 on PLSR axis 1 and/or PLSR axis 2 (Figure 23a). The response variable CWM body length was positively associated with soil quality variables and vegetation variables such as the amount of woody debris and leaf litter cover (Figure 23a) and was negatively associated with soil contamination and ground level vegetation variables (Figure 23a). In 2019, the local PLSR model also had significant Q2 scores for both PLSR axis 1 and PLSR axis 2. The PLSR axis 1 explained 35% and

51 the PLSR axis 2 explained an additional 14% of the variation for the response variable CWM body length in this model (Table 8). Fifteen variables had a VIP score ≥ 0.8 on PLSR axis 1 and/or

PLSR axis 2 (Figure 23b). The response variable CWM body length was also more strongly positively correlated with soil quality variables and vegetation variables such as the amount of woody debris and leaf litter cover and negatively associated with contamination and ground level vegetation variables (Figure 23b).

Ant species richness was correlated with impervious surface area, grass and shrubs, and canopy in 2019, but not 2018. In 2019, the landscape PLSR models had significant Q2 scores for either

PLSR axis 1 or PLSR axis 2. The PLSR axis 1 axis explained 15% of the variation, and the PLSR axis

2 explained an additional 19% (Table 8). In this model, 10 predictor variables had a VIP score ≥

0.8 on PLSR axis 1 and/or PLSR axis 2 (Figure 23c). The response variable ant richness was correlated with both PLSR axis 1 and PLSR axis 2, was negatively associated with impervious surface area and canopy cover and was weakly positively associated with percentage grass and shrub cover.

Table 8: Results of PLSR analyses examining the influence of local soil and vegetation variables, and landscape variables on ant species richness, ant dietary dispersion, and CWM body length.

Year Predictors Q2 R2Y R2X t1 t2 t1 t2 t1 t2 2018 Local 0.23 -0.36 0.35 0.15 0.58 0.035 2018 Landscape -0.10 -0.054 0.15 0.081 0.32 0.44 2019 Local 0.28 -0.038 0.35 0.14 0.60 0.091 2019 Landscape -0.18 0.041 0.15 0.19 0.48 0.28

Figure 23: Correlation maps for the PLSR of ants factors and local and landscape factors; A) 2018 pitfalls local, B) 2019 pitfalls local, and C) 2019 pitfalls landscape. Only variables with a VIP score of ≥ 0.8 for axes with a Q2 value ≥ 0.0975 are shown. Abbreviations are as follows: ant species richness (Ant Rich), CWM body length (Body Length), pollution load index (PLI), soil pH (pH), percentage moisture content (Soil M), percentage clay content (Clay), plant available nitrogen (PAN), Polycyclic aromatic hydrocarbons (PAH), heterocyclic aromatic compounds (HAC), permanganate oxidizable carbon (POX-C), plant richness (Plant Rich), bloom number (# Blooms), bloom richness (Bloom Rich), percentage grass cover (Grass), percentage forbs cover (Forb), vegetation height (V Height), percentage dead wood cover (Wood), percentage leaf cover (Litter), percentage bare ground (Bare Ground), impervious surface area at 500 m (Imp 500), impervious surface area at 1,000 m (Imp 1000), impervious surface area at 1,500 m (Imp 1500), impervious surface area at 2,000 m (Imp 2000), grass and shrub cover at 500 m (Grass 500), grass and shrub cover at 1,000 m (Grass 1000), grass and shrub cover at 1,500 m (Grass 1500), grass and shrub cover at 2,000 m (Grass 2000), Tree Canopy at 500 m (Canopy 500), Tree Canopy at 1,000 m (Canopy 1000), Tree Canopy at 1,500 m (Canopy 1500), and Tree Canopy at 2,000 m (Canopy 2000).

DISCUSSION

Urbanization causes habitat fragmentation, which often results in a reduced arthropod species richness within cities relative to the surrounding region (McIntyre 2000, McKinney 2002). Urban greening projects seek to enhance the arthropod richness of cities by improving habitat quality and quantity (Gardiner et al. 2013, Gardiner et al. 2014, Philpott et al. 2014). Although urban development has been shown to adversely affect ants (Sanford et al. 2009, Vepsäläinen et al.

2008), they are common inhabitants of cities (Santos 2016) and are utilized as biological indicators due to their sensitivity to habitat change (Uno et al. 2010) and ease of collection

(Agosti et al. 2000, Andersen et al. 2002). In this study, I hypothesized that both landscape-scale fragmentation and conservation actions would shape ant communities. I tested this hypothesis within pocket prairies; formerly vacant lots seeded with native vegetation. I predicted that this conservation investment would enhance species richness relative to existing vacant land dominated by early successional vegetation. Further, I contrasted these urban ant communities with those found in suburban Metroparks to determine how urbanization and deurbanization processes that result in soil degradation shape species and functional trait distributions. My key findings were that both soil legacy and landscape context influenced ant functional traits and species richness.

My findings did not support the prediction that establishing native pocket prairies on vacant land would enhance the species richness of ants occupying the patch. Instead, I found similar ant communities within these habitats. However, differences in species composition were detected in 2019 among vacant lots and pocket prairies. Surprisingly, when I found differences

54 among site treatments, vacant lots and pocket prairies hosted a greater number of species than

Metropark forests. These findings align with Ossola et al. (2015) who found a greater ant species richness in low complexity habitats with more frequent management. However, they contrast with Uno et al. (2010) who documented a greater ant species richness in urban forests, compared to vacant lots and urban gardens in Toledo, Ohio, and Detroit, Michigan USA.

Importantly, rarefaction illustrated that I severely under sampled the Metropark forest sites.

Therefore, I do not conclude that Metroparks are an inferior habitat and propose that additional sampling methods are necessary to effectively quantify the species occupying these sites. For example, smaller species may be more effectively sampled via Winkler sampling, which has been found to catch more ant species as well as more smaller ant species compared to pitfall traps (Ivanov & Keiper 2009). Further, pitfall traps were lost to vertebrates and heavy rains and utilizing covers or screens (Buchholz & Hannig 2009) is suggested for future studies.

Across rural to urban gradients, the average body size of many organisms has been documented to decline (Alaruikka et al. 2002, Gibbs et al. 2001, Merckx et al. 2018, Sadler et al

2006). For example, medium to larger sized beetles and spiders were more likely to be found in rural areas compared to urban or suburban areas (Alaruikka et al. 2002, Merckx et al. 2018). My findings illustrated that soil variables were strongly related to the average body size of ants occupying a habitat, with larger species negatively correlated with soil contaminants and positively correlated with soil quality variables such as plant available nitrogen, moisture, and organic matter content. Two of the largest ant species collected, C. pennsylvanicus and M. americana, dominated the Metropark forests and were scarce in the city illustrating that a

55 landscape legacy of soil degradation may influence the composition of species an urban habitat is able to support. Soil pollution by heavy metals are well documented to negatively impact ant species by reducing immune responses, reducing aggression, and have been shown to bioaccumulate in ant workers (Del Toro et al. et al. 2017, Sorvari & Eeva 2010, Sorvari et al.

2007). Research on the harmful effects on PACs and HACs on ant communities is sparse, and my findings illustrate that these contaminants may also pose a risk to conservation efforts of some biota in cities. Nonetheless species such as T. immigrans, L. neoniger and S. molesta were commonly found in vacant lots and pocket prairies as well as other urban habitats (Pećarević et al. 2010) and appear to tolerate soil degradation. These species are generalist and opportunistic species, which have large aggressive colonies that may allow them to outcompete other ants

(Chin & Bennett 2018, Carpintero et al. 2003, Silverman 2005, Uno et al. 2010). The greater aggression in T. immigrans would make it better adept at dominating food sources (Chin &

Bennett 2018). Impervious surface area was found to be a driver of ant species richness across urban greenspaces, as was found in Chapter 1, however, an inconsistent pattern was found in this study. While all urban greenspaces in this study are surrounded by impervious surfaces, this ranged from 45.5% to 67.0% at a 2,000 m buffer scale. As I documented in 2019, impervious surface area is frequently correlated with reduced species richness (Egerer et al.

2017, Lagucki et al. 2017, McKinney 2002, Su et al. 2015). For instance, solitary bees, and some groups of spiders are reduced within urban greenspaces embedded in dense impervious landscapes (Bennett & Lovell 2019, Geslin et al. 2016, Lowe et al. 2017, Lowe et al. 2014,

McKinney 2008, Moorhead & Philpott 2013). In addition to fragmenting habitat, impervious surfaces create heat islands, which could negatively impact ant dispersal and establishment

(Aronson et al. 2016, Diamond et al. 2015, Merckx et al. 2018). I was surprised to find that tree canopy cover was negatively correlated with ant species richness. Ants in urban landscapes are possibly foraging more in these trees rather than on the ground, resulting in this finding. Also, a greater grass and shrubs coverage in the surrounding landscape likely provided more favorable conditions for ant colonization in vacant lots and pocket prairies.

CONCLUSIONS

Vacant parcels are likely to continue increasing in numbers as shrinking cities continue to decline in population (Oswalt and Rienets 2006, Silva 2010). Vacant land has potential to conserve arthropods (Burkholder 2012, Kremer et al. 2013, Pallagst et al. 2019), and urban greening projects are likely to further improve its conservation value for some species (Gardiner et al. 2013). While my research did not show a significant increase in ant species or functional trait richness, I did find a small but significant change in community composition. Over time as these wildflower plantings establish, pocket prairies may host unique communities of ants compared to vacant lots. Pocket prairies could also be more beneficial to other arthropods such as pollinators as well as add aesthetic beauty to neighborhoods with vacant land (Turo &

Gardiner 2019). While both vacant lots and pocket prairies had mostly similar communities of ants, they do not replace the importance of maintaining urban forested sites which provide many ecosystem services such as pollution mitigation, reducing urban heat island effect, and reducing noise pollution (Carreiro et al. 2008, Escobedo et al. 2011). Additional studies focused on other arthropod groups or even on vertebrates are necessary to better understand how pocket prairies can improve the conservation value of vacant land. Soil contamination has been

57 shown to adversely affect plants and soil dwelling organisms, and therefore, could limit the conservation value of vacant land (Beniston 2013, Kozlowski 1999, Shuster et al. 2014). Also, impervious surface area is specifically known to limit conservation potential in urban areas

(Egerer et al. 2017, Lagucki et al. 2017, McKinney 2002). The larger goal of this research was to further influence city planners and urban conservationists on how to execute urban greening initiatives to further improve vacant land conservation, including planting native wildflowers.

Also, community concerns for aesthetics should be considered as pocket prairies can look unsightly to some residents, especially early in the growing season before flowers bloom. I hope that this will lead to more pocket prairie establishment initiatives taken up by cities to boost vacant land conservation as well as lead to other urban greening projects that benefit arthropods as well as urban communities.

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