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Influence of Vegetation on Invertebrate Communities in Grazed Freshwater Wetlands in South-central Florida
William R. Morrison III and Patrick J. Bohlen

Southeastern Naturalist, Volume 9, Issue 3 (2010): 453–464

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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. Introduction 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 - 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. Methods 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). Experimental design 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 per plot. Environmental Sampling 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 taxonomic unit. Data analysis 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 Results 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). Discussion 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. Environmental variables 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 southeastern US 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 wetlands. 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. Acknowledgments 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 Research Center. 2010 W.R. Morrison III and P.J. Bohlen 463 Literature Cited Ausden, M., M. Hall, and T. Strudwick. 2005. 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