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Theses and Dissertations
2016
Parasitoid Communities of Remnant and
Constructed Prairie Fragments in Western Ohio
Michael Drew Sheaffer
Wright State University
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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.
iv
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
v
Community Ordination ……. ……………………………………………25
ANCOVA …………. ……………………………………………………26
Edge vs interior habitat effects ……………. ……………………………28
Parasitism rates and orthopteran communities ……. ……………………29
V.
REFERENCES ……. ……………………………………………………………31
VI.
APPENDIX ………... ……………………………………………………………36
vi
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 largescale 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 prairiedependent 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 hostspecificity, 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.
7
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.
8
Table 1. List of study sites including patch size and coordinates.
Site Name
New Carlisle
Hebble Creek
Speedway
Karohl Park
Germantown
Oakes Quarry Park
Englewood
Sugarcreek
Carriage Hill
Possum Creek
Broad-wing Prairie
Sand Ridge Prairie
Goode Prairie Preserve
Stillwater Prairie Reserve
Huffman Prairie
Site Code
POF
HC
SWOF
KP
GT
OQ
ENG
SC
CH
PC
BWP
SRP
GP
SWPR
HP
Type
Size (ha)
Old Field
0.8
Old Field
1
Old Field
1.2
Old Field
3.5
Old Field
18.8
Constructed
4.52
Constructed
6.2
Constructed
11.1
Constructed
22.4
Constructed
24.9
Remnant
0.4
Remnant
1.1
Remnant
3.1
Remnant
6.4
Remnant
39.3
Coordinates
39.969871, -84.028856
39.830727, -84.00243
39.886872, -83.924263
39.740693, -84.04025
39.647128, -84.415845
39.817653, -83.993949
39.867591, -84.269053
39.617715, -84.092614
39.878592, -84.085804
39.713117, -84.265426
39.643028, -84.413313
39.716369, -84.210578
40.167111, -84.404563
40.156469, -84.392472
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
Site
richness
Broad-wing Prairie
34
Charriage Hill
40
Englewood
23
Goode Prairie Preserve
45
Germantown
34
Hebble Creek
31
Huffman Prairie
40
Karohl Park
28
Oakes Quarry Park
48
Possum Creek
33
New Carlisle
16
Sugarcreek
31
Sand Ridge Prairie
31
Speedway
38
Stillwater Prairie Reserve
15
Shannon's
diversity
1.932
2.728
1.856
2.646
2.219
2.332
2.434
2.054
2.617
2.468
1.775
1.589
1.988
2.477
1.459
Effective number
of species
6.9
15.3
6.4
14.1
9.2
10.3
11.4
7.8
13.7
11.8
5.9
4.9
7.3
11.9
4.3
GRASSHOPPER SAMPLING AND PARASITOID REARING
10
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 (RDevelopment-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.
12
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 withingroup 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.
15
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 **
16
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.
17
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.
18
A
B
19
C
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).
20
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)
21
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.
22
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.
24
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.
25
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 prairieassociated 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.
26
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 handcollected 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).
30
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35
VII.
APPENDIX
Table A1. Tachinid fauna collected by site type.
Species/morphospecies
Admontia sp.
Allophorocera sp.
Anisia sp.
Anoxynops aldrichi
Aplomyia theclarum
Archytas apicifer
Campylocheta sp.
Carcelia sp.
Celatoria sp.
Ceracia dentata
Chaetonodexodes vanderwulpi
Cryptomeigenia sp.
Cylindromyia binotata
Cylindromyia decora
Cylindromyia euchenor
Cylindromyia flumipenis
Cylindromyia interupta
Cylindromyia propusilla
Cyzenis sp.
Distichona sp.
Eucelatoria sp.
Euhalidaya genalis
Eulasiona sp.
Exorista sp.
Genea sp. 1
Genea sp. 2
Gymnosoma sp.
Hemisturmia sp.
Hemyda aurata
Leschenaultia sp.
Lespesia sp.
Leucostoma sp.
Linnaemya sp.
Old
Fields
1
1
1
24
3
1
13
2
2
1
18
1
1
1
2
1
Constructed
Prairies
1
2
1
1
1
2
12
2
1
1
1
3
2
7
2
1
3
1
1
1
1
36
Remnant
Prairies
17
1
1
2
3
3
1
1
1
7
1
1
1
2
Total
1
3
1
1
2
3
53
6
1
2
1
3
5
23
7
3
1
1
1
1
28
1
1
1
1
1
1
1
1
1
2
1
3
Lixophaga sp.
Lydina americana
Medina sp.
Myiopharus sp. 1
Myiopharus sp. 2
Nilea sp.
Neomintho celeris
Oxynops anthracinus
Paradidyma sp. 1
Paradidyma sp. 2
Peleteria sp.
Proopia sp.
Siphosturmia sp.
Spathidexia sp. 1
Spathidexia sp. 2
Tachinomyia sp.
Thelaira americana
Thelariodoria sp.
Vibrissina sp. 1
Vibrissina sp. 2
Voria sp.
Winthemia sp. 1
Winthemia sp. 2
Winthemia sp. 3
Winthemia sp. 4
Total Species
Total Individuals
2
18
1
1
4
3
4
2
1
1
27
1
1
2
1
2
0
32
144
1
28
5
1
1
2
6
1
1
1
2
1
1
34
98
14
1
1
5
5
1
3
1
1
10
1
1
2
8
1
1
30
98
2
1
60
1
1
2
5
4
13
1
1
5
7
2
1
1
43
3
1
2
5
11
3
2
1
58
340
Table A2. Orthopteran species collected.
Orthopteran Species
Acrididae
Dichromorpha viridis
Dissosteira carolina
Encoptolophus sordidus
Hippiscus ocelote
Melanoplus angustipennis
Melanoplus differentalis
Melanoplus femurrubrum
Melanoplus sanguinipes
Melanoplus scudderi
Number
Collected
w/ Dipteran
Parasitoids
w/ Nematomorph
Parasitoids
9
2
9
1
3
8
413
22
53
1
1
12
2
1
9
-
37
Melanoplus sp. (nymphs)
Trimerotropis sp.
Tetrigidae
Tettigidea armata
Tettigidea lateralis
Tettigoniidae
Conocephalus brevipennis
Conocephalus fasciatus
Conocephalus strictus
Conocephalus nemoralis
Neoconocephalus retusus
Orchelimum vulgare
Gryllidae
Oecanthus forbesi
Oecanthus sp.
16
1
-
4
1
1
1
-
-
5
2
26
1
1
1
-
-
1
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.
38