2010 SOUTHEASTERN NATURALIST 9(3):453–464
Influence of Vegetation on Invertebrate Communities in
Grazed Freshwater Wetlands in South-central Florida
William R. Morrison III1 and Patrick J. Bohlen2,*
Abstract - Grazing lands and rangelands are increasingly recognized as an important
alternative to other developed land uses for sustaining ecological communities in
Florida, the rest of the southeastern United States, and other regions. It is important
to understand factors that influence ecological communities on private grazing lands,
especially in areas with abundant wetlands, which are often sensitive habitats. This
study examined the effects of different vegetation types and simulated grazing (clipping)
on the abundance, diversity, and composition of the invertebrate community
in seasonally flooded, isolated wetlands on a cattle ranch in south-central Florida.
We compared invertebrate communities in wetland areas dominated by two different
types of vegetation: emergent macrophytes (Pontederia cordata [Pickerelweed])
and grasses (primarily Luziola fluitans [Southern Watergrass]). There was a trend
toward greater abundance and diversity of invertebrates in grass-dominated communities.
Experimental removal of vegetation to simulate heavy grazing significantly
decreased the abundance and diversity of invertebrates. It also shifted the community
composition of invertebrates to favor members of Diptera and Ostracoda. Management
practices in grazed wetlands that use light or intermediate levels of grazing,
or that foster a greater diversity of vegetative cover, may support more diverse and
populous wetland invertebrate communities.
Human land use can have positive and negative implications for organism
biodiversity and ecosystem services. A common type of land
management regime in Florida is grazing cattle (Main et al. 2000). Various
groups have been promoting the environmental benefits of ranching and
the importance of ranches to sustaining biodiversity and ecosystem function,
especially in major cattle-production areas of south-central Florida
(Bohlen et al. 2009, Swain et al. 2007). However, data for ecological communities
on ranches is lacking, especially for non-economic groups such as
wetland invertebrates. More information on factors that affect ecological
communities in these managed landscapes is needed to inform management
practices for conservation.
It is important to consider effects of livestock grazing on wetland ecology,
especially in regions like south Florida where wetlands are an important
part of the grazing landscape. Evidence suggests that cattle grazing in wetlands
can inhibit the establishment of shrubby or tree species (Vulink et
al. 2000). Presence of cattle alters the plant community, as cattle consume
1Evolution, Ecology, and Systematics, Biologie Department II, University of Munich,
Planegg-Martinsried, Germany. 2MacArthur Agro-Ecology Research Center,
Lake Placid, fl. *Corresponding author - firstname.lastname@example.org.
454 Southeastern Naturalist Vol. 9, No. 3
certain species while avoiding others (Ausden et al. 2005). The overall effect
of this selective grazing can either be positive or negative, depending on the
circumstances and desired outcomes (Marty 2005, Vavra 2005).
The impact of cattle on vegetation in wetlands raises the question of how
grazing influences other wetland trophic levels, such as aquatic macroinvertebrates,
which are integral to wetland foodwebs. Previous studies suggest
links between cattle grazing and invertebrates in various types of freshwater
habitats. For example, in prairie potholes located in south-central Canada,
removal of emergent vegetation by cattle grazing resulted in decreased
abundance and lower reproductive efforts of Odonata (Foote and Hornung
2005). In fens, grazing has been shown to significantly decrease both the
abundance and species richness of Mollusca (Ausden et al. 2005). In ditches
in the lowlands of the United Kingdom, invertebrate distributions have been
correlated with macrophyte occurrence (Painter 1999). On a cattle ranch
in south-central Florida, Steinman et al. (2003) reported that differences in
vegetation among wetlands had an effect on wetland invertebrate richness
and diversity. Wetlands dominated by the grass Panicum hemitomon J.A.
Schultes (Maidencane) had greater richness and diversity of insects than
wetlands dominated by more diverse mixtures of emergent macrophytes and
weedy wetland plants species. There were no significant differences in invertebrate
richness or diversity in wetlands exposed to grazing versus those
not exposed to grazing.
We investigated whether patches of different types of vegetation within
a wetland support different benthic and epibenthic invertebrate communities.
Further, to explore effects of grazing on invertebrate communities, we
removed vegetation from small plots within wetlands to simulate vegetation
removal by cattle. The study sites were located on a cattle ranch that
was used in a previous study examining effects of pasture type and grazing
on wetland invertebrate communities in whole wetlands (Steinman
et al. 2003). Our focus was to assess the influence of different patches of
vegetation and vegetation removal on variation of wetland invertebrate
communities within wetlands.
Study site and selection of wetlands
This study was performed in July of 2006 at the MacArthur Agro-
Ecology Research Center (Highlands County, fl; 27°08'57.42"N,
81°11'28.16"W), a 4250-ha working cattle ranch with over 600 embedded
depressional freshwater marshes, most of which are isolated, except in
cases where they are intersected by drainage ditches. Wetlands used in this
study were selected on the basis of whether approximately half of the wetland
was composed of Pontederia cordata L. (Pickerelweed)-dominated
communities and half of grass-dominated communities. Also, those wetlands
were chosen that were neither exceedingly large nor very small (>0.4
ha and <2.5 ha). A total of 10 wetlands were selected to fit these criteria.
2010 W.R. Morrison III and P.J. Bohlen 455
All sites were situated in semi-native pastures that have never been fertilized
and are mainly used for winter grazing of cow-calf pairs. These pastures
consist of varying mixtures of introduced forage species (mainly Paspalum
notatum Flueggé [Bahiagrass]), native tall grasses, such as Andropogon
virginicus L. (Broomsedge or Yellow Bluestem) and other Andropogon spp.,
Maidencane, and Panicum longifolium Vasey (Redtop Panicgrass), as well
as a diversity of other native shrubs and forbs (Swain et al. 2007).
The experiment consisted of a 2 x 2 replicated design in which invertebrate
communities were sampled in 2 different types of vegetation
community (Pickerelweed or grass-dominated) with vegetation left intact
or clipped to the base (to simulate cattle grazing; see: Butler et al.
2008, Multikainen et al. 1993). Due to logistical and time constraints
for this project, it was not possible to manipulate cattle for the grazing
treatment. Furthermore, the purpose of the treatment was to simulate
only removal of vegetation, not the other potential impacts of cattle, in
particular trampling of wetland sediments. The Pickerelweed-dominated
sites were chosen by first going to the geographic center of the given wetland.
A list of random numbers was generated for headings and distances
from that point. After reaching the center, the first random heading with
a random distance was used, utilizing magnetic north as 0 degrees. If that
random distance and heading did not locate the site in a patch dominated
by Pickerelweed, a site was chosen further along the same heading. Once
a Pickerel-dominated site was identified (>90% cover), a quadrat of vegetation
75- x 75-cm was clipped to the base of the stems where they meet
the substrate (referred to hereafter as PR for Pickerelweed removed). This
treatment was designed to simulate the effects of heavy grazing by cattle
in a wetland; the site was allowed to sit for one week before sampling to
allow the biota to equilibrate. The paired site for the Pickerelweed location
with vegetation that was not removed (Pickerelweed present; PP) was
chosen randomly within a radius of 5 m from the first site.
From the first randomly selected site, another random distance and
heading was chosen to select a grass-dominated site. The grass sites tended
to be located more on the periphery of the wetland, so headings that
led toward the deeper water in the center of the wetlands were eliminated.
Once the grass-dominated site was reached, a 75- x 75-cm area of vegetation
(GR) was clipped at the base of the plants, removing the clipped
vegetation from the wetland. A paired site of undisturbed grass vegetation
(GP) was chosen randomly within a radius of 5 m from the plot where
grass was clipped and removed.
Collection of invertebrates
Benthic and epibenthic invertebrates were chosen for this study as
these groups appear to be the most sensitive to variations in the plant
community (Steinman et al. 2003). Funnel traps were used to sample small
456 Southeastern Naturalist Vol. 9, No. 3
benthic invertebrates (Whiteside and Lindegaard 1980). Each funnel trap
consisted of nine inverted funnels held in place by a 40- x 40-cm piece of
0.64-cm-thick acrylic sheet. The sheet had a 3 x 3 grid of equally spaced
holes through which the funnel stems were inserted to hold them in place.
Plastic sample bottles (0.24 L) were attached to the protruding funnel
stems by drilling small holes in the bottle caps and pushing the capped
bottles down onto the stem. The resulting trap had inverted funnels on
one side of the plastic sheet and collection bottles on the opposite side.
Four funnel traps were set out in each wetland, one in each of the four
plots (PR, PP, GR, GP). The traps were deployed by filling the collection
bottles with water from a given sampling site and placing the trap funnelside
down on top of the wetland sediments. After clipping, we waited for 7
days before sampling to provide time for invertebrate communities in the
clipped areas to equilibrate. After this period, the traps were deployed in
the wetland for 24 hours. In the lab, the contents of the nine sample collection
bottles from each trap were poured through a 0.5-mm mesh, and
the retained invertebrates specimens were combined into a single sample
preserved in 70% ethanol.
In addition to the funnel traps, GP and PP plots were sampled by sweeping
a D-frame sweep net a distance of 1 m approximately 1m away from
the plots at the time that the funnel traps were set out. The direction for the
sweeping was done randomly based off of generated numbers for headings.
The contents of the net were placed in a whirl pack, and preserved in a 1:1
dilution of sample and 70% ethanol. Two sweep net samples were collected
Water depth, temperature, pH, and conductivity were measured at each
sampling site at the time that the traps were retrieved. These environmental
descriptors have been shown in other studies to affect the invertebrate community
(Batzer et al. 2004, Kay et al. 2001). Water depth was measured with
a meter stick, whereas temperature, pH, and conductivity were measured
with a YSI 556 Multiparameter Probe (YSI, Inc., Yellow Spring, OH).
Total phosphorus, ortho-phosphorus, and ammonia levels were analyzed
in each of the characterized wetlands, as these water quality measures have
been shown to affect organisms sensitive to their concentrations (Steinman
et al. 2003). Three water samples were taken from each wetland: one from
the geographic center of the wetland, and one at each of two 7-m intervals
from the center to the edge of the wetland. Samples collected for ortho-P
analysis were filtered immediately upon returning to the laboratory, stored
at 4 °C, and analyzed within 48 hours. The malachite green method was used
to analyze the samples using a microplate method (D’Angelo et al. 2001).
Those samples that were to be analyzed for the total phosphorus and ammonia
were preserved with concentrated H2SO4, stored at 4 °C, and analyzed
within 10 days. Total P was analyzed using a persulfate digestion followed
by the ascorbic acid method (Pote and Daniel 2000), and ammonium was an2010
W.R. Morrison III and P.J. Bohlen 457
alyzed using a modified ascorbic acid method (Sims et al. 1995). All samples
were analyzed in a microplate spectrophotometer (μQuant Microplate Spectrophotometer,
Bio-Tek Instruments, Winooski, VT).
Identification of invertebrates
Using a dissecting microscope, specimens from sweep nets and funnel
traps were assigned to a morphospecies based on physical differences and
counted. A representative specimen for each morphospecies was retained
in a scintillation vial filled with 70% ethanol as a reference to ensure that
morphospecies were not duplicated and for later identification. During this
first round of sorting, invertebrates were assigned taxonomically to order
using Voshell (2002) as a guide. In a second round of sorting, invertebrates
were assigned to family, where possible, using the same guide. Those organisms
whose family could not be identified were left at the lowest identifiable
All data were log or square root transformed as needed to conform to a
normal distribution. Subsequently, Shapiro-Wilk tests were performed to
ensure normality (McCune and Grace 2002). Analyses were performed using
one dataset comprising 20 sweep samples (GP and PP plots only) and 40
funnel samples (all plot types: GP, GR, PP, and PR). To evaluate treatment
effects on abundance and diversity, summary and diversity indices were
generated. Indices were calculated for every quadrat (e.g., site) and included
morphospecies richness, abundance, evenness, and Simpson’s and Shannon’s
indices (using morphospecies as a surrogate for species in these last
two and the program PC-ORD). A 2-way ANOVA was used to analyze these
measures, using vegetation type (grass or Pickerelweed) and occurrence of
vegetation (clipped/non-clipped) as explanatory variables. Community composition
of invertebrates was analyzed by producing Bray-Curtis similarities
between the sampled sites and visualizing them using nonmetric multidimensional
scaling (NMDS). An analysis of similarity (ANOSIM) was used
to evaluate whether the invertebrate community composition significantly
differed based on treatment, again using Bray-Curtis similarities. The ordination
and ANOSIM algorithm was performed with PAST v1.76 (Hammer
et al. 2001), and SPSS v11.5 was used to carry out the 2-way ANOVA procedures.
Graphs were generated using SigmaPlot v10.0.
To assess confounding effects from environmental variables among
wetlands on diversity indices and abundance, Spearman's rank correlation
coefficients were generated using the data from the 40 funnel traps. Spearman's
coefficients were produced for comparisons between depth, pH,
conductance, and temperature with the abundance, morphospecies richness,
Shannon's index, and Simpson's index at each site. Because of the large
amount of tests on these environmental data, a Bonferroni correction was
used with alpha = 0.0025. For all other analyses, alpha = 0.05.
458 Southeastern Naturalist Vol. 9, No. 3
Effects of vegetation community type on invertebrates
Grass-dominated plots appeared to have greater abundance and higher
diversity of wetland invertebrates than Pickerelweed-dominated plots,
although the trends were not statistically significant at an alpha = 0.05
(2-way ANOVA, means ± 95% confidence, sqrt [abundance]: Pontmean =
17.83 ± 5.5, Grassmean = 24.80 ± 6.50, F1,10 = 3.46, P = 0.068; Shannon’s index:
Pontmean = 1.31 ± 0.20, Grassmean = 1.57 ± 0.19, F1,10 = 3.92, P = 0.053;
sqrt [Simpson’s Index]: Pontmean = 0.62 ± 0.07, Grassmean = 0.70 ± 0.06,
F1,10 = 3.56, P = 0.064).
Invertebrate community responses to vegetation removal
Mean Simpson’s index, Shannon’s index, abundance, and diversity of
wetland invertebrates were significantly lower in plots where vegetation
was removed than where it was left intact (2-way ANOVA, sqrt [Simpson’s
index]: F1,10 = 10.90, P < 0.002; Shannon’s index: F1,10 = 13.52, P <
0.0006); sqrt [abundance]: F1,10 = 4.35, P < 0.042; diversity: F1,10 = 13.41,
P < 0.0006; Fig. 1).
Invertebrate communities differentiated along axis 1 of the NMS ordination
(Fig. 2); those samples where vegetation was absent tended to have
higher ordination scores along axis 1, while those where vegetation was
Figure 1. Difference in mean invertebrate abundance and diversity indices between
sites where vegetation was present (PP, GP) or clipped (PR, GR) to simulate grazing.
Error bars represent 95% confidence intervals. All measures are significantly different
between the simulated grazing and unclipped vegetation using a 2-way ANOVA
(P < 0.05).
2010 W.R. Morrison III and P.J. Bohlen 459
present tended to have lower scores (ANOSIM, 1000 permutations: R =
0.0881, P = 0.056). Moreover, there was more dispersion and greater variability
in the composition of the invertebrate community where vegetation
was present as opposed to where it was clipped as indicated by the tighter
clustering of values from the clipped areas (Fig. 2).
There were differences in the abundances of individuals in different orders
based on whether vegetation was present or removed (Table 1). Though
the most abundant three orders were the same in both treatments, Coleopterans
were second most abundant (23% of total abundance) when vegetation
was present, but third in abundance (9% of total abundance) after Ostracoda
when vegetation was absent. In both cases, Diptera was the most abundant
order composing 38% and 46% of the total individuals when vegetation
was present and absent, respectively. Taken together, Diptera and Ostracoda
comprised more than 75% of the total individuals found in samples where
vegetation was removed. Only seven orders were found when vegetation was
Figure 2. Optimal 2-D NMDS scatterplot of community data based on Bray-Curtis
similarities (stress: 17.3, instability: 0.00378) and categorized by clipped vegetation
(X’s) and unclipped vegetation (squares). Convex hulls were drawn based on 95%
confidence intervals for clipped and unclipped sites for better visualization.
460 Southeastern Naturalist Vol. 9, No. 3
clipped relative to 12 orders when the vegetation was present. Collembola,
Homoptera, Lepidoptera, Ephemeroptera, and Megaloptera were not found
when vegetation was clipped.
There were also differences in the percentage of morphospecies that
belonged to each order (Table 1). Diptera was the most diverse group
in morphospecies and number of families present regardless of whether
vegetation remained or was clipped; however, Dipteran morphospecies
increased from 28% of the total invertebrate diversity to 43% when
vegetation was removed. Coleoptera was the second-most diverse group
(23%) when vegetation was present, whereas Ostracoda (18%) was the
second-most diverse when vegetation was removed. When vegetation
was clipped, the rank of orders from least abundant to most abundant
followed exactly the rank of orders from least diverse to most diverse.
However, this relationship did not hold true when vegetation was present.
The water depth (range = 15.50–40.75 cm; mean = 28.45 ± 10.41 cm;
F9,40 = 3.87, P < 0.005), conductance (range = 26.5–144.3 μs; mean = 68.8
± 30.4 μs; F9,40 = 65.13, P < 0.001), amount of ortho-phosphorus (range =
0.024–0.052 mg/L; mean = 0.040 ± 0.008 mg/L; F9,30 = 98.66, P < 0.001),
and amount of total available biological phosphorus (range = 0.027–0.135
mg/L; mean = 0.058 ± 0.028 mg/L; F9,30 = 77.58, P < 0.001) were significantly different among wetlands. However, none of these environmental
variables were significantly correlated with the measures for the various
wetland invertebrate community (Table 2).
Effect of vegetation on invertebrate communities within wetlands
Trends observed in our study are consistent with previous findings that
grass-dominated plant communities appear to support a greater abundance
and diversity of wetland invertebrates when compared to areas dominated
Table 1. Order-level data from 40 funnel-trap samples expressed as a percent of the total
abundance of individuals for samples where vegetation was unclipped (sites PP, GP; n = 332),
where vegetation was clipped (sites PA, GA; n = 120); or expressed as a percent of the total
morphospecies richness (% diversity) described within orders for samples where vegetation was
unclipped (sites PP, GP; n = 332) and where vegetation was clipped (sites PA, GA; n = 120).
% abundance % diversity
Taxon Unclipped Clipped Unclipped Clipped
Ephemeroptera 1 0 3 0
Megaloptera 1 0 2 0
Odonata 1 1 8 3
Hemiptera 2 3 10 3
Trichoptera 7 8 12 12
Gastropoda 7 3 8 9
Ostrocoda 20 30 6 18
Coleoptera 23 9 23 12
Diptera 38 46 28 43
2010 W.R. Morrison III and P.J. Bohlen 461
by emergent macrophytes in these seasonally inundated freshwater marshes
(Steinman et al. 2003). Previous research has reported positive correlations
between plant structural diversity and invertebrate diversity (Batzer et al.
2004, Harvey et al. 2008), supporting the idea that diversity arises in part
from structural parameters such as landscape heterogeneity (Duelli 1997).
Grass communities have a higher density of stems than the less dense
patches of larger-stemmed emergent macrophytes, which may provide more
refugia and cover for small benthic and epibenthic invertebrates.
Influence of vegetation removal on invertebrate communities
Clipping vegetation was designed to mimic heavy grazing by cattle in a
wetland. Clipping treatment had a strong negative impact on the abundance
and diversity of wetland invertebrates, confirming the importance of vegetation
as habitat for benthic invertebrates in these systems. Moreover,
vegetation removal had different effects on different invertebrate groups,
with some orders, such as Diptera, being more abundant and contributing
more to overall diversity in clipped areas than in areas with intact vegetation.
Invertebrates were relatively depauperate in clipped sites probably because
of a decrease in the availability of refugia, food, and living spaces (Beckett
et al. 1992).
Though this experiment simulated vegetation removal by grazing cattle,
it did not simulate the trampling that accompanies such grazing, so we are
unable to make conclusions about potential effects on benthic invertebrates
of the large physical disturbance created by trampling. Responses observed
in this study were for a single instance of relatively small-scale vegetation
removal, and the results might have been more pronounced or different if
larger areas had been denuded, if the vegetation had been removed repeatedly
Table 2. Spearman’s rank correlations between the environmental variables (water depth, pH,
conductance, temperature, Ortho-P, and Total P) and the measures for the invertebrate communities
from 10 wetlands on a cattle ranch.
Water depth pH Conductance
Invertebrate measures n ρ P ρ P ρ P
Morphospecies richness 40 -0.103 0.528 0.009 0.956 0.102 0.528
Abundance 40 -0.101 0.531 -0.125 0.441 -0.160 0.325
Evenness 40 0.048 0.767 0.150 0.355 0.298 0.062
Simpson’s index 40 -0.056 0.730 0.177 0.275 0.369 0.192
Shannon index 40 -0.102 0.530 0.156 0.336 0.331 0.369
Temperature Ortho-P Total P
Invertebrate measures n ρ P ρ P ρ P
Morphospecies richness 40 0.115 0.479 -0.083 -0.609 -0.117 0.471
Abundance 40 0.073 0.654 -0.061 0.709 -0.025 0.877
Evenness 40 0.173 0.284 0.177 0.276 0.083 0.609
Simpson’s index 40 0.268 0.095 0.150 0.354 0.059 0.719
Shannon index 40 0.218 0.176 0.06 0.711 -0.011 0.944
462 Southeastern Naturalist Vol. 9, No. 3
over the course of the growing season, or if a surrogate to trampling had been
included (Kruess and Tscharntke 2002). For these reasons, the results obtained
from our vegetation removal treatment are likely to be a conservative
estimate of grazing impact on wetland invertebrate communities.
Implications for conservation of wetlands on cattle ranches in the
Previously, research has shown that small, isolated wetlands are important
sources of biodiversity (Kirkman et al. 1999). However, wetlands with
different vegetation and disturbance regimes will vary in the diversity of
aquatic invertebrates. Namely, wetlands with more submerged vegetation,
greater grass cover, less disturbance, and fewer open patches may have
greater abundance and diversity of invertebrates, which in turn could provide
more food resources for higher predatory trophic levels, including
birds, fish, and other vertebrates (Joyner 1980, Oliver et al. 1998).
Cattle often trample as well as consume varying amounts of vegetation
in a wetland, which creates open spaces devoid of vegetation. Cattle
also can change competitive dynamics between plant species leading to
shifts in wetland plant communities (Blanch and Brock 1994, Bohlen and
Gathumbi 2007, Vulink et al. 2000). Previous research has found that only
those invertebrates that are most tolerant to disturbance can withstand the
conditions created by cattle activity (Palmer et al. 2000). Although direct
localized effects of cattle may have negative consequences for invertebrate
communities, larger-scale effects for the whole wetland might include
shifts in plant communities and increased heterogeneity that could increase
invertebrate diversity. Assessing effects at this larger scale would require
longer-term studies in which cattle grazing were manipulated in whole replicated
Managing grazing lands to preserve diversity should include consideration
of the intensity, timing, and duration of grazing to mitigate effects
of cattle on wetland communities. Intermediate levels of grazing intensity
might ameliorate negative consequences of persistent grazing pressure and
could increase diversity of certain taxa by increasing environmental heterogeneity
(Marty 2005). Considering the large extent of grazing lands in
some regions of the southeastern US and their overlap with wetlands of
conservation value, more studies are needed of ecological communities and
interactions in these systems.
Thanks to A. Weiler, A. Tweel, and A. Peterson for their help in the field and
throughout the process; to E. Fraser, who helped with collecting the water quality
data. Thanks also to P. Duchen, C. Rupprecht, and M. Wittmann for reviewing the
manuscript, as well as the two anonymous reviewers for their helpful suggestions and
constructive criticism. This paper is contribution No. 127 of the MacArthur Agroecology
2010 W.R. Morrison III and P.J. Bohlen 463
Ausden, M., M. Hall, and T. Strudwick. 2005. The effects of cattle grazing on tallherb
fen vegetation and molluscs. Biological Conservation 122:317–326.
Batzer, D.P., B.J. Palik, and R. Beuch. 2004. Relationships between environmental
characteristics and macroinvertebrate communities in seasonal woodland ponds
of Minnesota. Journal of the North American Benthological Society 23:50–68.
Beckett, D.C., T.P. Aartila, and A.C. Miller. 1992. Contrasts in densities of benthic
invertebrates between macrophyte beds and open littoral patches in Eau Galle
Lake, Wisconsin. American Midland Naturalist 127:77–90.
Blanch, S.J., and M.A. Brock. 1994. Effects of grazing and depth on two wetland
plant species. Australian Journal of Marine and Freshwater Research
Bohlen, P.J., and S.M. Gathumbi. 2007. Nitrogen cycling in seasonal wetlands in subtropical
cattle pastures. Soil Science Society of America Journal 71:1058–1065.
Bohlen, P.J., S. Lynch, L. Shabman, M. Clark, S. Shukla, and H. Swain. 2009. Paying
for ecosystem services on agricultural lands: An example from the northern
Everglades. Frontiers in Ecology and the Environment 7:46–55.
Butler, D.M., N.N. Ranells, D.H. Franklin, M.H. Poore, and J.J.T. Green. 2008.
Runoff water quality from manured riparian grasslands with contrasting drainage
and simulated grazing pressure. Agriculture, Ecosystems and Environment
D’Angelo, E., J. Crutchfield, and M. Vandiviere. 2001. Rapid sensitive microscale
determination of phosphate in water and soil. Journal of Environmental Quality
Duelli, P. 1997. Biodiversity evaluation in agricultural landscapes: An approach at
two different scales. Agriculture, Ecosystems, and Environment 62:81–91.
Foote, A.L., and C.L.R. Hornung. 2005. Odonates as biological indicators of grazing
effects on Canadian prairie wetlands. Ecological Entomology 30:273–283.
Hammer, O., D.A.T. Harper, and P.D. Ryan. 2001. PAST: Palaeontological statistics
package for education and data analysis. Palaeontologia Electronica 4:9.
Harvey, J.A., W.H. van der Putten, H. Turin, R. Wagenaar, and T.M. Bezemer. 2008.
Effects of changes in plant species richness and community traits on carabid
assemblages and feeding guilds. Agriculture, Ecosystems, and Environment
Joyner, D.E. 1980. Influence of invertebrates on pond selection by ducks in Ontario.
Journal of Wildlife Management 40:700–705.
Kay, W.R., S.A. Halse, M.D. Scanlon, and M.J. Smith. 2001. Distribution and environmental
tolerances of aquatic macroinvertebrate families in the agricultural
zone of southwestern Australia. Journal of the North American Benthological
Kirkman, L.K., S.W. Golladay, L. Laclaire, and R. Sutter. 1999. Biodiversity in
southeastern, seasonally ponded, isolated wetlands: Management and policy perspectives
for research and conservation. Journal of the North American Benthological
Kruess, A., and T. Tscharntke. 2002. Contrasting responses of plant and insect diversity
to variation in grazing intensity. Biological Conservation 106:293–302.
Main, M.B., M.E. Swisher, J. Mullahey, W. DeBusk, A.J. Shriar, G.W. Tanner, J.
Selph, P. Hogue, P.J. Bohlen, and G.M. Allen. 2000. The ecology and economics
of Florida’s ranches. University of Florida IFAS Extension, Gainsville, fl.
464 Southeastern Naturalist Vol. 9, No. 3
Marty, J.T. 2005. Effects of cattle grazing on diversity in ephermeral wetlands. Conservation
McCune, B., and J.B. Grace. 2002. Analysis of Ecological Communities. MjM Software
Design, Gleneden Beach, OR. 304 pp.
Multikainen, P., M. Walls, and A. Ojala. 1993. Effects of simulated herbivory on
tillering and reproduction in an annual ryegrass, Lolium remotum. Oecologia
Oliver, I., A.J. Beattie, and A. York. 1998. Spatial fidelity of plant, vertebrate, and
invertebrate assemblages in multiple-use forest in Eastern Australia. Conservation
Painter, D. 1999. Macroinvertebrate distributions and the conservation value of
aquatic Coleoptera, Mollusca, and Odonata in the ditches of traditionally managed
and grazing fen at Wicken Fen, UK. Journal of Applied Ecology 36:33–48.
Palmer, M.A., A.P. Covich, S. Lake, P. Biro, J.J. Brooks, J. Cole, C. Dahm, J. Gilbert,
W. Goedkoop, K. Martens, J. Verhoeven, and W.J. van der Bund. 2000. Linkages
between aquatic biotic sediment and life above sediments as potential drivers of
biodiversity and ecological processes. BioScience 50:1062–1075.
Pote, D.H., and T.C. Daniel. 2000. Analyzing for dissolved reactive phosphorus in
water samples. Pp. 91–93, In G.M. Pierzynski (Ed.). Methods of Phosphorus
Analysis for Soils, Sediments, Residuals, and Waters. North Carolina State University,
Sims, G.K., T.R. Ellsworth, and R.L. Mulvaney. 1995. Microscale determinations of
inorganic nitrogen in water and soil extracts. Communications in Soil Science
and Plant Analysis 26:303–319.
Steinman, A.D., J. Conklin, P.J. Bohlen, and D.G. Uzarski. 2003. Influence of cattle
grazing and pasture land use on macroinvertebrate communities in freshwater
wetlands. Wetlands 23:877–889.
Swain, H., P.J. Bohlen, K.L. Campbell, L.O. Lollis, and A.D. Steinman. 2007. Integrated
ecological and economic analysis of ranch management systems: An example
from south-central Florida. Rangeland Ecological Management 60:1–11.
Vavra, M. 2005. Livestock grazing and wildlife: Developing compatibilities. Rangeland
Ecology and Management 58:128–134.
Voshell, J.R., Jr. 2002. A Guide to Common Freshwater Invertebrates of North
America. McDonald and Woodward Publishing Co., Blacksburg, VA. 454 pp.
Vulink, J.T., H.J. Drost, and L. Jans. 2000. The influence of different grazing regimes
on phragmites and shrub vegetation in the well-drained zone of a eutrophic wetland.
Applied Vegetation Science 3:73–80.
Whiteside, M.C., and C. Lindegaard. 1980. Complementary procedures for sampling
small benthic invertebrates. Oikos 35:317–320.