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2013 SOUTHEASTERN NATURALIST 12(1):121–136
Diversity and Community Similarity of Arthropods in
Response to the Restoration of Former Pine Plantations
John C. Burkhalter1, Daniel C. Moon1, and Anthony M. Rossi1,*
Abstract - Ecological restoration is becoming an increasingly important tool in humanity’s
attempt to manage, conserve, and repair the world’s ecosystems. In the current
study, the objective was to compare the effects of two restoration methods on arthropod
biodiversity and community composition in two former pine plantations; these treatments
included both intensive restoration effort (= cleared) and moderate restoration effort
(= thinned). For the cleared treatment, vegetation was clear-cut to the soil surface, and all
vegetation was removed from the plots, while the thinned treatment consisted of reducing
the Pinus elliotii (Slash Pine) density to that of a native ecosystem and removing of all
exotic plants from the plots as well. Arthropods were sampled by employing pitfall traps,
sticky traps, and sweep netting and identified to family and morphospecies; species richness,
diversity, and community similarity were compared between treatments and sites.
Experimental treatments quickly reached or exceeded arthropod diversity and richness
of an unmanipulated control treatment; however, the two sites produced non-overlapping
ordination plots, suggesting that the diversity of the two sites are either compositionally
different (alpha diversity) or community assemblage is incomplete and overall regional
(beta) diversity has not reached an equilibrium across sites. Additional long-term data
should reveal if these plots are proceeding along different successional trajectories in
terms of community species composition, or whether treatments, while having similar
richness, support different communities because the three types of plots used in this
study (control, thinned, and cleared) represent various successional stages which affect
arthropod species identity, but not overall richness.
Human activities are impairing the normal operation of ecosystem services
and functions on a large scale (Slobodkin 2000). This impairment has led to an
increase in the frequency and importance of attempts to restore degraded ecosystems.
The restoration of biodiversity in an environment can be a critical aspect of
improving ecosystem processes and functions, and the regeneration of community
structure is a vital part of the re-establishment process (Redi et al. 2005, Zerbe
and Kreyer 2006). For instance, managed forests, one type of anthropogenically
disturbed ecosystem, are often viewed negatively from a conservation standpoint
largely because they exhibit reduced biodiversity, and they may develop altered
community composition relative to natural forests (Freedman et al. 1996, Friend
1982). Managed pine plantations are now the most extensive ecosystem in North
Florida, comprising approximately 70% of the forested landscape (Clark et al.
2004). As researchers and land managers strive to understand and ameliorate the
effects of anthropogenic disturbances on ecosystem function and biodiversity,
1Department of Biology, University of North Florida, 1UNF Drive, Jacksonville, FL
32224. *Corresponding author - firstname.lastname@example.org.
122 Southeastern Naturalist Vol. 12, No. 1
there is the need to make informed decisions in a timely and cost-effective
manner. This study seeks to provide land managers with a simultaneous comparison
of multiple restoration techniques, using biologically relevant indicators of
ecosystem function in the form of biodiversity and community composition in a
former pine plantation.
To quantify biodiversity, a number of different groups of organisms have been
monitored including microbes, plants, and animals (Gaines et al. 2007, Kulmatiski
2011, York et al. 2011). Traditionally, restoration studies have focused on
establishing plant communities (Wheater et al. 2000), but it is also important to
examine effects on various fauna to determine the overall success of restoration
(Longcore 2003). Restoration of natural communities relies on the assumption
that with the reestablishment of natural vegetation, reestablishment of the fauna
normally associated with the habitat will follow (Gratton and Denno 2005). However,
little information is available on how arthropod assemblages are affected by
the reintroduction of native flora (Gratton and Denno 2005). It is assumed that
increased plant diversity is likely to result in higher herbivore diversity (Siemann
et al. 1999), but the reality may not be that simple. Insect life-history strategies
are closely attuned to vegetational characteristics of the habitat (Brown 1985),
but there could be a number of factors controlling arthropod community recovery,
e.g., plant-community structural complexity and life-history attributes of the
various insects of the communities (Lawton 1978, Siemann et al. 1998, Steffan-
Dewenter and Tscharntke 1997). Any one or a combination of these factors could
hinder an increase in arthropod diversity in spite of an increase in plant diversity
in a given habitat.
Assessing the progress of restoration projects with arthropods has many
advantages (Finnamore 1996, Kremen et al. 1993). For instance, the short generation
times of most arthropods make them ideal to track year-to-year changes
at a site, while their small size and relatively large population sizes make them
efficient indicators of subtle yet important variations that may influence the quality
of a habitat (Longcore 2003). Arthropod species diversity is a useful metric
of restoration success because of the important roles they fill, such as herbivores
(or plant regulators), pollinators, detritivores, predators, and parasites; they also
serve as important prey sources for reptiles, birds, and mammals (Gardner-Gee
and Beggs 2010, Price 1984, Ruzicka et al. 2010). Insects play a crucial role in
the decomposition of leaf litter, fallen trees, and standing deadwood, making
insects critical for nutrient cycling. Therefore, reduction in insect species abundance
and richness has important functional implications. (Kattan et al. 2006,
Spence 2001). In addition, insects can alter relationships between plant diversity
and ecosystem processes by altering relative abundances of plant species (Mulder
et al. 1999), particularly through pollination services. A critical review by
Maleque et al. (2009) provides a more exhaustive and thorough explanation of
the benefits of using arthropods as indicators of restoration su ccess.
Although species richness is an important attribute of an ecosystem, environmental
variation can produce disparate community composition, even when
species richness is similar. Community similarity measures can indicate what
2013 J.C. Burkhalter, D.C. Moon, and A.M. Rossi 123
effect restoration has on the community on a broader scale. For instance, local
(alpha) diversity, in which habitats are more likely to be homogenous, is typically
more similar than regional (beta) diversity. As a result, alpha and gamma diversity
may vary substantially between these scales (Ricklefs and Miller 2000). Thus,
measures of community similarity provide a metric by which we can determine
how resources are being utilized between sites based upon species composition
and their relative abundances in various communities or treatment groups.
Tree plantations often provide habitat characterized by a closed canopy and
tree densities that are greatly elevated above natural levels; these high densities
may affect underlying community composition and/or resource quality (Kattan
et al. 2006). For example, Zhu et al. (2007) found that plantation densities of
Pinus resinosa Aiton (Red Pine), which range from 89–332/ha, reduced average
tree ring width by approximately 50% in comparison to natural stands, thereby
reducing the growth of trees within plantations. In canopies of lower density, the
environment may be made more hospitable for understory species. Mixed species
stands are also likely to use nutrients more efficiently compared to pure stands
such as pine plantations because of differences among species in a number of
factors such as mycorrhizal associations, shade-tolerance, growth rate, form, nutrient
demands, and abilities to fix nitrogen. In addition, combinations of species
that differ in height, growth form, shade-tolerance, and phenology are also likely
to increase site productivity (Hartley 2002).
A common restoration method utilized to restore higher levels of biodiversity
in managed forest ecosystems is tree density reduction (= thinning). Thinning has
been shown to promote a more balanced stand structure where there is an excess
of smaller trees (Edminster and Olson 1996). Intermittent thinning can also preserve
tree and stand vigor while still maintaining structure (Edminster and Olson
1996). A study by Feeney et al. (1998) demonstrated that thinning produced
improved resource uptake, growth, and insect resistance in stands of Pinus ponderosa
Lawson (Ponderosa Pine). Thinning has also been shown to have positive
effects for species that benefit from burning such as Pinus elliotii Englem (Slash
Pine; Wilson and Watts 1999) and may be a viable alternative when burning is
not an option.
More intensive methods of restoration are sometimes used, ones in which
silviculturists or researchers do not simply remove certain species or decrease
density within the environment, but instead conduct an extensive removal of all
vegetation to restore ecosystem processes (Hobbs 2007). This treatment may
be especially useful for restoring commercially viable pine plantations because
it returns the site to an earlier successional stage and allows the community to
regenerate via secondary succession. The path of succession will be influenced
by a number of factors including seed-bank sources, as well as seed dispersal
from neighboring areas and competitive interactions among and between native
and invasive plants, resulting in ecosystem dynamics that are extremely complex
(Wallington et al. 2005).
Lastly, the least expensive method of “restoration” is passive in which researchers
or land managers do nothing to the site and allow existing successional
124 Southeastern Naturalist Vol. 12, No. 1
processes to occur. While this method of restoration is preferred by some because
it is both inexpensive and easy, it has some disadvantages (Morrison and
Lindell 2011). Passive restoration often takes a very long time, and the outcome
is relatively uncertain due to the fact that it is not directed; thus, the resulting
community may not return to its previous successional trajectory due to factors
such as dispersal limitation (Battaglia et al. 2008). Although factors such as
dispersal could also be an issue with active restoration, the ability of researchers
to “disperse” individuals via intentional planting, could obviate these issues.
The precise objective of the current study was to compare two active techniques
and one passive technique of restoring a former Slash Pine plantation, and the
resultant effects on abundance, diversity, and community similarity of arthropod
assemblages. As stated previously, managed ecosystems such as pine plantations
are often lower in diversity than unmanaged forests, and higher diversity within
the treated areas could indicate levels of diversity closer to a more natural forest.
Slash Pine is native to the southeastern US, but its density on former pine plantations
is significantly higher than that in undisturbed ecosystems. In native ecosystems
of North Florida, densities of Slash Pines in pure stands are typically 2–3
trees per 100 m2, but in the managed ecosystems of a pine plantation densities can
average 20 trees per 100 m2 (J.C. Burkhalter, unpubl. data). The current study utilized
two former pine plantations, McGirt’s Creek (30.26107°N, 81.81843°W) in
southwestern Duval County, FL and Tiger Point (30.5007941°N, 81.4937016°W)
in northeastern Duval County, FL, which have been acquired for conservation
by the city of Jacksonville, FL. Both sites were heavily planted with a Slash Pine
density varying from 20 to 50 trees per 100 m2, and both sites had low herbaceous
growth due to large amounts of leaf litter (for a more thorough description, see
Rossi et al. 2011). McGirt’s Creek is embedded in a mosaic of mostly residential
developments, and Tiger Point is surrounded primarily by a more or less contiguous
landscape of managed or formerly managed pine plantations.
We employed two experimental restoration treatments per site in this study
in addition to controls. One experimental treatment entailed removing all exotic
plant species and thinning the pine trees that previously made up a large amount
of the vegetation in the experimental plots. Each Slash Pine tree within a plot was
assigned a number, and then trees were randomly removed. Trees were removed
until stand density was reduced to 2–3 trees per 100 m2, with all remaining trees
being of similar size and condition so as to not introduce any confounding effects
of tree size. In the second experimental treatment, all vegetation was clear-cut
manually with chainsaws and loppers and then removed from the plot. Although
all vegetation was removed to the surface of the substrate, no below-ground
biomass was removed from the plots. Control plots were not manipulated and
contained a high density (approximately 20 trees/100 m2) of Slash Pine, as well
2013 J.C. Burkhalter, D.C. Moon, and A.M. Rossi 125
as native oaks and other small vegetation. All plots were 10 m x 10 m. Few, if any,
undisturbed reference sites remain in Florida and none near the study sites, and
thus we cannot directly compare the diversity and community similarity of our
restored sites to a pristine reference site. Due to the relatively homogenous biotic
and abiotic conditions within the study plots prior to restoration, it is assumed
that there were no differences in either invertebrate diversity or community similarity.
For each treatment group as well as the controls, there were five replicate
plots at McGirt’s Creek (15 total plots) and three replicates of each at Tiger
Point (9 total plots). All experimental manipulation was completed in February
of 2006, and afterwards plots underwent natural succession; invertebrate biodiversity
was compared approximately two years after the study was initiated. All
arthropod sampling was conducted by the same person throughout the entirety of
the monitoring phase of the experiment. For both the pitfall traps and the sticky
traps, one trap was used per plot. Sampling periods lasted for a month in the
spring of 2008 and 2009, alternating between the two restoration sites each week,
allowing for two weeks of sampling at each site per sampling period. During the
sampling period at each site, collection of the pitfall traps occurred every other
day, while sweep sampling and collection of the sticky traps was performed once
Pit fall traps
Ground-dwelling insects were captured using pitfall traps constructed from
small buckets that were approximately 15.2 cm in both diameter and depth. Pitfall
trap covers, constructed from pieces of plywood approximately 1.3 cm thick,
were used to exclude debris from the traps. The covers were approximately 20 cm
x 20 cm and were supported by two short blocks of 2.5-cm-thick wood that had
been screwed to the trap cover surface on opposing corners using wood screws.
Pitfall traps were placed in holes that were dug as close as possible to the center
of the plot with the lip of the trap being level with the surrounding ground, with
one pitfall trap per plot.
Flying insects were sampled using commercially prepared yellow sticky traps
(Sticky Strips Insect Traps, Olson Products) measuring 7.6 cm x 12.7 cm. Sticky
traps were hung from 1.3-cm-diameter PVC piping via a long binder clip (2.5
cm) that was attached to the PVC piping using 1.3-cm screws approximately
2.5 cm from the top of the PVC piping. The PVC piping was hammered into the
ground to a depth of 0.30 m, leaving approximately 1 m of piping exposed. The
sticky trap apparatus was placed approximately 2.5 cm to the right (if standing
with your back to the access road) of the pitfall trap in every sampling plot, and
the yellow sticky trap was oriented in the same direction (i.e., towards the access
road) on every trap.
Sweep netting was conducted by starting in one corner of the experimental
plot and walking diagonally through the center of the plot to the opposing corner
126 Southeastern Naturalist Vol. 12, No. 1
once. A fine mesh net was swept in a horizontal motion at chest height while
walking through the plot so as to capture any arthropods that might be resting
upon vegetation. Sweeps were made in each plot, and the contents were emptied
into vials (large insects such as butterflies were removed prior to collection in
vials). All samples were returned to the lab for identification.
In addition to mean species richness, diversity of treatment and control
plots were assessed using the Shannon index of diversity (H'). (See a complete
species list in Appendix 1). Furthermore, a community similarity value
for each pair-wise comparison of the different treatment groups and sites was
measured using a Bray-Curtis similarity matrix. Due to the fact that a large
amount of data was generated by the Bray-Curtis similarity matrix, the mean
similarity value of each cross treatment/site comparison (Table 1) was used
to construct a non-metric MDS ordination plot, so as to provide a concise 2-D
representation of the community assemblage groupings. To analyze for differences
between treatment groups in terms of diversity, a two-way ANOVA
was utilized using both treatment and site as the fixed factors. To meet
homogeneity assumptions of ANOVA, diversity data for the spring of 2008
was ℮(n)-transformed for the Shannon index. ANOVA was conducted using
SPSS v.15, and the Bray-Curtis and the NMDS ordination analyses were conducted
using PRIMER v. 5.
For both sites and survey years, mean species richness was not significantly
higher in the experimental plots compared to control plots. For each site,
mean species richness in cleared and thinned plots were 36% and 15% higher
Table 1. Average Bray-Curtis index values for cross treatment comparisons at McGirt’s Creek and
Tiger Point, Spring 2008. Con = Control, Clr = Clear, Thin = Thinned, Tig = Tiger Point, and McG =
Sites Con McG Clr McG Thin McG Con Tig Clr Tig
Clr McG 46.236560
Thin McG 45.317220 49.09561
Con Tig 22.477060 28.66109 24.930130
Clr Tig 15.436490 21.21960 14.075170 41.84704
Thin Tig 8.826945 13.84275 9.974205 38.67925 69.50629
Clr McG 9.523809524
Thin McG 28.571428570 10.344827590
Con Tig 5.504587156 21.656050960 10.19108280
Clr Tig 18.487394960 7.185628743 16.76646707 0.0
Thin Tig 14.516129030 5.813953488 11.62790698 6.451612903 6.060606061
2013 J.C. Burkhalter, D.C. Moon, and A.M. Rossi 127
than control plots, respectively (Figs. 1, 2) (F1,24 = 0.062; P α=0.05 = 0.807 and
F1,24= 2.379; P α=0.05 = 0.140, 2008 and 2009, respectively). Additionally, there
was no apparent effect on species richness due to treatments, although both
Figure 1. Mean species richness (± 1 SE) across all treatment groups at McGirt’s Creek,
Spring 2008 and 2009.
Figure 2. Mean species richness (±1 SE) across all treatment groups at Tiger Point, Spring
2008 and 2009.
128 Southeastern Naturalist Vol. 12, No. 1
cleared and thinned plots equaled or exceeded the number of species in control
plots after two years; once again these differences were not statistically
significant (F2,24 = 0.017 and Pα =0.05 = 0.983, and F2,24, = 0.884 and Pα =0.05 =
Figure 3. Mean Shannon diversity value (±1 SE) across all treatment groups at McGirt’s
Creek, Spring 2008 and 2009.
Figure 4. Mean Shannon diversity value (±1 SE) across all treatment groups at Tiger
Creek, Spring 2008 and 2009.
2013 J.C. Burkhalter, D.C. Moon, and A.M. Rossi 129
0.430 for 2008 and 2009, respectively; for a complete species list for both
years, see Appendix 1).
Richness and diversity trends for spring 2009 were similar to data from spring
2008, and treatment produced no significant difference compared to controls
(F2,24 = 0.174; P α=0.05 = 0.842 and F2,24,= 1.162; P α=0.05 = 0.335 for 2008 and 2009,
respectively). Lastly, species diversity (H') was initially approximately 30%
higher at McGirt’s Creek in 2008 than in comparison to 2009, a difference that
was statistically significant (F1,24 = 4.668; Pα =0.05 = 0.044); this trend was not the
same for Tiger Point, with the difference not reaching significance (F2,24 = 3.358
Pα =0.05 = 0.083) (Figs. 3, 4). No significant interaction was found between site and
treatment for any dependent variable for either year.
The mean similarity value of each cross treatment/site comparison and the
resulting ordination plot revealed that there were two main groupings separated
by site, with subsequent subgroupings, consisting of the cleared and thinned
plots grouping closer together in 2008 (Table 1; Figs. 5, 6). This high degree of
similarity between the cleared plots and the thinned plots was to be expected due
to the similar vegetation structure in these two experimental treatment groups in
comparison to the control replicates two years after the study began. These mean
similarity values were fairly disparate, even within sites (Figs. 5, 6).
Figure 5. Non-metric MDS ordination plot of community assemblages, Spring 2008.
M = McGirt’s Creek, T = Tiger Point.
130 Southeastern Naturalist Vol. 12, No. 1
In the current study, arthropod diversity was not significantly affected by
restoration treatments. However, both treatment groups (thinned and cleared),
which experienced biodiversity reduction through experimental treatments,
exceeded some measures of biodiversity (i.e., the Shannon index) compared to
control plots by the second year of the study at both sites. These results suggest
that insufficient time has passed to assess long-term (greater than 5 years) effects
of the treatments on arthropod biodiversity or that they have not yet reached an
equilibrium. Several reasons may account for differences in biodiversity between
treatment groups; certainly one possible reason is that, thus far, only short-term
responses have been assessed. Experimental manipulations were undertaken in
2006, and data collection began approximately 2 years later; it is possible that
an insufficient period of time elapsed to allow arthropod communities to reach
their maximum diversity. A series of studies concerned primarily with arthropod
community restoration have shown that recovery may take longer than expected.
For instance, a study investigating terrestrial arthropod recolonization following
riparian restoration found that abundance of all arthropods was lower at the
restored sites in comparison to reference sites, but all sampling was done within
two years of restoration, much like our study, in which there may not have been
enough time to see appreciable effects (Williams 2006). Blake et al. (1996)
found that five years after restoration, the carabid beetle fauna associated with
Figure 6. Non-metric MDS ordination plot of community assemblages, Spring 2009.
M = McGirt’s Creek, T = Tiger Point.
2013 J.C. Burkhalter, D.C. Moon, and A.M. Rossi 131
wildflower meadows exhibited fewer species and lower diversity than undisturbed
sites. Additionally, arthropod communities following mine reclamation
exhibited lower species richness, diversity, and evenness up to six years following
revegetation in comparison to control plots (Parmenter and MacMahon
1987, Parmenter et al. 1991). More recently, moth species richness following
prairie restoration was found to be lower with the newer restorations (1–3 years)
compared to older restorations (7–10 years) and control prairie remnants (Summerville
et al. 2007). While trends in overall species richness and diversity in
cleared and thinned plots is similar to the control plots at each site, community
similarity between sites is relatively low. These results seem to suggest that the
fixed size of the plots used in the study support a similar equilibrium species
richness (due to simple species-area relationships, similar number of niches,
etc.)—thus, plots may have reached an equilibrium for local diversity that varies
slightly between our sites, or the plots may contain subsets of the larger overall
regional diversity. However, these studies, as with the results reported here,
demonstrate that increased diversity or species richness are not likely to be an immediate
result of restoration. In a similar study, Rossi et al. (2011) measured the
plant community response to the restoration techniques used at McGirt’s Creek
and Tiger Point. They demonstrated that plant species diversity did increase as a
result of restoration at McGirt’s Creek, but the increase was recent; thus, too little
time may have elapsed for vegetation to fully affect the arthropod community.
Although treatments did not differ in species diversity within our study, they
did differ in terms of community similarity. This result raises the question of
whether or not community similarity is a useful, if not better, measure of restoration
success. Communities are considered important biological entities in
their own right, and conserving representative samples of communities is often
viewed as an efficient way of maintaining high levels of diversity (Hunter et al.
1988). Conservation of communities is considered to be a coarse-filter approach
as opposed to the fine-filter approach that attempts to conserve a single species
(Hunter et al. 1988). In the context of coarse-filter conservation, community
similarity measures may be more appropriate than measures of species richness
(Su et al. 2004). For instance, the restored treatments represent different successional
stages of the ecosystem and they support communities containing different
assemblages of species and thus contain a greater proportion of the species pool
for the area. In this study, different treatment groups support different arthropod
species assemblages as measured by the various community similarity indices,
and by taking this into consideration, it might be appropriate to implement multiple
restoration techniques so as to conserve the greatest species reservoir at the
two experimental sites.
Conversely, similarity will shed some insight into species identity and how
it compares and relates to ecosystem function in other communities. If one community
contains key species, such as keystone species or ecosystem engineers,
similarity measures can allow researchers to compare its abundance and proportion
to other communities and then provide researchers possible insights into
what may drive the presence or absence of such key species.
132 Southeastern Naturalist Vol. 12, No. 1
The authors would like to thank the following organizations for providing funding for
the project: The Nature Conservancy, United States Fish and Wildlife Service, Florida
Department of Environmental Protection, and Preservation North Florida. Additional
support was provided by the University of North Florida. We also thank Dale Casamatta,
Courtney Hackney, Julie Lockwood, and especially Keith Stokes, whose comments improved
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Appendix 1. Complete morphospecies list for both sites, 2008 and 2009.
Gryllus carolina Johansson Chaoboridae Sp. 1 Eurytomidae Sp. 1 Lycosid Sp. 9 Reduviidae Sp. 2
Gyrllus fasciata Haan Chrysomelidae Sp. 1 Flatidae Sp. 1 Lyssomanes viridesWakkenaer Rhapidophoridae Sp. 1
Acrididae Sp.1 Chrysomelidae Sp. 2 Forminicae Sp. 1 Mantidae Sp. 1 Salticidae Sp. 1
Acrididae Sp.3 Cicadellidae Sp. 1 Gastercantha cancriformis (L.) Meloidae Sp. 1 Salticidae Sp. 2
Actinae Sp.1 Cicadellidae Sp. 2 Geometridae Sp. 1 Meloidae Sp. 2 Salticidae Sp. 3
Aedes Sp.1 Cicadellidae Sp. 3 Geometridae Sp. 2 Meloidae Sp. 3 Salticidae Sp. 4
Agelenidae Sp.1 Cicadellidae Sp. 4 Geometridae Sp. 3 Mermeria bivittata Serville Salticidae Sp. 5
Alticinae Sp.1 Cicadellidae Sp. 5 Geometridae Sp. 4 Miridae Sp. 1 Salticidae Sp. 6
Anobiidae Sp.1 Cicadellidae Sp. 6 Geometridae Sp. 5 Miridae Sp. 2 Salticidae Sp. 7
Anopheles Sp.1 Cicadellidae Sp. 7 Gnaphosid Sp. 1 Miridae Sp. 3 Salticidae Sp. 8
Anthocoridae Sp.1 Cicujidae Sp. 1 Gnaphosid Sp. 2 Miridae Sp. 4 Salticidae Sp. 9
Anthribidae Sp.1 Cisseps fulvicollis (Hübner) Gnaphosid Sp. 3 Mogoplistidae Sp. 1 Scaphididae Sp. 1
Anthritidae Sp.1 Cleridae Sp. 1 Gnaphosid Sp. 4 Mordellidae Sp. 1 Scarabidae Sp. 1
Aphididae Sp.1 Cocinellidae Sp. 1 Graphocephala coccinea Forster Muscid Sp. 1 Scarabidae Sp. 2
Aphididae Sp.2 Coelopidae Sp. 1 Griburius larvatus (Newman) Muttlilidae Sp. 1 Scatopsidae Sp. 1
Aphididae Sp.3 Coenagrionidae Sp. 1 Gryllidae Sp. 1 Mycetophilidae Sp. 1 Sciaridae Sp. 1
Aradidae Sp.1 Coenagrionidae Sp. 2 Gryllidae Sp. 2 Mymaridae Sp. 1 Sciaridae Sp. 2
Araenidae Sp. 2 Coleoptera Sp. 1 Gryllus Sp. 1 Mymircinae Sp. 1 Sciomyzidae Sp. 1
Araenidae Sp.1 Colias eurytheme Boisduval Hemisphaerota cyanea (Say) Myridae Sp. 1 Sepsidae Sp. 1
Araenidae Sp.3 Colias philodice Godart Ichneumonidae Sp. 1 Myridae Sp. 2 Simullidae Sp. 1
Araenidae Sp.4 Collembola Sp. 1 Ichneumonidae Sp. 3 Myridae Sp. 3 Solenopsis invicta Buren
Araenidae Sp.5 Collembola Sp. 2 Ixodidae Sp. 1 Myrmicinae Sp. 2 Specidae Sp. 1
Araenidae Sp.6 Colydidae Sp. 1 Ixodidae Sp. 2 Nocturridae Sp. 1 Staphylinidae Sp. 1
Asilidae Sp.1 Crematogaster Sp. 1 Largus succintus L. Ochertidae Sp. 1 Staphylinidae Sp. 2
Bibionidae Sp.1 Crytocephalus Sp. 1 Largidae Sp. 1 Oedemeridae Sp. 1 Stratiomyidae Sp. 1
Blattidae Sp.1 Culicidae Sp. 1 Largidae Sp. 2 Oniscidae Sp. 1 Symphyta Sp. 1
Blattidae Sp.2 Curculionidae Sp. 1 Lauxaniidae Sp. 1 Oniscus asellus L. Tabanidae Sp. 1
Buprestidae Sp.1 Curculionidae Sp. 2 Lepidoptera Sp. 1 Opiliones Sp. 1 Tachinidae Sp. 1
Buprestidae Sp.2 Curculionidae Sp. 3 Leptura Sp. 1 Orchesella hexfasciata Harvey Temnothorax Sp. 1
Buprestidae Sp.3 Curculionidae Sp. 4 Lespidae Sp. 1 Orgyia leucostigma Smith Tenebrionidae Sp. 1
Buprestidae Sp.4 Curculionidae Sp. 5 Linphyiidae Sp. 1 Orthoptera Sp. 1 Tenebrionidae Sp. 2
Calliphoridae Sp.1 Curculionidae Sp. 6 Lycaenidae Sp. 1 Orthoptera Sp. 2 Tenebrionidae Sp. 3
136 Southeastern Naturalist Vol. 12, No. 1
Camponotus Sp.1 Delphacidae Sp. 1 Lycidae Sp. 1 Oxypidae Sp. 1 Tephritidae Sp. 1
Cantharidae Sp.1 Delphacidae Sp. 2 Lycosid punctulata Hentz Pardosa Sp. 1 Tetragnathidae Sp. 1
Carabidae Sp.1 Diptera Sp. 1 Lycosid Sp. 14 Pardosa Sp. 1 Tettigonidae Sp. 1
Carabidae Sp.2 Dolichopodidae Sp. 1 Lycosid Sp. 15 Pardosa Sp. 2 Theatops Sp. 1
Carabidae Sp.3 Dolichopodidae Sp. 2 Lycosid Sp. 16 Phaeothripidae Sp. 1 Theatops Sp. 2
Carabidae Sp.4 Dolichopodidae Sp. 2 Lycosid Sp. 1 Phaeothripidae Sp. 2 Thomisidae Sp. 1
Cassidinae Sp.1 Dolichopodidae Sp. 3 Lycosid Sp. 10 Phalacridae Sp. 1 Tipullidae Sp. 1
Catharidae Sp.1 Dolichopodidae Sp. 4 Lycosid Sp. 11 Philodromidae Sp. 1 Tortricidae Sp. 1
Cecidomyiidae Sp.1 Elateridae Sp. 1 Lycosid Sp. 12 Phoridae Sp. 1 Trochosa Sp. 1
Centruroides hentzi Banks Elateridae Sp. 2 Lycosid Sp. 13 Pipuncullidae Sp. 1 Trombididae Sp. 1
Cerambycidae Sp.3 Endomychidae Sp. 1 Lycosid Sp. 2 Pirata Sp. 1 Uca Sp. 1
Ceratopogonidae Sp.1 Epicauta floridensis Werner Lycosid Sp. 3 Pirata Sp. 2 Ulidiidae Sp. 1
Chaecididae Sp.2 Eucnemidae Sp. 1 Lycosid Sp. 4 Psyllobora Sp. 1 Zoraptidae Sp. 1
Chalcididae Sp.1 Eulophidae Sp. 1 Lycosid Sp. 7 Pteromalidae Sp. 1
Chaloididae Sp.1 Eurytides marcellus (Cramer) Lycosid Sp. 8 Reduviidae Sp. 1