Parasitoid Communities of Remnant and Constructed Prairie Fragments in Western Ohio

PARASITOID COMMUNITIES OF REMNANT AND CONSTRUCTED PRAIRIE FRAGMENTS IN WESTERN OHIO

A thesis submitted in partial fulfillment of the requirements for the degree of Master of Science

By:

Michael Drew Sheaffer B.S., University of Delaware, 2013

2016 Wright State University WRIGHT STATE UNIVERSITY GRADUATE SCHOOL

April 25, 2016 I HEREBY RECOMMEND THAT THE THESIS PREPARED UNDER MY SUPERVISION BY MICHAEL DREW SHEAFFER ENTITLED Parasitoid Communities of Remnant and Constructed Prairie Fragments in Western Ohio BE ACCEPTED IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF Master of Science.

______John Stireman, Ph.D. Thesis Director

______David L. Goldstein, Ph.D. Chair, Department of Biological Sciences Committee on Final Examination

______Tom Rooney, Ph.D.

______Volker Bahn, Ph.D.

______Greg Dahlem, Ph.D.

______Robert E. W. Fyffe, Ph.D. Vice President for Research and Dean of the Graduate School ABSTRACT

Sheaffer, Michael Drew. M.S. Department of Biological Sciences, Wright State

University, 2016. Parasitoids communities of remnant and constructed prairie fragments

in western Ohio.

The ability of organisms to disperse to, utilize, and persist in novel habitats in a fragmented landscape is vital to the success of many ecosystem restoration and construction efforts. With less than four percent of original tallgrass prairie persisting across its range, conservationists have made efforts to both protect and restore remnant prairies as well as to plant new prairies. Previous studies suggest that restored ecosystems do not support the same levels of biodiversity and ecosystems services as their remnant counterparts. In this study I measured tachinid fly diversity and orthopteran parasitism rates in order to assess ecological similarity of remnant and constructed prairie fragments and old fields in western Ohio.

Tachinid abundance and richness did not differ among the three site types. Plant richness was the only significant predictor of tachinid richness. Community ordination, while not significant, suggests that old field sites may have more shared species, while both prairie types may support more variable tachinid communities. Significant differences in tachinid abundance were found between habitat edges and interiors for all habitat types. This trend was especially pronounced in constructed prairies, where iii roughly three-fourths of all individuals were captured in edge traps, indicating that the interiors of these sites may be poorly colonized. Grasshopper parasitism rates were significantly lower in constructed prairies than in remnant prairies and old fields, suggesting that prairie planting efforts may not be successful in restoring more intricate ecological interactions. These findings demonstrate the need to consider not only taxonomic diversity but also biological interactions and ecosystem function in evaluating the success of restoration efforts.

TABLE OF CONTENTS

Page

I. INTRODUCTION AND OBJECTIVES ………………………………………....1

Ecological degradation and restoration …………………………………...1

Tallgrass prairie …………………………………… ……………………..2

Arthropods in remnant and restored prairies ……… ……………..………3

Parasitoids …………………………………. …...………………………...4

Objectives …………………………………………………. ……...……...6

II. METHODS ………………………………………………………... ……….…….8

Study sites ………………………………………………………………...8

Tachinid sampling and identification ……………………………………..9

Vegetation sampling ………………. ………...………………………….10

Grasshopper sampling and rearing ………………………………………11

Analysis …………………… ……………………………………………11

III. RESULTS ……………………………………………..…………... ……………13

NMDS community ordination …………….. ……………………………13

ANCOVA …………………………………………. ……………………15

Edge vs interior …………………………………………………….……17

Orthopteran parasitism …….. ……………………………………………23

IV. DISCUSSION ……………………... ……………………………………………25

Community Ordination ……. ……………………………………………25

ANCOVA …………. ……………………………………………………26

Edge vs interior habitat effects ……………. ……………………………28

Parasitism rates and orthopteran communities ……. ……………………29

V. REFERENCES ……. ……………………………………………………………31 VI. APPENDIX ………... ……………………………………………………………36

LIST OF FIGURES

Figure Page

1. Location of study sites ……………….... …………………………………………..…8

2. NMDS ordination plot ………………………… ……………………………………14

3. Tachinid richness and abundance ……………... ……………………………………17

4. Tachinid richness as a function of plant richness ……………... ……………………18

5. Edge vs. interior tachinid richness …………………………….. ……………………18

6. Edge vs. interior tachinid abundance ….. ……………………………………………19

7. Tachinid species accumulation curves ……….... ……………………………………21

8. Tachinid rarefaction across all sites …… ……………………………………………22

A1. Plant diversity metrics ………… ……………………………………………………38

vii

LIST OF TABLES

Table Page

1. List of study sites ……………………… …………………………………………….9

2. Plant diversity metrics ………….……………………………………………………10

3. Base ANCOVA model for tachinid richness ……...... ………………………... ……15

4. Final ANCOVA model for tachinid richness …………………………. …………....16

5. ANCOVA model for tachinid richness with three-way interactions ….. ……………16

6. ANCOVA model for tachinid abundance ……………... ……………………………17

A1. Tachinid fauna collected ……… ……………………………………………………36

A2. Orthopteran fauna collected …………………………………………... ……………37

viii

I. INTRODUCTION

Over the last three hundred years the percentage of Earth’s terrestrial biomes unaltered by human activity has decrease from more than fifty percent to approximately one quarter (Ellis et al., 2010). Considering the ubiquitous effects of anthropogenic climate change and pollution, the actual amount of unaltered landscape may be far lower, if such a thing truly exists at all. A growing body of evidence suggests that this large- scale degradation of the Earth and associated loss of biodiversity threatens to disrupt basic natural processes related to resource acquisition and biomass accumulation, destabilize ecological communities, and reduce the services imparted on humans by the natural environment (Cardinale et al., 2012). Fields of conservation, preservation, and ecological restoration have emerged as tools for combating the negative impacts of humans on the environment and for protecting our long-term societal interests. The need for informed, scientific, goal-oriented approaches in these fields is increasingly recognized (Hobbs & Harris, 2001).

Two major goals of ecological restoration are to maintain or increase biodiversity and to increase the number and quality of ecosystem services provided by the natural environment. A meta-analysis of restoration projects found that while biodiversity and ecosystem services increased significantly post-restoration, both metrics remained lower than in intact, reference ecosystems that had not experienced comparable degradation

(Benayas, 2009). In order to maximize the returns on restoration efforts it is important to

1 study how landscape factors and management decisions may aid or limit a restored or constructed ecosystem’s ability to function in the same manner as its remnant counterparts.

TALLGRASS PRAIRIE

One type of ecosystem which has been the focus of many conservation and restoration efforts in the Midwest United States is the tallgrass prairie. Prairies are grasslands distinguished by unique plant communities with evolutionary adaptations to tolerate drought, fire, and heavy herbivore grazing. Prairies are dominated by grasses and forbs with very little woody vegetation. Plant community composition, soil type, and moisture levels can vary greatly among different types of prairie and across different geographic regions.

Prior to European colonization, tallgrass prairie covered approximately 600,000 square kilometers of the eastern Great Plains region, ranging from central Texas to southern Manitoba and from central Nebraska to the Illinois-Indiana border, with scattered patches occurring farther east (Steinauer & Collins, 1996). It is estimated that prairie covered about three to five percent, or between 2,400 and 4,000 square kilometers, of the state of Ohio (Sears, 1926), which is situated on the eastern periphery of this ecosystem’s range. Studies of pollen in soil strata suggest that, following the last glaciation of the region, a cool, arid climate persisted, which allowed prairie species from the west to colonize areas previously dominated by coniferous forests (Sears, 1930). As the climate of Ohio has since warmed and humidified to its modern state, conifers have

2 been supplanted by deciduous species, and the prairies have receded due to natural succession, the absence of fire in the landscape, and human encroachment, especially in the form of conversion to agricultural land (Pyle et al., 1981). Today only a small fraction of Ohio’s naturally occurring prairies remain, many of which are small, highly fragmented, or heavily degraded. Across its entire range, only about four percent of original tallgrass prairie remains, with some states such as Iowa, Illinois, and Indiana retaining as little as 0.01% of their original prairie (Samson & Knopf, 1996).

Grassland habitats provide a wide variety of ecosystem services to people, including climate regulation, soil conservation, and pollination services, as well as repositories of biodiversity and genetic material and refugia for beneficial organisms such as the natural enemies of insect pests (Isaacs et al., 2008; Sala & Paruelo, 1997). Given these benefits and the increasing rarity of tallgrass prairie, conservationists have made efforts to both protect persisting remnant prairies and to construct new prairie habitat.

These newly constructed prairies may occur in novel locations that did not historically support prairie ecosystems. Questions regarding the ability of such created prairies to support biodiversity and provide ecosystem services on the same levels as their remnant counterparts remain largely unexplored.

ARTHROPOD DIVERSITY IN REMNANT AND RESTORED PRAIRIES

In prairies, as in most terrestrial ecosystems, the vast majority of biodiversity is comprised of insects and other arthropods, making them a logical group with which to study the impacts of restoration. While many insect species are known to be endemic to

3 or characteristic of prairies (Orlofske et al., 2010; Panzer et al., 2010; Wallner et al.,

2012), studies on the differences between the communities of remnant and restored or constructed prairies are deficient and are rarely replicated. Larsen and Work (2003) found ground beetles (Carabidae) to be less diverse in reconstructed prairies than in remnant prairies in Iowa and, additionally, found species that were unique to both prairie types.

Shepherd and Debinski (2005) found diversity and abundance of habitat-sensitive butterflies to be greater in remnant prairies than in integrated or isolated reconstructed prairies while others have noted highly similar lepidopteran communities between remnant and restored prairies (Summerville, 2008). Several groups of bees which were once abundant in and characteristic of tallgrass prairies were found to be rare or absent in reconstructed plots in Illinois (Petersen, 1996). Rowe and Holland (2012) found a positive relationship between plant richness and leafhopper (Cicadellidae) diversity in reconstructed prairies but noted that even communities of similar diversity were not particularly similar in composition to those of remnant prairies, with certain prairie- dependent species absent from reconstructed sites. Another study conducted in Nebraska found Auchenorrhyncha (Hemiptera) to be more diverse in restored than remnant prairies, but found grasshoppers and katydids to be more diverse in remnants. Bomar

(2001) also found grasshopper communities to be more diverse in remnant than reconstructed prairies in Wisconsin. An analysis of family-level arthropod diversity did not find any significant difference between remnant, restored, and reconstructed prairies in Iowa (Orloske et al., 2011).

Representatives of higher trophic levels – including predators and parasitoids – have received little attention in a remnant-versus-restored-prairie context. Parasitoids are

4 insects that lay their eggs on or within other arthropods, upon which their developing larvae feed, eventually killing their host. They are a diverse and integral component of most terrestrial ecosystems, acting as natural enemies of phytophagous insects and playing a significant role in shaping their ecological communities (Godfray, 1994). Due to their dependence on multiple lower trophic levels and their often high levels of host- specificity, parasitoids can be expected to be highly sensitive to ecological conditions and might be one of the final groups to recolonize an ecosystem following restoration

(Henson et al., 2009). On these grounds, parasitoid communities could serve as useful indicators of ecosystem health and of the success of restoration efforts. However, few systematic surveys of prairie parasitoids have been undertaken. Whitefield and Lewis

(2001) conducted a large-scale survey of braconid wasps in tallgrass prairies of Kansas,

Missouri, and Illinois. They collected 924 specimens comprising 184 braconid species, only 25 of which were shared with a nearby Missouri forest community (1,100 specimens, 103 species). Their steep species accumulation curves suggest that they still have not thoroughly sampled the tallgrass prairie braconid communities of that region.

Trailing only the hymenopteran families Braconidae and Ichneumonidae, the dipteran family Tachinidae is one of the largest groups of insect parasitoids with over

1,300 species occurring in North America north of Mexico (O’Hara & Wood, 2004). This super-diverse group of insects parasitizes an extremely broad range of hosts, including

Lepidoptera, Coleoptera, Hemiptera (mostly Heteroptera), Hymenoptera (mostly

Symphyta), Orthoptera, Diptera, and several other insect orders and non-insect arthropods (Eggleton & Belshaw, 1993; Stireman et al., 2006). While tachinids have been employed in numerous biological control programs (Grenier, 1988), they also provide

5 useful systems in which to study ecological questions of biogeography, phylogenetics, and the tritrophic interactions governing the co-evolution of plants, herbivores, and parasitoids (Stireman et al., 2006). Recent work suggests that tachinid diversity may be highly dependent upon remnant habitat fragments in the landscape (Inclán et al., 2014) which makes them an ideal group to study in the heavily fragmented prairie ecosystems of Ohio.

OBJECTIVES

The main objectives of our study were to: (i) broadly characterize and collect baseline data on the tachinid communities of Ohio prairies; (ii) compare tachinid community composition among remnant prairies, constructed prairies, and old fields; (iii) construct a model to determine how habitat type, plant community richness, habitat patch size, and the interactions thereof may affect tachinid richness and abundance; and (iv) compare parasitism rates of a common prairie-associated phytophagous insect group,

Orthoptera, among remnant prairies, constructed prairies, and old fields. Additionally, I analyzed edge-versus-interior habitat effects on tachinid richness and abundance, overall parasitoid abundance (including other dipteran and hymenopteran parasitoid families) among site types, orthopteran diversity and composition among site types, and potential indicator species (both tachinids and orthopterans) for remnant prairies.

I hypothesized that old fields would contain high abundances of relatively few, common species due to their disturbed, successional nature and that remnant prairies would contain the highest overall tachinid richness due to the presence of intact, prairie-

6 dependent communities. I predicted that constructed prairies would not be fully colonized by the natural insect communities of prairie remnants, thus exhibiting intermediate richness and abundance.

I hypothesized that old fields would contain a distinct community of common, generalist species, while remnant prairies would be equally distinct due to the presence of prairie-dependent taxa. I predicted that constructed prairies would not exhibit particularly distinct communities, sharing many of the common species from the old fields and only some taxa more characteristic of remnant prairie ecosystems.

I expected parasitism rates to be highest in both old fields, due to large numbers of common hosts, and in remnant prairies where host-parasitoid interactions would be relatively undisturbed. I expected low parasitism rates in the constructed prairies which may not have had time to establish natural host-parasitoid dynamics.

II. METHODS STUDY SITES

Insects were sampled at fifteen study sites (Table 1) in Greene, Clark, Miami, and

Montgomery counties in western Ohio. Five sites (Huffman Prairie, Goode Prairie

Preserve, Stillwater Prairie Reserve, Sand Ridge Prairie, and Broadwing Prairie) are remnant prairies; five additional sites (Oakes Quarry Park, Carriage Hill MetroPark,

Englewood MetroPark, Sugarcreek MetroPark, and Possum Creek MetroPark) contain constructed prairies; the five remaining sites are old fields or non-prairie meadows located at Karohl Park, Hebble Creek Reserve, Germantown MetroPark, the Speedway

LLC headquarters in Enon, and a private residence in New Carlisle (Figure 1).

Fig. 1. Approximate locations of study sites in Ohio. Orange = remnant prairie, pink = constructed prairie, red = old field. See Table 1 for site codes.

Table 1. List of study sites including patch size and coordinates.

Site Name Site Code Type Size (ha) Coordinates New Carlisle POF Old Field 0.8 39.969871, -84.028856 Hebble Creek HC Old Field 1 39.830727, -84.00243 Speedway SWOF Old Field 1.2 39.886872, -83.924263 Karohl Park KP Old Field 3.5 39.740693, -84.04025 Germantown GT Old Field 18.8 39.647128, -84.415845 Oakes Quarry Park OQ Constructed 4.52 39.817653, -83.993949 Englewood ENG Constructed 6.2 39.867591, -84.269053 Sugarcreek SC Constructed 11.1 39.617715, -84.092614 Carriage Hill CH Constructed 22.4 39.878592, -84.085804 Possum Creek PC Constructed 24.9 39.713117, -84.265426 Broad-wing Prairie BWP Remnant 0.4 39.643028, -84.413313 Sand Ridge Prairie SRP Remnant 1.1 39.716369, -84.210578 Goode Prairie Preserve GP Remnant 3.1 40.167111, -84.404563 Stillwater Prairie Reserve SWPR Remnant 6.4 40.156469, -84.392472 Huffman Prairie HP Remnant 39.3 39.806721, -84.056997

PARASITOID SAMPLING Eight yellow pan traps were deployed at each site approximately every two weeks from late May to mid-September, 2014, for a total of eight sampling dates. Four traps were placed along a transect on the edge of the prairie approximately 20 meters apart.

Habitat edges bordering a forest or a tree line were favored over edges bordering lawn or roadsides in order to maintain consistency among sites. Four additional traps were placed in the prairie interior, approximately 20 meters apart. Traps were recovered after 48 hours and the contents transferred to 70% ethanol for storage. Adult Lepidoptera were disposed of in the field to prevent their scales from contaminating samples. In total 929 samples were collected (15 sites x 8 pan traps per site x 8 sampling dates; 31 traps were disturbed in the field and thus failed to capture any insects).

In the laboratory, parasitoids of the family Tachinidae (Diptera) were sorted from the samples, pinned, dried, and identified to species or morphospecies using available keys (Aldrich, 1926; Guimaraes, 1972; Wood et al., 1987). Hymenopteran and dipteran

9 parasitoids of other families were separated from the samples, counted, and stored separately. Specimens are deposited in the Wright State University Insect Collection,

Dayton, Ohio.

In September and October of 2014, vegetation surveys were conducted at all 15 sites. Sixteen one-square meter plots were selected at each site, and the percent cover of each plant morphospecies occurring in each plot was estimated. Vegetation samples were taken and preserved as necessary in order to facilitate identification. Vegetation data was used to calculate Shannon diversity indices for the plant community from which the effective number of species (Hill, 1973) was calculated for each site (Table 2).

Table 2. Summary of plant community diversity metrics for study sites.

Observed plant Shannon's Effective number Site richness diversity of species Broad-wing Prairie 34 1.932 6.9 Charriage Hill 40 2.728 15.3 Englewood 23 1.856 6.4 Goode Prairie Preserve 45 2.646 14.1 Germantown 34 2.219 9.2 Hebble Creek 31 2.332 10.3 Huffman Prairie 40 2.434 11.4 Karohl Park 28 2.054 7.8 Oakes Quarry Park 48 2.617 13.7 Possum Creek 33 2.468 11.8 New Carlisle 16 1.775 5.9 Sugarcreek 31 1.589 4.9 Sand Ridge Prairie 31 1.988 7.3 Speedway 38 2.477 11.9 Stillwater Prairie Reserve 15 1.459 4.3

GRASSHOPPER SAMPLING AND PARASITOID REARING

In September and October of 2015, Orthopterans (primarily grasshoppers) were sampled at all 15 sites. Specimens were hand-collected along transects until approximately 50 individuals had been captured from each site. The time it took to collect the specimens was recorded in order to calculate a “search effort” index that could be correlated to orthopteran density for each site. These insects were then reared in the laboratory until they died of parasitoid emergence or other causes. Grasshoppers were fed a mixture of lettuce and natural, weedy vegetation and kept under grow-lights on a 16 hours light/8 hours dark cycle. Deceased grasshoppers and emerged parasitoids were removed daily. The experiment was concluded on November 16th, 2015, with any remaining living insects being dispatched in ethanol. Parasitoids (including dipterans, hymenopteran hyperparasitoids, and nematodes) and grasshoppers were preserved and identified (Kirk & Bomar, 2005). Orthopteran diversity and parasitism rates were then calculated for all sites.

ANALYSIS

Shannon’s diversity indices were calculated for the tachinid assemblages and plant communities of each site, from which effective number of species estimates were calculated. Analysis of Covariance (ANCOVA) was performed in R 3.1.2 (R-

Development-Core-Team, 2016) to test for effects of habitat type on tachinid richness and abundance with habitat patch size and plant community richness as covariates, and stepAIC was used on a generalized linear model to determine the optimal model.

The R package vegan (Oksanen et al., 2016) was used to conduct NMDS community ordination analysis on tachinids assemblages. Singletons, which are often the result of stochastic variation or external forces in ecological data (Legendre et al., 1985)

11 were removed from this analysis as they made up a significant portion of the observed taxa and could obfuscate broader trends in tachinid-habitat associations. Vegan was also used to build species accumulation curves for each site as well as to estimate the true tachinid diversity among all sites by rarefaction. Indicator species analyses were performed using the R package indicspecies (De Caceres & Legendre, 2009) to determine whether the presence of any tachinid or orthopteran species were significant predictors of habitat type. A generalized linear model was used to test for differences in parasitism rates between site types and additional ANCOVAs wree used to assess orthopteran diversity and abundance.

III. RESULTS

A total of 340 individual tachinid flies were captured in pan traps during 2014, comprising 45 genera and 58 morphospecies (Appendix 1). Eighteen (40%) genera and

28 (48%) species were represented by a single specimen, while the four most abundant genera (Medina, Campylochaeta, Thelaira, and Cylindromyia) represented 57% of all specimens. Species accumulation curves within each site type were generally quite steep, with no curves approaching apparent asymptotes (Fig. 7). Sample sizes were too low at some sites to obtain meaningful accumulation curves. For example, the constructed prairie site Englewood yielded five individuals representing five species, while the constructed site Carriage Hill yielded nine individuals representing four species.

Rarefaction using each study site as a sample estimated a total tachinid richness of between 70 (bootstrap) and 100 (second-order jackknife) (Fig. 8). Give the large numbers of rare species in our samples, the Chao estimate of 88 may be most appropriate. The

Chao estimator considers the ratio of singletons to doubletons in a sample in calculating and estimated total richness for an assemblage (Chao, 1984; Magurran, 2013).

TACHINID COMMUNITY ORDINATION ANALYSIS

NMDS ordination was performed using Bray-Curtis dissimilarities and random starting configurations. The final solution contained 2 dimensions and had a stress level of 17.5. Stress is a representation of the variance in the data not explained by the model.

Levels above 20 are considered to be unreliable and above 30 are considered entirely

13 arbitrary. Ordination models with a stress under 10 are considered to be a good fit with little risk of making false inferences (Clarke, 1993). Old field and constructed prairie tachinid communities (with the exception of Oakes Quarry) appear to occupy distinct regions of the community ordination plot (Fig. 2). Old fields cover a particularly compact space on the second axis, while both prairie types are spread relatively widely over both axes. An analysis of similarity (ANOSIM) did not find the communities to be significantly differentiated from each other, however, an F test of Bray-Curtis dissimilarities within each site category revealed that old fields may have lower within- group variance in ecological distances than remnant prairies (ratio = 0.32, p = 0.054).

Fig. 2. NMDS ordination of tachinid communities with convex hulls for old fields (1-5), constructed prairies (6-10), and remnant prairies (11-15) with singletons excluded (stress = 17.5). The distance between any two sites in the plot is relative to the ecological

14 distance (number of shared versus unique taxa) of those sites. The names represent tachinid taxa and are positioned nearest the sites of which they are most representative.

ANCOVA

Remnant prairies, constructed prairies, and old fields did not differ in their raw tachinid richness (means of 10, 10, and 11.4, respectively), effective number of species

(8.2, 7.3, 8.4), or abundance (19.6, 19.6, 28.8) (Fig. 3). Large numbers of singletons, steep species accumulation curves (Fig. 7), and exceptionally low abundances at certain sites suggest that tachinid communities were not thoroughly sampled at all locations.

Our base ANCOVA model tested tachinid richness as a function of site type, patch size, and plant community richness with two-way interactions included (Table 3).

Performing stepAIC on our base model returned only plant community richness as being a significant predictor of tachinid richness (p=0.018) (Table 4). A separate regression found plant richness to account for 21% of the variance in tachinid richness (p = 0.086) when other factors were not controlled for (Fig.4). Performing stepAIC on our model with three-way interactions included did not eliminate any elements from the model, returning varying levels of significance on nearly all predictors (Table 5). StepAIC on a model testing tachinid abundance as a function of type, size, and plant richness, found significant effects of site type and plant richness and a number of interactions on tachinid abundance (Table 6).

Table 3. Base ANCOVA model testing tachinid richness as a function of site type, patch size, plant richness, and the two-way interactions of these factors. Null deviance = 25.906 on 14 degrees of freedom.

Coefficients: Estimate Std. Error z value Pr(>|z|) (Intercept) -0.037273 1.190036 -0.031 0.9750 TypeOF 1.936757 1.243155 1.558 0.1192 TypeRP 1.057308 0.924351 1.144 0.2527 Size 0.121080 0.129135 0.938 0.3484 Plant.rich 0.071131 0.028551 2.491 0.0127 * TypeOF:Size 0.022736 0.028720 0.792 0.4286 TypeRP:Size 0.043594 0.022831 1.909 0.0562 . TypeOF:Plant.rich -0.055361 0.032501 -1.703 0.0885 . TypeRP:Plant.rich -0.037827 0.022308 -1.696 0.0900 . Size:Plant.rich -0.004012 0.003473 -1.155 0.2481

Table 4. Final ANCOVA model after stepAIC. Tachinid richness as a function of site type, size, and plant richness. Residual deviance = 20.117 on 13 degrees of freedom. Coefficients: Estimate Std. Error z value Pr(>|z|) (Intercept) 1.625843 0.322232 5.046 4.52e-07 *** Plant.rich 0.021666 0.009131 2.373 0.0177 *

Table 5. StepAIC with three-way interactions included failed to eliminate any predictors from the model. Residual deviance = 2.0822 on 3 degrees of freedom. Coefficients: Estimate Std. Error z value Pr(>|z|) (Intercept) -1.777308 1.424394 -1.248 0.212117 TypeOF 3.653024 1.713491 2.132 0.033013 * TypeRP 6.135957 2.128304 2.883 0.003939 ** Size 0.370587 0.172364 2.150 0.031553 * Plant.rich 0.115887 0.035167 3.295 0.000983 *** TypeOF:Size -0.204559 0.721819 -0.283 0.776875 TypeRP:Size -0.804272 0.318236 -2.527 0.011495 * TypeOF:Plant.rich -0.099471 0.045197 -2.201 0.027749 * TypeRP:Plant.rich -0.172526 0.056392 -3.059 0.002218 ** Size:Plant.rich -0.010911 0.004816 -2.266 0.023477 * TypeOF:Size:Plant.rich 0.006247 0.021126 0.296 0.767462 TypeRP:Size:Plant.rich 0.022056 0.008327 2.649 0.008076 **

Table 6. Final ANCOVA model after stepAIC for tachinid abundance as a function of site type, patch size, and plant richness. Null deviance = 147.188 on 14 degrees of freedom. Residual deviance = 27.948 on 6 degrees of freedom. Coefficients: Estimate Std. Error z value Pr(>|z|) (Intercept) 0.21077 0.67162 0.314 0.753661 TypeOF 2.65853 0.79048 3.363 0.000770 *** TypeRP 2.14760 0.77619 2.767 0.005660 ** Size -0.02319 0.01399 -1.658 0.097263 . Plant.rich 0.08010 0.01404 5.706 1.16e-08 *** TypeOF:Size 0.09089 0.01798 5.056 4.29e-07 *** TypeRP:Size 0.03648 0.01531 2.383 0.017177 * TypeOF:Plant.rich -0.07902 0.02008 -3.936 8.28e-05 *** TypeRP:Plant.rich -0.06661 0.01800 -3.700 0.000216 ***

I found no significant differences in tachinid richness between habitat interiors and habitat edges, although tachinid richness was suggestively higher along edges in both old fields (t = 1.7, df = 4, p = 0.085) and constructed prairies (t = 1.5, df = 5, p = 0.097)

(Fig. 5). A Chi-square test found that proportions of tachinids captured in edges and interiors was significantly different than 1:1 for old fields (p = 0.010), constructed prairies (p < 0.0001), and remnant prairies (p = 0.014) (Fig. 6).

Fig. 3. Raw tachinid richness (left) and abundance (right) in old fields, constructed prairies, and remnant prairies with standard deviations.

Fig. 4. Relationship between plant community richness and tachinid richness among all sites (old fields red, constructed prairies pink, remnant prairies orange).

Fig. 5. Tachinid richness at habitat edges (E) and habitat interiors (I) for old fields, constructed prairies, and remnant prairies with standard errors.

Fig. 6. Proportion of individual tachinids caught on habitat edges (bottom) versus in habitat interiors (top) for old fields (A), constructed prairies (B), and remnant prairies (C).

Fig. 7. Species accumulation curves for old fields (A), constructed prairies (B), and remnant prairies (C) with individuals as samples. (See Table 1 for site codes)

Fig. 8. Estimated richnesses of regional grassland tachinid fauna using each study site as a sample for rarefaction.

PARASITISM RATES AND ORTHOPTERAN COMMUNITIES

Across all sites, 577 orthopteran specimens were collected representing: 6 genera and 10 species of grasshopper (Acrididae); 1 genus and 2 species of pygmy grasshopper

(Tetrigidae); 3 genera and 6 species of katydid (Tettigoniidae); and 1 genus and 2 species of tree crickets (Gryllidae: Oecanthinae) (Appendix 2). A single species, the red-legged grasshopper, Melanoplus femurrubrum (De Geer 1773), accounted for over 70% of all individuals.

I did not find distinct differences among site types in an NMDS community ordination of orthopteran communities, however, an indicator species analysis suggested the species Melanoplus scudderi (Uhuler 1864) is an indicator of remnant prairies (p =

0.017). M. scudderi was collected at four out of five remnant prairies and in only one out of five sites for both constructed prairies and old fields.

Parasitoids reared from grasshoppers included tachinid and sarcophagid flies, hyper-parasitoid wasps in the family Perilampidae, and horsehair worms

(Nematomorpha). Parasitism rates were low: 3.5% for dipterans (n = 20), 2.8% for horsehair worms (n = 16), and 6.3% overall. Orthopterans parasitized by Diptera generally contained more than one maggot, yielding a total of 51 individual dipteran parasitoids (2.55 per host), the majority of which failed to either pupate or eclose as adults. Successfully reared species consisted of the tachinid Ceracia dentata (Coquillett,

1895) and sarcophagid Blaesoxipha atlantis (Aldrich 1916). One C. dentata was reared from the grasshopper Encoptolophos sordidus (Burmeister 1838). Two C. dentata emerged without causing the immediate mortality of their host, and thus cannot be associated with a particular orthopteran species. Both successfully reared B. atlantis were from the same Melanoplus femurrubrum adult female.

Four out of five old field sites and four out of five remnant prairies yielded at least one parasitism event. Parasitism was only observed in a single constructed prairie, although, orthopteran abundance was so low at two constructed prairies that reasonable sample sizes could not be obtained. Our generalized linear model suggests that parasitism in constructed prairies is lower than that in old fields (F = 1.49, df = 12, p = 0.0528) and remnant prairies (F = 1.86, df = 12, p = 0.0128). No differences in orthopteran richness

23 between site types was observed in our ANCOVA model, and the results for orthopteran abundance remain ambiguous due to the inability of stepAIC to remove factors from the model.

IV. DISCUSSION

Community Ordination

A visual interpretation of our community ordination analysis suggests that old field tachinid communities may be distinct from those of both remnant and constructed prairies. The constructed prairie communities also appear to group closely together with the exception of Oakes Quarry Park (site 8) which was an anomaly, having by far the highest species richness (22) and second-highest abundance (55) of all sites. However, analysis of Bray-Curtis dissimilarities among sites did not find these clustering to be significant. Old field communities were not significantly more similar to each other than they were to those of remnant or constructed prairies. Lower within-category variance among old fields compared to remnant prairies suggests that old fields may have more homogenous species compositions, while there are more unshared taxa within the remnant prairie type. The lack of distinctness between remnant and constructed prairies may suggest that prairie-dependent fauna are capable of colonizing constructed prairies, although, the insufficient separation of both prairie types from old fields calls into question whether any of the species captured are truly characteristic of a prairie ecosystem. It is also possible that prairie-adapted species can utilize old fields and constructed prairies just as well as remnants. An indicator species analysis didn’t find any tachinid taxa to be predictive of any of the three site types.

ANCOVA

Existing literature on prairie arthropod communities generally supports the idea that constructed prairies harbor lower biodiversity than their remnant counterparts and often remain uncolonized by prairie-dependent taxa. I predicted that tachinid richness would be higher in remnant than constructed prairies due to the presence of prairie- associated species that had not yet colonized novel sites, but found no differences in tachinid richness among site types while controlling for patch size and plant community richness. Several potential explanations exist for this finding. It could be that prairie construction efforts have been successful in restoring the tachinid communities characteristic of Ohio’s remnant prairies. It is also possible that the remnant sites studied have been degraded to such an extent that they have already lost any prairie-specialized taxa. Finally, it may be that few tachinid species are strongly associated with or truly dependent upon prairie ecosystems and that tachinids which visit open sites may have no preference between prairies and old fields. The fact that neither prairie type possessed communities that were distinct from those of old fields suggests that the tachinids found in constructed prairies are not necessarily colonizing them from remnants.

Insect abundance and diversity, especially those of herbivores, have been shown to be highly correlated with plant community richness (Haddad et al., 2001; Murdoch et al., 1972). When accounting for site type, habitat patch size, and plant community richness in our model with two-way interactions, plant richness was the only significant predictor of tachinid richness across all site types, accounting for 21% of the observed variation. Tachinids may be dependent on plant communities for many reasons including their influence on the availability of suitable hosts, floral resources, and mating sites.

Plant community traits besides taxonomic richness may be good or better predictors of tachinid richness such as flowering phenology and functional groups.

The ANCOVA model with three-way interactions included did find significant effects of site type, plant richness, and the interaction thereof on tachinid richness. Due to our small sample sizes, however, I cannot say decisively that these effects are real or make inferences about the relationships between these factors.

Our ANCOVA model for tachinid abundance did suggest that there are significant effects of site type and plant richness on tachinid densities, with potentially marginal effects of patch size as well (Table 6). Old field sites on average yielded 47% more individual tachinids than both remnant and constructed prairies (Fig. 3). Old fields may support high densities of common phytophagous insect species which in turn support large parasitoid populations. Interactions between site type and plant richness or patch size also had significant influence on tachinid abundance (Table 6), however, like before,

I lack the sample size to determine the nature of these potential relationships. Increasing the number of sites sampled would help to elucidate these interactions.

Many additional factors besides those included in our model are likely to influence tachinid communities in prairie ecosystems. Habitat fragmentation and isolation as well as surrounding landscape use have both been shown to influence communities of tachinids (Inclán et al., 2015) and non-parastioid insects (Stoner & Joern, 2004) in grassland habitats. The age of a constructed prairie is also likely to play a large role in how many species have colonized it, as is the particular assemblage of plant species chosen by land managers for restoration. Additional management practices such as burning and grazing (Panzer, 2002; Vogel et al., 2007) or mowing may also significantly

27 alter species composition.

EDGE VS. INTERIOR HABITAT EFFECTS

Our analysis of tachinids caught in edge versus interior traps indicates that tachinids may not be using prairie interiors as heavily as edges. This trend is especially pronounced in the constructed prairies where 75% of all individuals were captured in edge pan traps. This concentration along the edge of prairies may be an indication that many of the individuals sampled are not truly characteristic of prairie ecosystems and that more attention should be given to those species found in the prairie interiors when attempting to compare communities of remnant and restored sites. Other studies have suggested that open habitats may exhibit large abundances of relatively few species

(Stireman, 2008), and that habitat edges may be particularly diverse, having species representative of both ecosystem types (Inclán & Stireman, 2011). Although I found no differences in tachinid richness between habitat interiors and edges bordering forest, our abundance patterns seem to agree with these previous studies. Given our steep species accumulation curves, and large numbers of singletons, it is possible that greater sampling could reveal differences in tachinid richness between prairie edges and interiors. It is also possible that differences in trap visibility between edges and interiors lead to a sampling bias and that alternative sampling methods or procedures could have produced different results.

While a large number of tachinid species and individuals were captured overall, certain sites produced surprisingly low numbers (five individuals representing five species at Englewood, for example.) It is possible that the dense, tall vegetation of some

28 sites significantly reduced the visibility of the pan traps, which were placed on the ground, reducing the number of individuals captured. It is also possible that additional sampling methods such as malaise traps, hand-collecting, or large-scale rearing would have yielded additional species not seen in the pan traps or greater numbers of species captured only rarely in the pan traps. Tschorsnig (2002) found that in certain locations, pan-trapping yielded significantly more tachinid species than could be collected by hand, while in other locations pan traps failed to capture many species that were easily hand- collected on vegetation directly adjacent to the pan traps. Our steep species rarefaction curves as well as the high numbers of rare species suggest that many of our sites were not thoroughly sampled.

PARASITISM & ORTHOPTERAN COMMUNITIES

Our parasitism data appears to support the hypothesis that constructed prairies do not support the same species interactions found in remnant prairies and old fields. In order for more complex interspecies dynamics such as parasitism to develop, not only do environmental conditions need to be acceptable for both host and parasitoid, but both species must be capable of locating and dispersing to the novel habitat in a complex, fragmented landscape. The diversity of such ecological interactions is arguably as important to ecosystem health as raw species diversity in assessing the success of restoration efforts and implications for the provision of ecosystem services (Dyer et al.,

2010; Henson et al., 2009; Singer & Stireman, 2005).

Analysis of orthopteran diversity among site types and ordination analysis did not reveal any clear patterns. Indicators species analysis did, however, highlight the

29 grasshopper Melanoplus scudderi as being an indicator of remnant prairies, being found in four out of five remnants in our study and in only one of five constructed prairies and old fields. This species is brachypterous, possessing reduced wings incapable of flight, and suggesting it may have a reduced ability to disperse over large distances to colonize novel habitats. While another known prairie-dependent insect, the delphacid planthopper

Aethodelphax prairianus Bartlett & Hamilton, 2011, was observed in multiple constructed sites, it is clear that taxa have differing dispersal abilities, and this must be considered when planning and evaluating restoration efforts. In heavily fragmented landscapes like Ohio’s tallgrass prairie, it may be necessary to create additional patches for the conservation of habitat-specific taxa (Diekötter et al., 2008; Schwartz & van

Mantgam, 1997).

V. REFERENCES

1. Aldrich, J. M. 1926. North American two-winged flies of the genus Cylindromyia Meigen (Ocyptera of authors). Proceeding of the United States National Museum 68: 1-27.

2. Benayas, J. M. R., Newton, A. C., Diaz, A., & Bullock, J. M. (2009). Enhancement of biodiversity and ecosystem services by ecological restoration: a meta-analysis. Science.

3. Bomar, C.R. (2001) Comparison of Grasshopper (Orthoptera: Acrididae) communities on remnant and reconstructed prairies in western Wisconsin. Journal of Orthoptera Research, 10(1):105-112.

4. Cardinale, B. J.; Duffy, E.; Gonzalez, A.; Hooper, D.U.; Perrings, C.; Venail, P.; Narwani, A.; Mace, G.M:; Tilman, D.; Wardle, D.A.; Kinzig, A.P.; Daily, G.C.; Loreau, M.; Grace, J.B.; Larigauderie, A.; Srivastava, D. and Naeem, S. (2012) Biodiversity loss and its impact on humanity. Nature. Volume: 486, Number: 7401, pp 59-67.

5. Chao, A. (1984). Nonparametric estimation of the number of classes in a population. Scandinavian Journal of statistics, 265-270.

6. Clarke, K. R. (1993). Non‐parametric multivariate analyses of changes in community structure. Australian journal of ecology, 18(1), 117-143.

7. De Caceres, M., Legendre, P. (2009). Associations between species and groups of sites: indices and statistical inference. Ecology, URL http://sites.google.com/site/miqueldecaceres/

8. Diekötter, T., Billeter, R., & Crist, T. O. (2008). Effects of landscape connectivity on the spatial distribution of insect diversity in agricultural mosaic landscapes. Basic and Applied Ecology, 9(3), 298-307.

9. Dyer, L. A., Walla, T. R., Greeney, H. F., Stireman III, J. O., & Hazen, R. F. (2010). Diversity of interactions: a metric for studies of biodiversity.Biotropica, 42(3), 281-289.

10. Eggleton, P., & Belshaw, R. (1993). Comparisons of dipteran, hymenopteran and coleopteran parasitoids: provisional phylogenetic explanations. Biological Journal of the Linnean Society, 48(3), 213-226.

11. Ellis, E. C., Klein Goldewijk, K., Siebert, S., Lightman, D., & Ramankutty, N. (2010). Anthropogenic transformation of the biomes, 1700 to 2000. Global ecology and biogeography, 19(5), 589-606.

12. Grenier, S. (1988). Applied biological control with tachinid flies (Diptera, Tachinidae): a review. Anzeiger für Schädlingskunde, Pflanzenschutz, Umweltschutz, 61(3), 49-56.

13. Godfray, H. C. J. (1994). Parasitoids: behavioral and evolutionary ecology. Princeton University Press.

14. Guimaraes, J. H. (1972). A revision of the genus Winthemia Robineau-Desvoidy in America north of Mexico (Diptera, Tachinidae). Arquivos de Zoologia, 22(2), 27-112.

15. Haddad, N.M., Tilman, D., Haarstad, J., Ritchie, M. & Knops, J.M.H. (2001) Contrasting Effects of Plant Richness and Composition on Insect Communities: A Field Experiment. The American Naturalist, 158(1):17-35.

16. Henson, K. S., Craze, P. G., & Memmott, J. (2009). The restoration of parasites, parasitoids, and pathogens to heathland communities. Ecology,90(7), 1840-1851.

17. Hill, M. O. (1973). Diversity and evenness: a unifying notation and its consequences. Ecology, 54(2), 427-432.

18. Hobbs, R. J., & Harris, J. A. (2001). Restoration ecology: repairing the earth's ecosystems in the new millennium. Restoration ecology, 9(2), 239-246.

19. Inclán, D. J., Cerretti, P. & Marini, L. (2014). Interactive effects of area and connectivity on the diversity of tachinid parasitoids in highly fragmented landscapes. Landscape Ecology, 29(5):879-889.

20. Inclán, D. J., Cerretti, P., & Marini, L. (2015). Landscape composition affects parasitoid spillover. Agriculture, Ecosystems & Environment, 208, 48-54.

21. Isaacs, R., Tuell, J., Fiedler, A., Gardiner, M., & Landis, D. (2008). Maximizing arthropod-mediated ecosystem services in agricultural landscapes: the role of native plants. Frontiers in Ecology and the Environment, 7(4), 196-203.

22. Kirk, K., & Bomar, C. R. (2005). Guide to the grasshoppers of Wisconsin. D. J. Watermolen (Ed.). Madison, WI: Bureau of Integrated Science Services, Wisconsin Department of Natural Resources.

23. Ladd, D. (2005). Tallgrass Prairie Wildflowers: A Field Guide to Common Wildflowers and Plants of the Prairie Midwest. Globe Pequot.

24. Larsen, K.J., & Work, T.W. (2003) Differences in ground beetles (Coleoptera: Carabidae) of original and reconstructed tallgrass prairies in northeastern Iowa, USA, and impact of 3-year spring burn cycles. Journal of Insect Conservation, 7:153–166.

25. Legendre, P., Dallot, S., & Legendre, L. (1985). Succession of species within a community: chronological clustering, with applications to marine and freshwater zooplankton. American Naturalist, 257-288.

26. Magurran, A. E. (2013). Measuring biological diversity. John Wiley & Sons.

27. McAlpine, J. F., Peterson, B. V., Shewell, G. E., Teskey, H. J., Vockeroth, J. R., & Wood, D. M. (1987). Manual of Nearctic Diptera. Volume 2 (No. 28).

28. Murdoch, W.W., Evans, F.C. & Peterson, C.H. (1972) Diversity and pattern in plants and insects. Ecology, 53(5):819-829.

29. Nemec, K. T., & Bragg, T. B. (2008). Plant‐Feeding Hemiptera and Orthoptera Communities in Native and Restored Mesic Tallgrass Prairies. Restoration Ecology, 16(2),324-335.

30. O’Hara, J. E., & Wood, D. M. (2004). Catalogue of the Tachinidae (Diptera) of America north of Mexico. Memoirs Entomology International, 18, 1-410.

31. Jari Oksanen, F. Guillaume Blanchet, Roeland Kindt, Pierre Legendre, Peter R. Minchin, R. B. O'Hara, Gavin L. Simpson, Peter Solymos, M. Henry H. Stevens and Helene Wagner (2016). vegan: Community Ecology Package. R package version 2.3-5. 32. http://CRAN.R-project.org/package=vegan

33. Orloske, J.M., Ohnesorg, W.J., & Debinski, D.M. (2010) Potential Terrestrial Arthropod Indicators for Tallgrass Prairie Restoration in Iowa. Ecological Restoration, 28(3), 250-253.

34. Orloske, J.M., Ohnesorg, W.J., & Debinski, D.M. (2011) A comparison of the arthropod communities in remnant, restored, and reconstructed Iowa tallgrass prairies. Natural Areas Journal, 31(2),148-155.

35. Panzer, R. (2002). Compatibility of prescribed burning with the conservation of insects in small, isolated prairie reserves. Conservation Biology, 16(5), 1296- 1307.

36. Panzer, R., Gnaedinger, K. & Derkovitz, G. (2010). The prevalence and status of conservative prairie and sand savanna insects in the Chicago wilderness region. Natural Areas Journal, 30(1), 73-81.

37. Petersen, C.E. (1996) Bee Visitors of Four Reconstructed Tallgrass Prairies in Northeastern Illinois." Proceedings of the 15th North American Prairie Conference.

38. Pyle, R., Bentzien, M. & Opler, P. (1981). Insect conservation. Annual Review of Entomology, 26(1):233-258.

39. R Core Team (2014). R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. URL http://www.R- project.org/.

40. Rowe, H.I., & Holland, J.D. (2013) High Plant Richness in Prairie Reconstructions Support Diverse Leafhopper Communities. Restoration Ecology, 21(2):174-180.

41. Sala, O. E., & Paruelo, J. M. (1997). Ecosystem services in grasslands.Nature’s services: Societal dependence on natural ecosystems, 237-251.

42. Samson, F. B., & Knopf, F. L. (1996) Prairie conservation: preserving North America's most endangered ecosystem. Island Press.

43. Schwartz, M. W., & van Mantgem, P. J. (1997). The value of small preserves in chronically fragmented landscapes. In Conservation in highly fragmented landscapes (pp. 379-394). Springer New York.

44. Sears, P.B. (1926) The Natural Vegetation of Ohio II. The Prairies. The Ohio Journal of Science, 26(3), 128-146.

45. Sears, P. B. (1930). A record of post-glacial climate in northern Ohio. The Ohio Journal of Science, 30(4),205-217.

46. Shepherd, S., & Debinski, D.M. (2005) Evaluation of isolated and integrated prairie reconstructions as habitat for prairie butterflies. Biological Conservation, 126:51–61.

47. Singer, M. S., & Stireman, J. O. (2005). The tri‐trophic niche concept and adaptive radiation of phytophagous insects. Ecology Letters, 8(12), 1247-1255.

48. Steinauer, E. M., & Collins, S. L. (1996). Prairie ecology: The tallgrass prairie. Prairie Conservation. Island Press, Washington, 39-52.

49. Stireman III, J. O., O'Hara, J. E., & Wood, D. M. (2006). Tachinidae: evolution, behavior, and ecology. Annu. Rev. Entomol., 51, 525-555.

50. Stireman, J. O. (2008). α and β diversity of a tachinid parasitoid community over space and time. Annals of the Entomological Society of America, 101(2), 362-370.

51. Stoner, K. J., & Joern, A. (2004). Landscape vs. local habitat scale influences to insect communities from tallgrass prairie remnants. Ecological Applications, 14(5), 1306-1320.

52. Summerville, K. S. (2008). Species diversity and persistence in restored and remnant tallgrass prairies of North America: a function of species’ life history, habitat type, or sampling bias? Journal of Animal Ecology, 77(3), 487-494.

53. Tschorsnig, H.P. (2002) How effective are yellow pan traps for collecting Tachinidae? Tachinid Times, 15, 4-6.

54. Vogel, J. A., Debinski, D. M., Koford, R. R., & Miller, J. R. (2007). Butterfly responses to prairie restoration through fire and grazing. Biological Conservation, 140(1), 78-90.

55. Wallner, A. M., Molano-Flores, B., & Dietrich, C. H. (2013). Using Auchenorrhyncha (Insecta: Hemiptera) to develop a new insect index in measuring North American tallgrass prairie quality. Ecological Indicators, 25, 58- 64.

56. Whitfield, J.B., & Lewis, C.N. (2001) Analytical Survey of the Braconid Wasp Fauna (Hymenoptera: Braconidae) on Six Midwestern U.S. Tallgrass Prairies. Annals of the Entomological Society of America, 94(2):230-238.

VII. APPENDIX

Table A1. Tachinid fauna collected by site type. Old Constructed Remnant Species/morphospecies Fields Prairies Prairies Total Admontia sp. - 1 - 1 Allophorocera sp. 1 2 - 3 Anisia sp. - 1 - 1 Anoxynops aldrichi - 1 - 1 Aplomyia theclarum 1 1 - 2 Archytas apicifer 1 2 - 3 Campylocheta sp. 24 12 17 53 Carcelia sp. 3 2 1 6 Celatoria sp. - 1 - 1 Ceracia dentata - 1 1 2 Chaetonodexodes vanderwulpi - 1 - 1 Cryptomeigenia sp. - 3 - 3 Cylindromyia binotata 1 2 2 5 Cylindromyia decora 13 7 3 23 Cylindromyia euchenor 2 2 3 7 Cylindromyia flumipenis 2 - 1 3 Cylindromyia interupta - - 1 1 Cylindromyia propusilla - - 1 1 Cyzenis sp. 1 - - 1 Distichona sp. - 1 - 1 Eucelatoria sp. 18 3 7 28 Euhalidaya genalis - - 1 1 Eulasiona sp. 1 - - 1 Exorista sp. - 1 - 1 Genea sp. 1 1 - - 1 Genea sp. 2 - - 1 1 Gymnosoma sp. - 1 - 1 Hemisturmia sp. - 1 - 1 Hemyda aurata 1 - - 1 Leschenaultia sp. - 1 - 1 Lespesia sp. 2 - - 2 Leucostoma sp. - - 1 1 Linnaemya sp. 1 - 2 3

Lixophaga sp. 2 - - 2 Lydina americana - 1 - 1 Medina sp. 18 28 14 60 Myiopharus sp. 1 - - 1 1 Myiopharus sp. 2 1 - - 1 Nilea sp. 1 - 1 2 Neomintho celeris - - 5 5 Oxynops anthracinus 4 - - 4 Paradidyma sp. 1 3 5 5 13 Paradidyma sp. 2 - 1 - 1 Peleteria sp. - - 1 1 Proopia sp. 4 1 - 5 Siphosturmia sp. 2 2 3 7 Spathidexia sp. 1 1 - 1 2 Spathidexia sp. 2 1 - - 1 Tachinomyia sp. - - 1 1 Thelaira americana 27 6 10 43 Thelariodoria sp. 1 1 1 3 Vibrissina sp. 1 - 1 - 1 Vibrissina sp. 2 1 - 1 2 Voria sp. 2 1 2 5 Winthemia sp. 1 1 2 8 11 Winthemia sp. 2 2 - 1 3 Winthemia sp. 3 0 1 1 2 Winthemia sp. 4 - 1 - 1 Total Species 32 34 30 58 Total Individuals 144 98 98 340

Table A2. Orthopteran species collected. Number w/ Dipteran w/ Nematomorph Orthopteran Species Collected Parasitoids Parasitoids Acrididae Dichromorpha viridis 9 1 - Dissosteira carolina 2 - - Encoptolophus sordidus 9 1 - Hippiscus ocelote 1 - - Melanoplus angustipennis 3 - - Melanoplus differentalis 8 - 1 Melanoplus femurrubrum 413 12 9 Melanoplus sanguinipes 22 - - Melanoplus scudderi 53 2 -

Melanoplus sp. (nymphs) 16 - 4 Trimerotropis sp. 1 - 1 Tetrigidae Tettigidea armata 1 - - Tettigidea lateralis 1 - - Tettigoniidae Conocephalus brevipennis 5 - - Conocephalus fasciatus 2 - - Conocephalus strictus 26 - - Conocephalus nemoralis 1 - - Neoconocephalus retusus 1 - - Orchelimum vulgare 1 - - Gryllidae Oecanthus forbesi 1 - - Oecanthus sp. 1 - 1

Total: 577 20* 16

*In four instances parasitoids emerged without causing the immediate mortality of their hosts, preventing us from determining, with full certainty, the host species for each individual. The column total includes these four parasitism events.

Fig. A1. Mean plant richness (R) and effective number of species (E) in old field, constructed prairie, and remnant prairie sites.