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Amphibian and avian species composition of forested depressional wetlands and circumjacent habitat

Howard T. Odum Center for Wetlands
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PAGE 1

AMPHIBIAN AND AVIAN SPECIES COMPOSITION OF FORESTED DEPRESSIONAL WETLANDS AND CIRCUMJACENT HABITAT: THE INFLUENCE OF LAND USE TYPE AND INTENSITY By JAMES A. SURDICK JR. A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL OF THE UNIVERSITY OF FLOR IDA IN PARTIAL FULFILLMENT OF THE REQUIREMENTS FOR THE DEGREE OF DOCTOR OF PHILOSOPHY UNIVERSITY OF FLORIDA 2005

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Copyright 2005 by James A. Surdick Jr.

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iii ACKNOWLEDGMENTS I am not only grateful for the opportunity provided to me by my major advisor Dr. Mark Brown, but also for his catalytic appro ach to learning which allowed me to pursue my interests. I have gained high qual ity information, and an amazing experience exploring the systems of Florida. I th ank my committee memb ers Graeme Cumming, Richard Franz, and Clay Montague for their time, effort, and diligent input. I am forever thankful to Kris Sullivan for her tireless en ergy, curiosity, and support. Friends essential to this study include, but are not limite d to, Matt Cohen, Tony Davanzo, Chuck Lane, Kelly Reiss, and Ben Vivas. There are t oo many individual land owners, land managers, and facilitators to mention, but without their effort and consent this work would not have been possible. To the land owners who di d not allow me on to their property you now can see I did not have ulterior motives and re ally did just want to tally swamp critters. The Florida Department of Environmental Protection provided the funding to Dr. Brown and is the ultimate supporte r of this research.

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iv TABLE OF CONTENTS page ACKNOWLEDGMENTS.................................................................................................iii LIST OF TABLES............................................................................................................vii LIST OF FIGURES.............................................................................................................x ABSTRACT......................................................................................................................x ii CHAPTER 1 INTRODUCTION........................................................................................................1 Statement of the Problem..............................................................................................1 Plan of Study.................................................................................................................4 Study Systems...............................................................................................................5 Wildlife Response to Land Use Type...........................................................................8 “Reference” Habitats.............................................................................................8 Agriculture...........................................................................................................12 Amphibians and agricultural landscapes......................................................13 Birds and agricultural landscapes.................................................................16 Silviculture..........................................................................................................19 Amphibians and silvic ultural landscapes.....................................................19 Birds and silvicultural landscapes................................................................22 Urban/Residential/Roadways..............................................................................24 Amphibians and urban landscapes...............................................................25 Birds and urban landscapes..........................................................................28 Roads............................................................................................................34 Studies on the Response of Amphibians a nd Birds to Multiple Land Use Types36 Habitat Fragmentation................................................................................................41 Amphibian and Avian Assemblages as Indicators of Land Use Intensity..................43 Landscape Development In tensity (LDI) Index.........................................................45 2 METHODS.................................................................................................................47 Plan of Study...............................................................................................................47 Site Selection..............................................................................................................47 Field Sampling............................................................................................................51 Visual and Auditory Encounter Surveys.............................................................52

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v Automated Recorder Surveys..............................................................................53 Combined Sampling Effort..................................................................................53 Dip-net Sampling.................................................................................................54 Terrestrial Insect Sampling..................................................................................55 Other Biotic Sampling.........................................................................................55 Environmental Variables............................................................................................56 Water Sampling...................................................................................................56 Fire History..........................................................................................................56 Landscape Variables............................................................................................57 Landscape Development Intensity Index.....................................................57 Landscape indices........................................................................................58 Data Analysis..............................................................................................................60 Water Chemistry and Landscape Variables........................................................61 Species Richness.................................................................................................61 Species Composition...........................................................................................62 Indicator Species.................................................................................................65 Community Characteristics.................................................................................66 Amphibian community.................................................................................66 Avian community.........................................................................................67 3 RESULTS...................................................................................................................68 Wetland Characteristics..............................................................................................68 Wetland Vegetation.............................................................................................68 Wetland Water Depth..........................................................................................69 Wetland Water Chemistry...................................................................................70 Wetland Terrestrial Insects..................................................................................70 Circumjacent Land Use Characteristics......................................................................71 Landscape Development Intensity..............................................................................72 Amphibians.................................................................................................................73 Amphibian Species Richness..............................................................................73 Amphibian Species Frequency of Occurrence....................................................76 Amphibian Breeding Effort.................................................................................78 Amphibian Species Composition........................................................................78 Amphibian Indicator Species..............................................................................87 Amphibian Community Characteris tics and Fish Composition..........................90 Birds.......................................................................................................................... ..94 Bird Species Richness.........................................................................................94 Bird Species Composition...................................................................................97 Bird Indicator Species.......................................................................................103 Bird Community Characteristics.......................................................................105 4 DISCUSSION...........................................................................................................113 Species Richness.......................................................................................................114 Landscape Versus Wetland Species Richness..........................................................115 Species Composition................................................................................................118

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vi Land Use Type and Intensity.............................................................................118 Species Composition Thresholds and Gradients...............................................119 Correlates of Species Composition...................................................................123 Correlates of amphibian species composition............................................123 Correlates of avian species composition....................................................127 Indicator Species.......................................................................................................128 Amphibian Indicator Species............................................................................129 Amphibian indicators of reference co ndition and low intensity land uses.130 Amphibian indicators of high intensity land uses and ubiquitous species.134 Avian Indicator Species.....................................................................................136 Species Guilds and Land use....................................................................................139 Amphibian Guilds.............................................................................................139 Avian Guilds......................................................................................................141 Landscape Development Intensity (LD I) Index and Species Composition..............145 Variability in LDI index....................................................................................146 Variability in the Landscape..............................................................................147 Conclusion................................................................................................................149 APPENDIX A SITE SURVEY LENGTHS......................................................................................151 B CORRELATES OF AMPHIBIA N SPECIES COMPOSITION..............................154 C CORRELATES OF AVIAN SPECIES COMPOSITION........................................156 D SITE CHARACTERISTICS....................................................................................158 E SAMPLED AMPHIBIAN SPECIES ATTRIBUTES..............................................164 F SAMPLED AVIAN SPECIES ATTRIBUTES........................................................166 LIST OF REFERENCES.................................................................................................171 BIOGRAPHICAL SKETCH...........................................................................................194

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vii LIST OF TABLES Table page 2-1 Mean wetland size (ha) and the mi nimum and maximum wetlands sampled per land-use category and region...........................................................................48 2-3 Independent variables included in stepwise regressions predicting the amphibian based NMS axes scores........................................................................65 3-1 The mean and standard deviation of the wetland macrophyte species richness, percent canopy cover, and tree ba sal area (>10.2 dbh) by land use.......................69 3-2 The mean proportion and standard devi ation of the basal area of the wetland overstory vegetation by land use............................................................................69 3-3 The mean and standard deviation of the maximum wetland water depth at sites with standing water by land use, a nd the count of site s without standing water at the time of sampling by land use..............................................................70 3-4 Mean and standard deviation of wate r chemistry parameters by land use. .........71 3-5 The mean and standard deviation (in parentheses) of attributes of the landscape surrounding the sample wetlands by land use type...............................72 3-6 The range, mean, and standard deviati on of LDI scores by LDI quartile, and the number of sites by land use type for each LDI quartile...................................72 3-7 The Pearson correlation coefficients between the site’s 100m buffer LDI score and selected environmental variables....................................................................74 3-8 Amphibian species richness by land us e, 100 meter LDI quartile, and Florida region.....................................................................................................................75 3-9 Amphibian frequency of occurrence by land use type and all sites combined.a....79 3-10 MRPP test for significant differences in amphibian species composition between land use types and LDI quartiles.............................................................82 3-11 Coefficients of determination (R2) between amphibian based NMS ordination distances and distances in the original space.........................................................83

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viii 3-12 Pearson’s correlation coefficients ( r) and Kendall’s Tau comparisons with NMS ordination axes based on amphibian species composition and LDI, FWCI, and maximum water depth.........................................................................84 3-13 Independent variables included in stepwise regressions predicting the amphibian based NMS axes scores........................................................................84 3-14 The variables and regression coeffici ents selected (p<0.05) from a stepwise regression using the Landscape variables listed in Table 313 to represent the amphibian based NMS ordination axes.................................................................85 3-15 The variables and regression coeffici ents selected (p<0.05) from a stepwise regression using the Landscape, Focal Wetland, and Sampling variables listed in Table 3-13 to represent the am phibian based NMS ordination axes.................86 3-16 The variables and regression coeffici ents selected (p<0.05) from a stepwise regression using the Landscape, Focal Wetland, Sampling, Water Chemistry variables listed in Table 3-13 to represent the amphi bian based NMS ordination axes........................................................................................................................86 3-17 Significant (p<0.10) amphi bian indicator species va lues by LDI quartile with quartiles three and four combined..........................................................................89 3-18 Significant (p<0.10) amphibian indi cator species values by land use...................90 3-19 Significant (p<0.10) amphibian indi cator species values by land use (agriculture and urban versus reference and silviculture)......................................90 3-20 The mean proportion of amphibian sp ecies independent of wetlands for breeding, dependent on ephemeral wetlands, and will facultatively use ephemeral wetlands by land use and LDI quartile.................................................91 3-21 The number of sites where fish species occurred by land use...............................93 3-22 Avian species richness by land use, LDI quartile, and Florida region..................96 3-23 Mean number and standard deviation of bird species detected within the wetlands, landscape, total, and landscap e to wetland ratio during two 15 min surveys by land use, LDI quartile, Flor ida region and all sites combined.............97 3-24 MRPP test for significant differences in bird species composition between sites by land use and LDI score quartiles...............................................................98 3-25 Coefficients of determination (r2) between avian based NMS ordination distances and distances in the original space.......................................................100

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ix 3-26. Pearson’s correlation coefficients (r ) and Kendall’s Tau comparisons with NMS ordination axes based on avian species composition and the environmental variables with the strongest correlation...............................................................101 3-27 Independent variables in cluded in stepwise regressi ons to predict the avian based NMS axes scores........................................................................................101 3-28 The selected variables and regression coefficients selected (p<0.05) from a stepwise regression using the Landscape variables listed in Table 3-27 to represent the avian base d NMS ordination axes..................................................102 3-29 The selected variables and regression coefficients selected (p<0.05) from a stepwise regression using the Lands cape, Focal Wetland, and Sampling variables listed in Table 3-27 to represent the avian based NMS ordination axes...........................................................................................102 3-30 The significant (p<0.10) avian indicator values of land use................................104 3-31 The significant (p<0.10) avian in dicator values of LDI quartiles........................105 3-32 Mean proportion of sampled bird sp ecies predominate foraging strategy by land use and LDI quartile.....................................................................................108 3-33 Mean proportion of sampled bird sp ecies predominate nesting strategy by land use and LDI quartile.....................................................................................109

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x LIST OF FIGURES Figure page 1-1 Systems diagram of landscape biotic co mposition storages and driving energies, including human influence.......................................................................................2 2-1 The distribution of the survey si tes by land use and Florida region......................48 2-2 The mean proportion of land cover type within 200 meters of each focal wetland by land use categories..............................................................................50 3-1 Mean and standard deviation of am phibian species richness within the wetland and within the wetland and the 200m buffer by land use.........................77 3-2 Mean amphibian species richness wi thin the wetland and within the wetland and the 200m buffer by LDI quartile.....................................................................77 3-3 The mean tadpole count, mean number of tadpole species, mean number of anuran species calling and mean maximum intensity of anuran species calling by land use.............................................................................................................80 3-4 The mean tadpole count, mean number of tadpole species, mean number of anuran species calling on cassette tape s, and mean maximum intensity of anuran species calling by LDI quartile..................................................................80 3-5 Unrotated joint plot of the amphibian based NMS ordination site scores for axis one and axis three by land use with the strength and re lationship of three environmental variables (FWCI, LDI, and maximum depth)................................83 3-6 Unrotated joint plot of the amphibian based NMS ordination species scores for axes one and three with the strength a nd relationship of three environmental variables (FWCI, LDI, and maximum depth)........................................................88 3-7 The proportion of all sites and only sites that had standing water in 2003 where fish were encountered by land use..............................................................92 3-8 The proportion of the amphibian community that attaches eggs to substrates, has free floating eggs, attaches or has floating eggs, and ha s terrestrial egg development by land use........................................................................................93

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xi 3-9 The proportion of the amphibian community that attaches eggs to substrates, has free floating eggs, attaches or has floating eggs, and ha s terrestrial egg development by LDI quartile.................................................................................94 3-10 Unrotated joint plot of the avian based NMS ordinati on site scores for axes one and three with the strength and rela tionship of the landscape variables LDI and FWCI.......................................................................................................99 3-11 Unrotated joint plot of the avian based NMS ordinati on site scores for axes one and two with the strength and relati onship of three landscape variables LDI, FWCI, and proportion of the landscape within 200 meters of the wetland that was used for agriculture..................................................................100 3-12 The mean proportion of avian species within each feeding guild by LDI quartile..........................................................................................................107 3-13 The mean and standard deviation of the proportion of avian species that maintain feeding territories at each site by LDI Quartile and Land Use.............108 3-14 The mean proportion and standard e rror of the exotic and Neotropical migrant birds recorded by LDI Quartile and Land Use.......................................110 3-15 The mean proportion and standard error of bird species th at have significant (p<0.1) population trends in the Breedi ng Bird Survey data for Florida since 1966......................................................................................................................111 3-16 The mean weight (g) and standard e rror of avian species detected at the study sites by LDI quartile and Land use.............................................................112 3-17 Log10 of avian mass versus log10 of avian territory or home range size..............112

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xii Abstract of Dissertation Pres ented to the Graduate School of the University of Florida in Partial Fulfillment of the Requirements for the Degree of Doctor of Philosophy AMPHIBIAN AND AVIAN SPECIES COMPOSITION OF FORESTED DEPRESSIONAL WETLANDS AND CIRCUMJACENT HABITAT: THE INFLUENCE OF LAND USE TYPE AND INTENSITY By James A. Surdick Jr. August 2005 Chair: Mark T. Brown Major Department: Environmental Engineering Sciences Wetlands provide wildlife habitat; howe ver, circumjacent anthropogenic land uses influence abiotic and biotic aspects of wildlif e habitat and in turn possibly which species will persist in a landscape. The main objectiv e of this study was to record the amphibian and avian species composition of depressional forested wetlands embedded in landscapes of varying human land use intensity and rela te compositional differe nces to wetland and landscape characteristics. Presence/absen ce surveys for amphibian and avian species were conducted at 111 small (<2.5 ha) forest ed depressional wetlands throughout Florida within four common land use types, natural or reference, silvicul ture, agriculture, and urban/residential. Despite a modest relationship between land use and species richness, the results of nonmetric multidimensional scaling ordi nations and multi-re sponse permutation procedures indicated a strong association w ith amphibian and avian species composition and land use. Indicator species, species sens itive and tolerant of human land uses, were

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xiii identified. The relevance of these results may broaden with the finding that not only were there species-specific responses, but also predominate life histor y traits of the avian and amphibian community varied with land us e context. For example, amphibians that were obligatory ephemeral pond breeders decreas ed with increasing land use intensity. Within the avian community, insectivores, bark gleaners, canopy gleaners, territorial species, ground nesters, and cavity nester s decreased, while omnivores, herbivores, ground gleaners, canopy nesters, and exotic species increased, with increasing land use intensity. The Landscape Development Intensity index was a remotely measured predictor of amphibian and avian species assemblages. Amphibian species composition was also significantly associated with the Florida We tland Condition Index (FWCI), distance to the nearest wetland, maximum wetland water depth, wetland water pH, wetland water specific conductivity, and geogr aphical location of the st udy sites. Avian species composition was also significantly associat ed with the proportion of agricultural land within 200 meters, FWCI, proportion of rete ntion ponds/canals within 200 meters, and geographical location of the study sites. The results of this study suggest that merely preserving wetland habitat in developed landscapes will not be enough to support a wildlife comm unity indicative of natural settings. However, wetlands embedde d in even the most intensive land uses provide valuable wildlife habitat.

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1 CHAPTER 1 INTRODUCTION Statement of the Problem The increasing spatial extent and intensity of human uses of land worldwide are affecting both biotic and abiotic aspects of ecological systems and altering ecological processes (e.g., Vitousek et al. 1997). The result of these alte rations is a landscape that is becoming dominated by ecological system s that Odum (1962) termed, emerging ecosystems associated with man, and late r as interface ecosystems, whose landscape conditions and driving energies are different from their na tural counterp arts found in more remote landscapes (Odum 1971). It is suggested that biological integrity of a system is maintained when the species composition and functional organization are comparable to the natural ha bitats of the region (Karr a nd Dudley 1981). The biological integrity of ecological communities within hum an dominated landscapes is a concern of many, especially those who are interested in conserving native biota. Despite this concern, there is still no clear understanding of the changes in w ildlife composition along a gradient of human land use intensity. In just 100 years Florida’s population has increased from 530 thousand to nearly 17 million people (U.S. Census Bureau 2003). By the year 2025, almost 21 million people are projected to live in Flor ida (Cambell 1997). With the past and projected population increases in Florida come major habitat tr ansformations and changes to energy flows within and between different land systems. Ev en more changes in the future are likely as the intensity and spatial extent of Flor ida’s population increase s continue. Human

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2 induced changes to ecological communities are complex, and include both direct and indirect effects. Figure 1-1. Systems diagram of landscape biotic composition, storages, and driving energies, including human influence. Figure 1-1 is a systems diagram of an aggregated food web that includes main driving energies and human influence. Show n are aggregated compartments of the food web, each containing many species that combined make up landscape biodiversity. Every landscape has a suite of species that are adapted to its driving energies and habitat structure, with specific speci es responding to subtle differe nces in its physical structure and the timing and intensities of driving en ergies. The diagram shows the effect of humans on each compartment and ultimately on species composition as both direct (harvesting of species or stress as a result of pollutants or toxins) and indirect (changing runoff characteristics or increa sing nutrient inflows). In a ddition, the direct effects can have indirect affects as individual species within compartments adjust to changes in

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3 others. For instance, Polis et al. (1997) suggest bottom up and top down indirect effects that may result from direct effects. Terbor gh et al. (1999) give as an example, hunting pressure on top carnivores that may affect th e entire ecosystem indirectly by influencing species abundances and persistence in lowe r levels of the food web. The interplay of direct, indirect, and cumulative effects makes the identification of sp ecific mechanisms of shifts in species composition difficult. With the changes in land use that Florida has experienced in the past 100 years (in both spatial extent and intensit y of use) presumably there ha ve been changes in wildlife use and persistence. However little resear ch has been conducted that would lend insight into these changes especially related to the impacts of differing intensity of uses. The main research question that this dissertati on addresses is directed at this lack of information: How do amphibian and avian communities respond to the anthropogenic influences associated with di fferent land use intensities? While this question begs to be answered for the entire landscape and for all wildlife species, it was obvious from the beginning that the study of human in fluences on wildlife communities should be narrowed to a specific ec osystem type and a subset of wildlife in order to reduce the size of the study to one that was manageable within the time and resource constraints available. Isolated wetlands offer an inte resting point of reference in that they are embedded in a landscape matrix a nd thus can be easily treated as individual cases within varying landscape intensities. Fu rther, isolated wetlands offer the potential to study avian and amphibian populations as th ey are relatively productive habitats for these species. Finally, the macroinvertebrate and vegetative communities of isolated wetlands have been shown to respond to land use intensity of the circumjacent landscape

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4 (Lane 2003, Reiss 2004); thus, changes in amphibian and avian communities might be expected. Plan of Study In this dissertation, depressional forested wetlands (Cowardin et al. 1979) and their immediately surrounding landscape (within 200m) that are dominated by human land uses to varying degrees were studied to be tter understand wildlife responses to human land use intensity. The study took place thr oughout Florida and involved 111 wetlands in landscapes that were a priori characterized as either refe rence (no human dominated land uses), silvicultural, agricultural, or urba n. The main objective of this study was to measure avian and amphibian species composition in these wetlands and their immediately adjacent uplands and relate it to human land use intensity. In Florida small (<2.5ha) isolated forested wetlands or dome swamps persist within primarily four land use categories, reference or natural, silviculture, agriculture, and urban/residential (Monk and Brown 1965, Ka utz 1993). These wetlands serve many functions including the often cited, suppor t of wildlife (e.g., Moler and Franz 1987, Semlitsch and Bodie 1998, Russell et al. 2002). Yet, the self-organi zation of the wildlife community of dome swamps within the four main land use categories, or along a gradient of human land use intensity, is little understood. There have been many studies comparing the pair wise wildlife compos ition amongst common land uses of a region (e.g., Smith and Schaeffer 1992, Delis et al. 1996, Daily et al. 2001, Guerry and Hunter 2002, Renken et al. 2004); however, no pub lished accounts could be found that simultaneously explore the wildlife composition of agriculture, natural, silviculture, and urban/residential. Similarly, studies have explored the bio tic response to an implied or qualitative gradient of anthropogenic land us e intensity (e.g., Blair 1999, Chambers et al.

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5 1999, Miller et al. 2003, Norris et al. 2003); however, rarely is the land use intensity gradient in these studies a qua ntitatively based measure. Af ter a review of the available information Hart and Newman (1996) conclude d that the impacts of human land use on wildlife, especially amphibians, of isolated wetlands in Florida, should be a priority for future research in Florida. This dissertation explores se veral questions related to amphibian and avian species composition in Florida. How similar are sp ecies assemblages within four common land use types (agriculture, reference, silvicul ture, and urban)? Do indices of land use intensity or wetland condition correspond with di fferences in species composition? What environmental variables are significantly associated with amphibian and avian species composition? Which species are indicative of specific land uses or levels of land use intensity? Finally, are certain guilds (gr oups of species that share life history characteristics) sensitive or tolerant to di fferent land use types or levels of land use intensity? Study Systems No part of Florida’s 1,400,000 ha lies above 105m (mean sea level). A relatively flat landscape and temperate to subtropical climate that averages approximately 1.35 to 1.54m of rainfall a year lead to the formati on of numerous wetlands and water bodies. In 1780 approximately 54% of Florida was wetla nds, but land development reduced the proportion to 29% by 1980 (Dahl 1990). Of the remaining wetlands in Florida nearly one third are forested freshwater wetlands with a large proportion typed as stillwater swamp forests which consist primarily of cypr ess basin swamps and gum ponds (Ewel 1990). The focal habitat type in this study was dome swamps which are a subset of stillwater swamp forests. Dome swamps include cypress domes, cypress ponds, cypress

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6 heads, cypress galls, gum ponds, bayheads, pine barren ponds, and Citronelle ponds (Florida Natural Areas Inventory 1990). Do me swamps tend to be circular with the tallest trees occupying the center and tree he ights decrease towards the outer edges (Clewell 1986). The surface water of dome sw amps is hydrologically isolated from other waterbodies and wetlands. The hydroperiod of dome swamps varies but most have standing water periodically throughout th e year (Clewell 1986, Heimburg 1986, Ewel 1990, Folkerts 1997). A general estimate of the hydroperiod for dome swamps is between 200 and 300 days (Florida Natural Ar eas Inventory 1990). Variability in dome swamp hydroperiod can result from differen ces in geomorphology, vegetation type, soil type, ground water connectivit y, local precipitatio n patterns, and human disturbance. Typically, water levels are deeper in the center of dome swamps where peat accumulations are greatest and decrease towards the periphery where mineral soils dominate (Spangler et al. 1976, Ewel 1990). Maximum water depths are usually no more than a meter (Monk and Brown 1965, Clewell 1986) The center of a dome may develop a treeless pond surrounded by marsh ve getation giving the dome a doughnut like appearance (Vernon 1947). Likewise, freshwater marsh vegetation or shrubs fringe many domes where water level fluctu ation and fire frequency are greatest. Like hydroperiod, the fire frequency within dome swamp wetlands is partially dependent on wetland type and the surr ounding landscape composition. Cypress dome swamps may be dependent on periodic fire wh ere the outer edges may burn as frequently as every 3-5 years and the interior every 100 to 150 years. Gum pond and bay wetlands burn less frequently (Florida Natura l Areas Inventory 1990).

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7 Isolated basin wetlands have low productiv ity and growth rates compared to other wetland habitats (Mits ch and Ewel 1979, Brown 1981). Pr oductivity in dome swamps may be influenced by both the available nutrients and hydroperi od (Brown 1981, Ewel 1990). Depression wetlands are primarily dependent on rainwater or groundwater infiltration for recharge and have relatively small catchment basins; thus, wetland surface waters tend to be oligotrophic (Brown 1981). Peaty soils aid in lowering the surface water pH which further decrea ses nutrient availability to dome swamp vegetation (Monk and Brown 1965, Ewel 1990). It is not uncommon to have pH levels in natural dome swamp wetlands range between 3.5-4.5 (Brown 1981, Florida Natural Areas Inventory 1990, Warner and Dunson 1998). Still water swamps are prim arily classified by the domi nant overstory vegetation. For instance, cypress domes ar e dominated by pond cypress ( Taxodium ascendens ), gum ponds by black gum ( Nyssa biflora ), and bayheads by three species of bay tree ( Magnolia virginiana Persea palustris and Gordonia lasianthus ). However, all three often share vegetation components of the other hab itat types (Monk and Brown 1965, Ewel 1990). Other typical overstory species in dome swamps include Pinus elliottii P. serotina P. taeda and Acer rubrum Typical midstory and shrub species include Ilex myrtifolia I. cassine Lyonia lucida Itea virginiana Annona glabra Myrica cerifera Cephalanthus occidentalis Cyrilla racemiflora and Vaccinium corymbosum (Monk and Brown 1965, Florida Natural Areas Inventory 1990). Flatwoods, clayhills, sandhills, and wet a nd dry prairies covered more then 60% of Florida (Kautz et al. 1993) and are the typical unadulterated ha bitats surrounding dome swamps (Clewell 1986, Abrahamson and Hartnett 1990). These systems have evolved to

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8 periodic fire every one to te n years and frequently have Pinus spp. as the dominant overstory vegetation (Florida Natural Areas Inventory 1990). In areas that receive periodic fire, flatwoods, sandhills, and clayhi lls achieve an open savanna like habitat structure and in the absence of fire succeed into hardwood forest (Engstrom et al. 1984, Clewell 1986, Wolfe et al. 1988). Nutrient cycling and availability varies by upland habitat type, but primarily these systems ar e oligotrophic with re latively low productivity (Art and Marks 1971). For at least 5,000 years much of Florid a was pine savanna with interspersed wetlands (Watts 1971, Clewell 1986, Watts et al 1992). It is only within recent times that many of the habitats and driving energies that much of Florid a’s biodiversity has adapted to are experiencing rapi d human induced alteration. Wildlife Response to Land Use Type “Reference” Habitats Throughout the rest of the text “referen ce” refers to undeveloped habitats that have received relatively low anthropogenically derived inputs and al terations, and are the least impacted sites in this study. The reference sites are reasonable examples of Florida’s native communities (e.g., pine fl atwoods, sandhills, dry prairies, cypress domes). There is a set of amphibian a nd avian species that are commonly found in Florida reference habitats. Florida has 53 native amphibians (Enge 1997) and three established exotic anuran species. Depression wetlands are important habitats for many of Florida’s amphibians and a subset requires isolated depressional we tlands for larval development. There are 14 amphibian species that require, and 21 speci es that will utilize, small isolated depressional wetlands as breeding sites in Florida (Moler and Franz 1987).

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9 There have been 196 confirmed breeding bird species in Florida (FFWC 2003). More specifically, Rowse (1980) recorded 52 summer resident bird species in north Florida flatwoods, and Workman (1996) reco rded 14 summer bird species in 9 north Florida cypress domes. No avian species in Florida is dependent on small isolated wetlands; however, facultativ e use is common among several species (Hart and Newman 1995). At any one location in Florida there is a regional pool of species available to respond to changes in the landscape. Regi onal species richness increases with the increasing number of habitat types, and this ha s been suggested as a reason why there is a general pattern of decreasing species richne ss for amphibians and birds from north to south throughout the Florida peninsula (M eans and Simberloff 1987, Engstrom 1993). For example 19 out of 26 of Florida’s salamander species are found only in the Panhandle and/or north Florid a (Enge 1997). This suggests there may be more species sensitive to development in north Florida si mply due to a larger species pool. At the same time South Florida’s warmer annual te mperatures and lower frequency of days reaching 0 C may allow for the establishmen t of more temperatur e sensitive exotic species capable of colonizing bot h disturbed and natural habitats (Simberloff et al. 1997). Although there is an attemp t to utilize “natural” habitats (e.g., parks and preserves) as benchmarks of reference conditi on in studies of ecosystem change even the benchmarks themselves are experiencing, and often have experience d, varying degrees of human induced alterations that have the ab ility to influence biotic composition (e.g., Rooney et al. 2004).

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10 The extinction and extirpation of species is the most obviou s alteration to a region’s biotic composition. Some wildlife species now missing from Florida’s systems due to direct or indirect effects of humans include the passenger pigeon ( Ectopistes migratorius ), Carolina parakeet ( Conuropsis carolinensis ), ivory-billed woodpecker ( Campephilus principalis ), and the Florida red wolf ( Canis rufus floridanus ) (Kautz 1993). Similarly, the geographical ranges of numerous species have been reduced, extirpating them from habita ts where they once interact ed (e.g., Florida panther, Felis concolor coryi from most of Florida). Without replacement these species’ role in systems is lost from both developed and re ference landscapes. Several extinct or extirpated species are top carnivores whose ab sence from their hierarchical positions has allowed a shift in food chain dynamics wh ere lower level carnivores and omnivore populations increase with concomitant potential to affect lower levels of the food chain (e.g., Crooks and Soule 1999). The vegetation structure of today’s referen ce habitats may be very different than in the past. Virtually all of Florida’s old growth timber was harvested between 1870 and 1925 (Clewell 1986). Most of today’s preserves, national forests, national wildlife refuges, and military bases were clear-cut im mediately before their designation (Clewell 1986). Thus the average age of the oldest oversto ry trees in nearly a ll of Florida’s natural areas are no more than 80 years and the trees in the majority of Florida’s forested systems are probably much younger (Wolfe et al. 1988). The effect on forest wildlife of the decrease in the density of older large trees and the loss of their distinct structural characteristics is largely unknown (Harris and Vickers 1984). For instance, cavities occur more often in older and larger trees, and populations of cavity dependent species

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11 may be very different from what they we re before widespread human alteration of Florida’s forested habitats. Timing and natural procession of fire is another forcing function in Florida’s landscapes that has also been significantly altered by human actio ns. Starting in the 1920s large scale programs to stop natural forest fires and prescribed burns were enacted by federal officials (Clewell 1986). Many of Florida’s natural syst ems are dependent on fire for their maintenance. Obstructions su ch as roads, canals, and open fields inhibit fires from covering the extent that traditi onal fires would have. Even the timing of landscape fires has largely changed in recent tim es where winter burns are more frequent. The historical fire season in Florida was mo st likely between late April and June when habitats were driest and the advent of su mmer convective rainstorms brought frequent lightning strikes (Robbins and Myers 1992). Several pineland plants require summer fires to stimulate flowering (Wolfe et al. 1988, Robbins and Myers 1992). Many of Florida’s native species have adapte d to habitats that experience frequent ground fires that help to maintain an open sa vanna like habitat with numerous grasses and forbs (Engstrom et al. 1984). For example, the size, density, and productivity of redcockaded woodpecker (an endangered southeaste rn pine savanna specialist) social units were highly correlated with ground cover comp osition and natural pi ne regeneration, both of which were surrogate measures of local fire history (James et al. 1997). Furthermore, red-cockaded woodpeckers laid larger clutches the year after their territories had been burned (James et al. 1997). Engstrom et al (1984) documented the gains and losses of bird species over a 15 year period of fire suppression in a north Florida pinewoods. Notably, loggerhead shrike, Bachman’s sparro w and blue grosbeak disappeared within a

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12 few years of fire suppression and species that favor brushy areas dominated, common yellowthroat, indigo bunting, rufous-sid ed towhee, and white-eyed vireo. There are relatively few recent baseline studies of amphibian and avian assemblages in reference Florida landscapes ; however, mounting evidence suggests some species decline or even disappe ar as human inputs to habita ts increase (e.g., Blair 1996) which implies a subset of species exists that is indicative of the lowest intensity land uses in Florida. Agriculture Approximately 1000 years ago Apalachee and Timucuan Indians maintained croplands in Florida, and in parts of the fe rtile clayhills of the northern panhandle their fields stretched “as far as the eye could re ach” (in Clewell 1986). Spanish colonizers brought cattle, hogs, and other grazing animals and for almost 300 years much of Florida was open rangeland where the only fen ces were around croplands (Clewell 1986, Daugherty 1989, Ewel 1990). The amount of land area influenced by grazing was reduced with the advent of the 1940s fence law requiring pastures to have fences, and rangeland started becoming less extensive and more intensive (Clewell 1986). During this same time period cropland intensificati on increased with the a dvent of widespread manufactured fertilizer use and essentially the traditional fallow field rotation system ceased (Wolfe et al. 1988). High rates of land conversion to agricu lture occurred from 1936 to 1987 when agricultural land increased from approximately 17.5% to 29.7 % of the state (Kautz 1993). Today, the amount of agricultural land appear s to have leveled off and may even be decreasing (Economic Research Service 2002). Based on data from Vesterby and Krupa (2001) approximately 36.5% or 5.1 million ha of Florida is cropland or grazed by

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13 domestic animals. Cropland is a higher in tensity land use and it comprises 29% of all agricultural lands or 1.5 million ha. In Florid a the most common types of agriculture are rangeland, improved pasture, dairy, citrus, ro w crops, sugar cane, and greenhouse/nursery (FASS 2005). Given the large percent of la nd area utilized for ag riculture there are surprisingly few studies a ddressing amphibian or avia n use of Florida farmland landscapes. Amphibians and agricultural landscapes Meshaka (1997) described the herpetofauna of a ranch in central Florida and documented 19 amphibian species in various ha bitats. Many of the species utilized the abundant ephemeral ponds on the property. Eastern spadefoot ( Scaphiopus holbrookii ) was expected, but reported missing from the ra nch. Hydrological manipulations of the improved pasture which maintain water levels high during the dry season and lower than normal during the wet season were stated as th e best explanation for the absence of this highly fossorial species. A co mparison with a lower or higher intensity land use was not made. Folkerts (1997) commented on the relati vely low occurrence of amphibians in Citronelle ponds (abrupt topographical depressi ons within the Citrone lle formation). The lower than expected amphibian species richne ss was attributed to extensive disturbance to the surrounding terrestrial habitat. Unlike many ephemeral isolated wetlands Citronelle ponds are surrounded by relatively fe rtile sandy loam soils used extensively for row cropping (Wolfe et al. 1988, Folkerts 1997). Amphibian species frequently encountered in Citronelle ponds included Acris gryllus, Hyla cinerea, Rana catesbeiana, Rana grylio, Pseudacris nigrita Pseudacris crucifer, and Gastrophryne carolinensis (Folkerts 1997).

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14 In locations other than Flor ida, agricultural landscapes have been implicated in altering amphibian species composition, ric hness, and abundance. For example, in Australia, Jansen and Healey (2003) attribut ed observed decreases in frog communities, species richness and some indi vidual species populations to increased grazing intensity. Knutson et al. (2004) did not compare natura l to agricultural wetla nds but did find frog species richness and reproductive success was similar in Minnesota ponds and wetlands surrounded by row crops and non-grazed pasture. Ponds used by livestock for water had higher turbidity, higher phosphorus and lower amphibian reproductive success. Species richness was highest in ponds with lower to tal nitrogen levels (K nutson et al. 2004). A significantly greater number of two spadefoot toad species was found in isolated playa wetlands surrounded by crops than in those surrounded by lightly grazed pasture. Species richness and abundance for the rema ining species did not differ by agricultural land use type (Gray et al. 2004). Amphibians exhibit species specific res ponses to agricultural landscapes that depend on life history characteristics and la ndscape attributes. Kolozsvary and Swihart (1999) studied amphibian use of forest patches and found American toads ( Bufo americanus ) and gray treefrogs ( Hyla chrysoscelis ) were tolerant to most agricultural landscapes, ranid frogs responded to the nearness of adjacent wetlands, redback salamanders ( Plethodon cinereus ) were correlated with the amount of forest area, and several species responded to wetland hydroperiod. Ray et al. (2002) found presence of common toads ( Bufo bufo ) declined with cultivated fi eld area and alpine newt ( Triturus alpestris ) presence declined with vineyard area surrounding ponds.

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15 American toads ( Bufo americanus ) and northern leopard frogs ( Rana pipiens ) were negatively associated with the amount of fo rested area within 1 km of small (<0.5ha) isolated wetlands in a Maine agricultural la ndscape. Five other amphibian species were positively associated with forested area a nd two species were most likely to occupy ponds with adjacent forest (Guerry and Hunter 2002). The presence of four Ohio amphibians had a positive associ ation with the amount of fo rest cover within 200m of breeding ponds (Porej et al. 2004). The remaining species had negative responses to cumulative lengths of paved roads within 1km of the breeding ponds and the distance to the nearest five wetlands (e.g., Notophthalmus viridescens )(Porej et al. 2004). Herbicide, pesticide, and fertilizer use is prevalent in agricultural landscapes. Nutrients from fertilizers and animal waste have the potential to alter wetland biogeochemical processes and species co mposition. Reproductive success of the northwestern salamander ( Ambystoma gracile ) and northern red-legged frog ( Rana aurora ) were found to be significantly lower in wetlands receiving agricultural runoff (de Solla et al. 2002). High ammonia concen trations and high biochemical oxygen demand are suspected in causing the observed diffe rence in hatching success. Algal mats frequently form in eutrophic waters a nd Palis (1997) observed that Flatwoods salamanders ( Ambystoma cingulatum ) are not found in wetlands with excessive amounts of algae. Houlahan and Findlay (20 03) found the presence/absence of 12 of 13 amphibian species was negatively correlated to total nitrogen levels in 74 Ontario wetlands. Nitrates and nitrog en fertilizers at ecologically relevant levels are known to have toxic and lethal effects on amphibian species (Hecnar 1995, A dolfo et al. 1999).

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16 In a study in the Prairie Pothol e wetlands of Canada, 288,000 wetlands (approximately 17% of the wetlands in the ar ea) were estimated to exceed Canadian pesticide guideline levels for the protection of aquatic life (Donald et al. 1999). Berrill et al. (1998) found lethal and subl ethal affects on three larval anuran species exposed to relatively low doses of a comm on agricultural pesticide, endosulfan. American toad ( Bufo americanus ) tadpoles were more tolerant to endosulfan exposure than wood and green frogs ( Rana sylvatica and R. clamitans ). Bridges and Semlitsch (2000) also found species specific lethal and sublethal responses as well as population specific responses within Rana sphenocephala to Carbaryl, a common crop and garden insecticide. Finally, larval Xenopus laevis exposed to ecologically relevant le vels (0.01 to 40 ppb) of atrazine, the most widely used herbicide in the U.S ., exhibited hermaphroditism and demasculation (Hayes et al. 2002). Not only do the active ingredients in pest icides have an apparent effect on amphibians but the surfactants used in common pesticides were found to have a narcotic effect. For instance, Mann and Bidwell (2001) found that after expos ure to surfactants tadpole feeding activity wa s reduced and the effect s were compounded under low dissolved oxygen conditions. Birds and agricultural landscapes A comparison of avian richness on four agri cultural land use types in west-central Florida revealed a relationship between bird species richness and land use type (Cutright 1981). Species richness was greatest in nati ve grazed forest habitats (included dome wetlands), followed by unimproved pasture, improved pasture, and cropland supported the lowest number (Cutright 1981). The most abundant species during the summer season in the pasture habita ts were eastern meadowlar k, cattle egret, and northern

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17 bobwhite quail, and in the cropland mourning dove and redwinged blackbird were the most common. Shrub dependent species (e.g., Carolina wren and white-eyed vireo) and vultures dominated the grazed forest habitats. Hirth and Marion (1979) found the eastern meadowlark and northern bobwhite quail were the most common su mmer residents in native rangeland (grazed flatw oods) of south Florida, and summer granivore densities were twice that of insectivores. A single species study by Morrison and Humphrey (2001) found higher densities and higher nesting success of the fede rally threatened crested caracara ( Caracara cheriway ) within Florida cattle ranches than on lands managed as natural areas. The loggerhead shrike is often positively associat ed with lower intensity agricultural lands and other open areas, and population declin es in Florida and elsewhere have been attributed to decreases in low intensity agricultural lands since the 1940s (Cade and Woods 1997). Several studies conducted in tr opical areas indicate avia n species richness is higher in or near remnant forest patches than in the surrounding agricultural matrix (Daily et al. 2001, Luck and Daily 2003, Naidoo 2004). Avian species composition is also distinct in tropical regions between lower and higher intensity agricultur al land uses (Daily et al. 2001, Naidoo 2004). DesGranges and Boutin (1996) analyzed avian population trends along Breeding Bird Survey (BBS) routes in Quebec and f ound species experiencing population declines were associated with lower intensity agri culture (e.g., pastureland) while increasing species were associated with the recent incr eases of more intensive agriculture (e.g., cash crops). In England population d eclines of farmland birds were associated with a trend in

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18 increasing agricultural intensity (Chamberlain et al. 2000). In Iowa bird species richness declined on a continuum of landscapes types from forest to row-crop monocultures (Best et al. 1995). In North Dakota the best i ndicator of duck abundance in shallow wetlands was the proportion of upland ar ea comprised of agricultural land (Austin et al. 2001). Three levels of grazing intensity were a ssociated with changes in avian species composition and richness in Utah. Metrics of land use intensity were created using species dominance and richness. However, th ese metrics were only significant for the high impact sites (Bradford 1998). Finally, Beecher et al. (2002) reported higher bird species richness on organic farmland than on nonorganic farmland, 54 and 39 species respectively. They attributed the higher bird species richness to greater foraging opportunities provided by the increased biomass of non-crop vegetation at sites that were not treated with herbicide. Ford et al. (2001) reviewed why many sp ecies of birds have declined in the eucalyptus woodlands throughout the agricultu ral zone of Australia. Although species from several feeding guilds have declined, insectivores dependent on native vegetation experienced the biggest declines. Hypotheses of mechanis ms responsible for declines in some species included poor dispersal capabilities between habitat patches, competition with increasing populations of larger gene ralist species, habitat specificity and the inability to adapt to vegetation changes in fragments, increased nest predation by mesopredators in habitat fragme nts, and degradation of food re sources (Ford et al. 2001). Declining and increasing avian species populations to changes in land use and varying avian species richness by land us e types suggest ther e are avian species indicative of land use intensity within agricu ltural habitats. For in stance, brown-headed

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19 cowbird, eastern meadowlark and Brewer’s spar row were associated with lower intensity agriculture and ring billed gull, rock dove and horne d lark with higher intensity agriculture (DesGranges and Boutin 1996, Bradfo rd et al. 1998). No studies were found that utilized birds as statistically signifi cant indicators of agricultural land use or agricultural land use intensities. Silviculture Approximately 43% of Florida is managed for commercial forestry and more then 27% of that area is plan ted in dense stands of Pinus elliottii P. taeda or P. clausa (Kautz 1993). A mature stand of planted pine superfic ially appears to be th e least altered of the anthropogenic land-use types discussed in this paper, and perhaps silvicultural areas are less intensively managed than both agricultura l and urban areas. However, mechanical site preparation, fertilizer augmentation, he rbicide application, re latively shor t rotation cycles, and densely stocked stands cause a significant shift in the composition and structure of forest vegetation (Clewell 1986 ). Likewise, the conversion from native landscapes to pine plantations a ppears to be associated with a shift in vertebrate species composition. Amphibians and silvicultural landscapes The flatwoods salamander ( Ambystoma cingulatum ) may be sensitive to silviculture practices in Florida. Means et al. (19 96) observed a drastic decline in a large A. cingulatum population over a 22-year period. Al though unable to demonstrate a direct link, they attributed the declin e to the silvicultural practi ces occurring in the surrounding landscape. The direct effect of mechanical site preparation (e.g., roller chopping, disking, and bedding) on fossorial species is poorly understood. Wolfe et al. (1988) stated the amphibians found in pine plantations, such as oak toad ( Bufo quercicus ) and pinewoods

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20 treefrog ( Hyla femoralis ), entered from adjacent communities following large scale silvicultural related disturba nces. Similarly, a recovery of flatwoods herpetofauna populations three years after a north Florida clear-cut was at tributed to interspersed cypress dome habitats that provided refugia during the early stages of forest succession (Enge and Marion 1986). Ephemeral wetlands in silvicultural areas may be important refugia and habitat for several amphibian species. Ru ssell et al. (2002) recorded 20 amphibian species utilizing edge habitat of five ephemeral wetlands em bedded in a South Carolina silvicultural landscape, while in the adjacent uplands only 8 species were recorded. Amphibian use of the wetlands was negatively correlated to th e amount of hardwoods in the uplands and the distance to the nearest wetland, and positivel y correlated to the conifer basal area surrounding each wetland (Ru ssell et al. 2002). A common practice in flatwoods managed for silviculture is to ditch the isolated wetlands to increase connectivity and flow. Furthermore, frequently wetlands are encircled with a ditch to retard fire. As a result, wetland hydroperiods are altered. On lands in north Florida used for timber produc tion, Vickers et al. (1985) found nearly 4 times as many terrestrial reptiles and toads in ditched cypress ponds and 1.4 times more frogs and salamanders in unditched cypress pond s. Furthermore, ditching can connect once formerly isolated wetlands to sources of predatory fish (Babbitt and Tanner 2000). The species specific response of amphibians to fish predation influences the amphibian species composition of wetla nds (Kats et al. 1988). Herbicides and fertilizers are also pot ential anthropogenically induced stressors within silvicultural landscapes. Intensively managed Florida pine pl antations often have

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21 herbicide and fertilizer applie d two times during each stand rotation. Hatch et al. (2001) observed avoidance behavior in several we stern amphibians exposed to paper towels dosed with urea (a common forest fertilize r). However, the amphibians did not avoid soils dosed with urea, and exposed individua ls experienced higher mortality and lower feeding rates than controls. An insecticide, fenitrothion, used to control spruce budworm in boreal forests caused paralysis in several tadpole species. Bullfrog ( R. catesbeiana ), green frog ( R. clamitans ), and spotted salamander ( Ambystoma maculatum ) were the most sensitive followed by the wood frog ( R. sylvatica ) and leopard frog ( R. pipiens ). The American toad tadpoles ( B. americanus ) were the most resilient (Ber rill et al. 1995). Similarly Berrill et al. (1994) found three tadpole species died or were paralyzed upon exposure to a herbicide, triclopyr, and an insecticide, fenitrothion, and the order of increasing sensitivity was R. catesbeiana R. clamitans and R. pipiens Florida pine plantations are generally clear-cut on a 25-30 year rotation cycle (Clewell 1986). Several researchers have documented differences in the amphibian community in clear-cut areas versus forested areas. Enge (1984) found that amphibian abundance was reduced 10-fold in North Fl orida clear-cuts while amphibian species richness was not affected Major reductions in Scaphiopus holbrookii Gastrophryne carolinensis and Rana sphenocephala were observed. North Carolina salamander abundance was five times greater and species richness two times greater in mature forests (50-70 years) than in clear-cuts (P etranka et al. 1993). Amphibian response to clear-cuts appears to be species specific where some species are sensitive and others tolerant of tree harvesting (e.g., deMa ynadier and Hunter 1998).

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22 Intuitively, arboreal species are impacted by the removal of trees, but clear-cutting also reduces relative humidity and increases inso lation and evaporative water loss from the soil (Raymond and Hardy 1991). The “drying out” of the habitat may partly explain the observed declines in amphibian abundan ce and richness within clear-cut areas. In most industrial and commercial forests in Florida clear-cut ar eas are immediately replanted with an even aged, densely stocke d monoculture. Even aged stands of trees have been reported to have lower local a bundances of amphibians than uneven aged mixed forest; however, a reduction in species richness or compositi on was not detected (Renken et al. 2003). Mature even aged stands have higher densities of redback salamanders ( Plethodon cinereus ) than regenerating and sapling even aged stands (DeGraaf and Yamasaki 2002). Amphibian species richne ss and composition among diffe rent isolated wetlands within managed forests of a region appear to be consistent (Russell et al. 2002) that suggests amphibian composition may be used as an indicator of this land use. Birds and silvicultural landscapes Avian abundance and species richness is hi gher in Florida silvicultural landscapes that contain forested wetlands (Marion and O’Meara 1982, Harris and Vickers 1984). Furthermore, the ecotone of small cypre ss wetlands and clear-cuts supports higher breeding bird abundances and species richness th an the ecotone with planted pine stands (Harris and Vickers 1984). The vegetation heig ht heterogeneity of cypress domes is the suspected reason why bird density and divers ity are greater than in the surrounding pine plantations or clear cuts (M arion and O’Meara 1982). Cypress domes also have more cavities than pine stands a nd approximately 30% of the flatwoods/cypress dome avian community are cavity nesting species (Row se 1980, McComb et al. 1986). Finally,

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23 forested wetlands within a pine plantati on matrix may experience pulses of increased avian richness by providing refugia for forest dependent species after adjacent pines are harvested (Marion and O’Meara 1982). Comparisons of north Florida pine planta tions divided into three levels of site preparation intensities indicat e that avian species richness and abundance are highest in the low and intermediate intens ity stands (Harris et al. 1975). Furthermore, all plantation sites (n=9) had lower avian richness and a bundance than mature stands of longleaf ( Pinus palustris ) or slash ( Pinus elliottii ) pine. Finally, mature longl eaf pine stands had more species than mature slash pine stands (Harris et al. 1975). In Oregon, Chambers et al. (1999) reported the avian res ponse to thr ee intensities of silvicultural treatments and uncut contro l stands. The avian community of the low intensity treatments was most similar to the control stands. Avian community composition in the moderate and highest intensity treatments was significantly altered. The number of species that declined was approximately equal to the number that increased in the moderate and highest intensity treatments. Similarly, Tittler et al. (2001) found the abundance of 10 species increased, 7 decreased, and 10 remained the same in silvicultural treatments relative to controls in Alberta, Canada. Four species were only found within the silviculture treatments. Keller et al. (2003) found species richness increased 2-6 years after clear-cutting then decreased between 7-25 years after cutting and then increased again after 25 years. They concluded avian guilds respond to successional stages of forest structure and productivity (Keller et al. 2003). Estades and Temple (1999) included the la ndscape mosaic in exploring the effect of habitat fragmentation on Chilean avian communities in managed forest systems.

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24 Fragmentation effects were species specific. The avian community was significantly correlated to the vegetation adjacent to the study forest fragments. For example, the abundance of cavity nesting species was positivel y correlated to the proximity of native forest patches. Snags are often removed or at very low de nsities in southeastern pine plantation forests. A habitat manipulation experime ntal study in loblolly pine forests ( P. taeda ) found avian abundance, richness, and diversity, density of woodpecker territories, density secondary cavity nesting species and the density of Neotropical migrants to be greater in plots that had snags than those where snags were removed (Lohr et al. 2002). Several studies have inves tigated the response of avian guilds to silviculture. For instance, ground nesters, cavity nesters, fl ycatchers, and long distance migrants were negatively associated with monotypic spruce plantations in Norway (Hausner et al. 2003). Insectivorous birds consistently exhi bit decreases in sele ctively logged forests relative to primary or unlogged forests (e.g., Thiollay 1992, Taylor and Haseler 1995, Mason 1996, Stratford and Stouffer 1999). More specifically, sallying and leaf gleaning insectivores have significantly lower species richness in se lectively logged forests while bark gleaning insectivores, frugivores and omnivores increase (Owiunji and Plumptre 1998). Finally, ground nesting species in no rthern Minnesota forests were most susceptible to nest-predation w ithin 100 meters of clear-cut areas, so much so, that for one species the edge habitat was a po pulation sink (Manolis et al. 2003). Urban/Residential/Roadways The fastest growing land use in Florida is urban and residential development. From 1949 to 1987 urban land area increased almost 500% (Kautz 1993). Currently 13.4 % of

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25 Florida’s land area is composed of roads or urban/residential developments (Vesterby and Krupa 2001). Amphibians and urban landscapes Very few studies explore amphibian species richness, composition, and or abundance in urban landscapes. However a pa rticularly relevant study by Delis et al. (1996) compared anuran species richness and abundance within wetlands surrounded by residential housing with wetlands surrounded by undeveloped land near Tampa, Florida. Species richness and abundances of anurans were greater in the wetlands surrounded by undeveloped land. Furthermore, anuran species richness and composition in the wetlands surrounded by undeveloped land were more si milar to a 1974 amphibian study conducted in the same area prior to the residential deve lopment. Delis et al. (1996) claimed that four species were sensitive to development because they were only found in the wetlands surrounded by undeveloped land, Bufo quercicus Scaphiopus holbrookii Hyla femoralis and H. gratiosa Three species, Rana sphenocephala R. grylio and R. catesbeiana had higher abundances in the wetlands surrounded by development. Changes to both the upland and wetland habitats in the urban envi ronment were thought to be responsible for the loss of anuran species at the developed sites (D elis et al. 1996). Life history traits of the extirpated (Sensitive) and the relatively more common (Tolerant) species may help to explain potential causal mechanisms for the observed differences. The four species Delis et al. (1996) found extirpated from wetlands surrounded by development were fossorial ( Bufo quercicus, Scaphiopus holbrooki and Hyla gratiosa ) and/or arboreal ( Hyla femoralis and H. gratiosa ). Changes in the surrounding terrestrial habitat may have reduced the available burrowing substrate for fossorial anurans. In a burrowing experiment, adult Scaphiopus holbrookii were unable

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26 to burrow into grass sod, and were much sl ower burrowing into la ndscaping gravel than sand. Juvenile Scaphiopus holbrookii did not burrow into comm on residential substrates such as sod, gravel, and satu rated soils (Jansen et al. 2001) Reduction in the amount of area affording refugia has the po tential to increase this speci es exposure to predators and the elements. A study in Italy of 84 wetlands within developed landscapes found that the wetlands with the highest amphibian species richness were free of fish, received high levels of insolation, and were near other wetlands that had high amphibian species richness (Ficetola and De Bernardi 2004). The most common species were pollution tolerant, did not require large patches of terrestrial habitat, and were capable of moving through human dominated landscapes by us ing hedgerows and canals. The most sensitive species appeared to be influen ced by a combination of wetland isolation and habitat alteration. In general the sensitive species were dependent on terrestrial habitats for part of their life cycle, highly susceptible to road tra ffic, and required water bodies free from fish predation (Fi cetola and De Bernardi 2004). Along a rural to urban gradient, a 10 km l ong 2 km wide transect in Connecticut, two amphibian species, Rana sylvatica and Ambystoma maculatum were found only in landscapes above 30% forest cover, Notophthalmus viridescens persisted only in landscapes above 50%, forest cover while two species, Plethodon cinereus and Pseudacris crucifer were found along the entire urba n gradient (Gibbs 1998). The species dispersal abilities were inversely corr elated to their persistence in fragmented habitats. That is, the most sedentary speci es persisted in habitat fragments and were found in the highest densities (Gibbs 1998).

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27 There are several potential mechanisms influencing amphibian species abundance and composition within the urban environment. Wetlands that remain in urban areas are often incorporated into the cities’ stormwater systems or utilized for tertiary wastewater treatment. Furthermore, very little of the native terrestrial habitat surrounding wetlands is allowed to persist, and road density, herbicide and fertilizer use all increase. Urbanized watersheds as well as wetla nds utilized for stormwater control experience highly fluctuating water depths, wh ich have been correlated to the amount of impervious surface in catchment basins (Richter and Azous 1995). Azous and Horner (1997) captured fewer amphibians in wetlands where water level fluctuation exceeded 20 cm. Egg masses attached to vegetation were suspected to be incapable of withstanding extreme water depth fluctuations. Heavy metals bioaccumulate in organisms using stormwater wetlands. Three species of fish ( Lepomis microlophus L. macrochirus and Micropterus salmoides ) collected from stormwater ponds in the Or lando, Florida area had significantly higher concentrations of heavy metals (Ag, Cd, Ni, Cu, Pb and Zn) than fish caught in natural lakes (Campbell 1995). Furthermore, some amphibians experience increased embryonic and larval mortality rates, slower developmental rates, reduced feeding rates, and lower breeding success when exposed to different heavy metals. These effects may be compounded under low pH conditions (Horne and Dunson 1994, 1995a,b,c, Rowe et al. 1996). Large amounts of water are used each day in urban areas. The resulting wastewater is often partially treated and then released into the environment. Ephemeral ponds within effluent-irrigated fields in Penns ylvania are used to treat wastewater high in

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28 chlorine, boron and nitrate. High boron con centrations were found to reduce hatching success of B. americanus and cause deformities in R. sylvatica A. jeffersonianum and A. maculatum (Laposata and Dunson 1998). When co mpared to natural temporary ponds the wastewater-irrigated pond amphibians ha d significantly fewer egg masses, lower hatching success and lower larval survival (Laposata and Dunson 2001). In Florida, Jetter and Harris (1976) noted initially high anuran use of cypress depression wetlands receiving wastewater. However, anuran us e nearly stopped as dissolved oxygen levels dropped with increasing inputs of waste. Birds and urban landscapes There have been numerous studies of avian composition, abundance, and species richness in urban environments throughout the world. However, studies of avian diversity in urban environmen ts within Florida or the Southeastern U.S. are limited. Woolfenden and Rohwer (1969) describe d the breeding birds within three study plots located in the suburbs of St. Petersburg and Gulfport, Florida. Older residential suburbs had three times as many breeding pair s than newer residential suburbs, 600 and 200 pairs per 100 acres respectively. The pr oportion of vegetative cover, a potential mechanism explaining the observed trend, in the new suburb was estimated to be 1.4% while that in the more mature suburbs were 28.3 and 10.3%. Th e breeding bird density in the mature Pinellas county suburbs were also twice the densities found in grassland, marsh, brush and scrub, oak hickory forests, beech maple forest, and mixed coniferousdeciduous forests (Udvardy 1957). A total of el even species were r ecorded breeding in the suburbs and the house sparrow, mourni ng dove, blue jay and northern mockingbird comprised 90% of all sightings (Woolfenden a nd Rohwer 1969). Based on earlier studies of native habitats that had been repla ced by suburban development Woolfenden and

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29 Rohwer (1969) listed 10 species that respond negatively to development; red-tailed hawk, red-shouldered hawk, bald eagle, yellow -billed cuckoo, chuck-will’s-widow, redcockaded woodpecker, scrub jay, brown-head ed nuthatch, pine warbler and Bachman’s sparrow. In a comparison of avian use of two forest ed riparian corridors, in a north Florida urban area and a state preserve, community and species specific differences were detected (Smith and Schaefer 1992). Densities of five insectivores were lower in the riparian forest surrounded by urban developmen t while the density of four omnivores and an insectivore increased (Smith and Schaefer 1992). In general, bird abundance and biom ass increases and species richness and community evenness decreases in an urban ma trix relative to less fragmented natural habitat areas (Woolfenden and Rohwer 1969, Emlen 1974, Mills et al. 1989, Blair 1996, Clergeau et al. 1998, Sorace 2002, Traut and Ho stetler 2004). There are many potential mechanisms responsible for this trend. Huma n supplied feeding stations were cited as a likely reason why Emlen (1974) recorded a 65-fold increase in herbivorous birds’ biomass in Tucson, Arizona relative to the surrounding native habitat. Species that do not maintain feeding territories and those species that utili ze human structures for nesting also tend to be more successful in urban environments (Emlen 1974). Studies focusing on the abundance of birds in urban environments have found that the population response is often species specific and vari es with the habitat qualities within the urban environment (Emlen 1974, Williamson and DeGraaf 1981, Mancke and Gavin 2000, Hostetler and Knowles-Yanez 2003) The population densities of 21 of 36 bird species were affected by the prox imity of buildings to study woodlots in

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30 Pennsylvania. The density of 10 species increased while 11 species decreased in woodlots adjacent to buildings (Mancke and Gavin 2000). Hostetler and Knowles-Yanez (2003) surveyed birds in older residential, newer residential, parkland, and golf courses in Phoenix, AZ, and found abundance of 4 of the 26 species encountered were significantly correlated with the amount of specific urban land use type. For instance, abundance of killdeer varied with golf course area, hous e sparrow with single lot residential and mourning dove with medium density residentia l. The remaining species that did not show a correlation with any specific land use type were thought to be associated with the unique vegetation characteristics found within the different study ar eas (Hostetler and Knowles-Yanez 2003). Avian species that share life histor y characteristics (i.e., guilds) have demonstrated population responses to urban land uses (e.g., Flather and Sauer 1996). The density of houses within 100 m of forest remnants has been shown to have a strong affect on Neotropical migrant birds while short-distance migrants and permanent residents showed no observable trend. Furthe rmore, urban 25 ha woodlots had a smaller Neotropical migrant community than 4 ha woodlots surrounded by lower intensity land uses but without nearby houses (Friesen et al. 1995). In New England, Kluza et al. (2000) also found Neotropical migrants and forest interior species responde d negatively to the density of hous ing in a forest matrix. More specifically, abundances of ground nesting species were greatest in forest areas with the lowest housing density while abundances of bl ue jays, a reported ne st predator, were greatest in areas of moderate housing densit ies. Vegetation characteristics were not significantly different between areas (Kluza et al. 2000). Nest predators were suspected

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31 to be responsible for the observed trends, and nest predation rate may increase as the land use intensity contrast between remnant fore st patches and adjacent land uses increases (e.g., Hanski et al. 1996, Suarez et al. 1997). Mills et al. (1989) found the densities and richness of na tive territorial birds were strongly correlated to the volum e of native vegetation in urban areas and to a lesser extent a negative correlation with housing density. Na tive territorial birds were not correlated with exotic vegetation volume. However, e xotic and nonterritorial birds were positively correlated with exotic vegetation volume and th e amount of lawn area (Mills et al. 1989). There is an assumed gradient in urban land use intensity from rural areas to the center of cities. Several studies have utilized the rural to urban gradient to examine the role humans have on avian communities (McDonnell and Pickett 1990). The urban gradient approach also enables researchers to identify, often species-specific, thresholds of human development on bird communities (M iller et al. 2001). Studies conducted in different regions of the world exploring the rural to urban gradient have indicated there are avian species that are urban avoiders urban adapters, an d urban exploiters (McKinney 2002). Blair (1996, 2001) ranked urban land areas from the least altered to the most altered by humans; biological preserve, recr eational area, golf course, residential neighborhood, office park and business district. One or more bird species reached their maximum density in each of the urban land uses sampled (Blair 1996). Native species occurred predominately in the lower intensit y land uses while generalists and exotic species were most common in the business dist ricts. The avian community similarity was much higher between the most heavily devel oped sites in Ohio a nd California then the

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32 similarity between native habitats in the two ecoregions (Blair 2001) Species richness, diversity, and biomass were gr eatest at sites of intermed iate development (Blair 1996, 2001). The community similarity between winteri ng birds in five town centers (30%) of Finland was lower than the similarity betw een apartment complexes (54%) and single family residential areas (54%)(Jokimaki a nd Kalsanlahti-Jokimaki 2003). A threshold was detected in the impact on the avia n community in towns between 35,000 and 105,000 residents, suggesting the regional land us e context has an affect on avian species composition within urban landscapes. Finall y, along with city size and urban land use type, the local habitat structur e is thought to be especially important in shaping avian community structure (Jokimaki and Kalsanlahti-Jokimaki 2003). Crooks et al. (2004) compared avian a bundance and richness in native habitat, fragments, and urban transects in southern California. Frag ments had the highest species richness and abundance. The significant vegetation differen ces between native and urban landscapes were suspected to be responsib le for the trends in avian populations. Furthermore, they detected species indicative of either urban or core habitat, and many of the urban species were also encountered by Bl air (2001) in Ohio and northern California. Different bird species reached peak abunda nce within forest fragments in Seattle surrounded by three different le vels of urbanization (urb an, suburban, and exurban) (Donnelly and Marzluff 2004). Most species associated with native forest habitat remained in fragments greater than 42 ha wh ile urban dependent species were associated with fragments surrounded by greater than 40% urban land cover. Urban dependent

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33 species were positively correlated and native bird species were negatively correlated with the amount exotic forb and shrub cover (Donnelly and Marzluff 2004). Avian species composition changed with th e number of native trees and shrubs, ground cover, and tree density within Colora do riparian lowland areas along a rural to urban gradient (Miller et al. 2003). Settlement intensity was negatively associated with the migrant and low nesting species, while resident and cavity ne sting species were positively associated urbanization (Miller et al. 2003). The species richness and density of birds in riparian forests of Californi a along an urban gradie nt decreased with increasing number of bridges within 500m, decreasing dist ance to nearest building, and decreasing volume of native vegetati on (Rottenborn 1999). The avian species composition was influenced by the same variab les with the addition of the proportion of building and pavement surface cover with in 500m. Foliage and bark gleaning insectivores, cavity nesters, and ground nes ting species all showed significant declines with increasing urbanization. Finally, a se t of tolerant and sensitive species to urbanization was identifi ed (Rottenborn 1999). Significantly different bird communities were also found in four urban habitats along a development continuum in Australia. Notably, insectivorous and nectivorous bird species declined along th e gradient from parks with na tive vegetation, to residential areas with native vegetation, to residentia l areas with exotic vegetation to newly developed areas denuded of most vegetation (White et al. 2004). Clergeau et al. (1998) observed a de crease in insectivorous species and an increase in omnivorous species on two rura l to urban gradient studies on different continents. Two types of bi rd groups capable of utilizing urban environments were

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34 identified. Omnivorous species adapte d to food resources supplied in urban environments and species that exploit resour ces found in the urban environment and their usual habitat. Several of the urban adapted species, Eu ropean starling, rock dove, and house sparrow were found in cities in Canada and Fran ce (Clergeau et al. 1998). A literature review by McKinney (2002) identified the traits of avian species groups indicative of areas along the rural to urban gradient. Species predominately found in urban centers, urban expl oiters or synanthropes, ar e often dependent on human resources, can feed on wind dispersed seeds or are omnivorous, and nest on rocky cliff like surfaces, or are cavity nesters capable of inhabiting human structures. Urban adapters are often considered edge species fo r their tendency to be located at the ecotone of different habitats. Urban adapters are pr edominately granivores, omnivores or aerial sweepers, and tree, shrub, or cavity nesters. Urban avoiders, often ca lled forest interior species, tend to be tree fora ging insectivores, neotropical migrants and or ground nesting species. Roads Roads and the vehicles that use them tr averse through many different landscapes and habitat types. The 3.8 million km of road s in the contiguous United States cover 1% of the land area, but it has been estimated that the ecological e ffect spans over 22% (Forman 2000). Roads have both direct and indirect affects on wild life populations. The direct affects include vehicular induced mortalit y and suitable wildlife habitat is lost with the construction of roads. Indirect affects include modi fication of animal behavior, habitat fragmentation, alteration of the surr ounding chemical and physical environment, corridors for the spread of exotics, and increased access by humans (see Trombulak and Frissell 2000).

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35 Road density within two km of wetlands in Canada was negatively correlated with the wetland’s plant, amphibian, reptile, and bird species ri chness (Findlay and Houlahan 1997). Furthermore, the correlation was stronge r when older road densities were used suggesting the full effect of road density on the species richness of amphibians and birds in wetlands can take decades (F indlay and Bourdages 2000). The probability of six different amphibian species being killed while attempting to cross a Denmark road with a traffic load of 134 vehicles/hour was estimated to be 0.34 to 0.61 and on roads with 625 vehicles/hour the probability rose to 0.89 to 0.98 (Hels and Buchwald 2001). For two of the species it was estimated that 10% of the population within 250 m of the road was killed annually (Hels and Buchwald 2001). Fahrig et al. (1995) found reduced amphibian abundance and density, and increased mortality with increasing traffic intensity. Finally, the probability that a moor frog ( Rana arvalis ) population is present in suitable habitat drops from 55% to 30% when the study area is adjacent to a highway (Vos and Chardon 1998). Road density is often correlated with ot her aspects or human development making it difficult to determine the independent e ffect of roads on wildlife populations. When the effects of isolated forest patch charac teristics were corrected for, avian species richness was lower in forest fragments with in 2 km of a major highway (Brotons and Herrando 2001). At the same time, forest sp ecies richness in forest fragments was negatively correlated to the distance to ro ads with higher traffic loads while some ubiquitous species showed a positive relationship with highways and residential roadways (Brotons and Herrando 2001). Sp ecies richness and re lative abundance of birds in Big Bend National Park, Texas was nega tively associated with distance to nearest

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36 road, distance to nearest development, and the interaction between these two variables (Gutzwiller and Barrow 2003). Finally, a signi ficant effect of traffic intensity was found on the avian population (reduced 12-54%) w ithin 100 m of roads with 5,000 cars a day and within 500 m for roads with 50,000 cars per day in ag ricultural landscapes without any other apparent land use influences (Reijnen et al. 1996). Studies on the Response of Amphibians and Birds to Multiple Land Use Types Studies of the wildlife response to lands capes of varying and multiple land use intensities indicate biodiversity is primarily affected by the proportion of land use types and distances between land use types of differe nt intensities. There appears to be a wildlife compositional gradient that roughly corresponds to a la nd use intensity gradient. Wildlife response to land use type is species specific, and often an individual species response is consistent between studies of si milar land uses. Where species response, as well as measures of biodiversity, seems to di ffer between studies of similar land uses (e.g., agriculture) the intensity of land use within the land use type is often the explanatory difference. Fina lly, there is also evidence s uggesting species that share certain life history characte ristics respond in a similar ma nner to habitat fragmentation, land use types and intensities (Whitcomb et al. 1981, O’Connell et al. 2000, Rodewald and Yahner 2001, Norris et al. 2003). Anuran species richness and abundance within Midwestern landscapes was negatively associated with the amount of urban land and positively associated with upland forest, wetland forest, a nd emergent wetlands. There was not a consistent trend with the amount of agricultural land. Anurans had a positive relations hip with agriculture in Wisconsin and a negative trend in Iowa The predominance of more intensive agriculture in Iowa (e.g., ro w crops) and, in general, le ss intensive agriculture in

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37 Wisconsin (e.g., pasture and hayfields) was o ffered as an explanation for the observed trends (Knutson et al. 1999). Amphibian species richness of ephemeral to semipermanent wetlands located in an agricultural (n=9) and urban (n =12) matrix in Minnesota wa s negatively associated with road density, distance to nearest wetland, a nd a very strong effect with the proportion of urban land adjacent to focal wetlands (L ehtinen et al. 1999). The proportion of agricultural land was not selected as a predictor of species richness. Species thought to be particularly sensitive to urban influences included Notophthalmus viridescens and Hyla chrysoscelis The American toad ( B. americanus ) in urban landscapes was positively correlated with the amount of forest cover (Lehtinen et al. 1999) while Guerry and Hunter (2002) found a negative association with forest cover for this species in an agricultural landscape. Anurans of the New Jersey Pine Barrens appear to respond in a predictable manner to agricultural and urban development surr ounding breeding wetlands. Two species were considered pineland specialists ( Hyla andersonii and Rana virgatipes ), four species were wide ranging throughout New Jersey (e.g., Rana sphenocephala Rana clamitans and Bufo fowleri ), and four species were considered bo rder entrant or thos e only capable of entering the pineland community in habitats disturbed by humans (e.g., Hyla versicolor and Acris crepitans ). The anuran community gradient corresponded to an environmental gradient. An increase in pH and a d ecrease in specific conductance and floating vegetation were associated with increasi ng human disturbance and amphibian species composition (Bunnell and Zampella 1999).

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38 The occurrence of 10 of 13 amphibians in 74 Ontario wetlands was positively correlated with the proportion of forest cover and wetlands on adjacent lands and negatively correlated with road density and nut rient levels (Houlahan and Findlay 2003). The effects of adjacent land uses on amphibian composition were strongest within 200m of the focal wetland. Bufo americanus was the only species positively correlated with land use intensity and nutrients levels. All other species we re negatively correlated with water nutrient levels. Two species, Rana clamitans and R. catesbeiana did not show an association with land use in tensity. Although the proportion of surrounding land use in either agriculture or urban was compared with amphibian species richness, abundance, and species frequency of occurrence the propor tion of forest cover was a better predictor in all models (Houlahan and Findlay 2003). The proportion of forest cover may be a surrogate measure of land use intens ity because the replacement land use was conglomeration of often high intensity land uses (e.g., cropland, highways, urban areas). A multi taxa study in Minnesota observed bird richness and diversity generally decreased with increasing ag riculture surrounding 15 ripari an wetlands in Minnesota. Amphibian abundance increased with the pr oportion of wetlands and decreased with increasing rangeland and urban land uses. Amph ibians were also influenced by specific wetland characteristics, such as ditching or eutrophication. Land uses within 500 and 1000 m of the wetland, the shortest distances studied, had the strongest relationship with amphibian and bird divers ity (Mensing et al. 1998). Landscape scale studies of avian compos ition indicate bird communities respond to the patch area size of forest or wetland cover, proportion of forest cover, and edge density

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39 (Miller et al. 1997, Brennan and Schnell 2005). Yet the effect of land use intensity with in the nonforested land uses on adjacent fo rested habitats is little understood. Determining the biological effect of land us es of different land use intensities and energy flows could have important implications in land use planning. For instance, in the American west there has been some debate as to whether ranch lands or residential development has a lower impact on native bi odiversity (see Knight et al. 1995 and Wuerthner 1994). Maestas et al. (2003) compared avian, mesopredator, and plant communities across a gradient from reserves to ranches to exurban development. Seven bird species had higher dens ities in residential areas while two species had higher densities on ranches, one species on reserv es, and three species on both ranches and reserves. The plant and mesopredator comm unities were more similar on reserves and ranches than residential developments. Bird community response appears to be co rrelated to assumed anthropogenic land use intensity gradients. The species abunda nces and composition of a subandean bird community in continuous native forest was mo re similar in forest fragments surrounded by lower intensity exotic tree plantations th an pastures (Renjifo 2001). Avian community response was species-specific and independe nt of trophic group and foraging strata (Renjifo 2001). The more heterogeneous vege tation structure of th e plantations versus pastures was hypothesized as the mechanism for the gradient in bird community composition. Gradual changes in avian spec ies composition along a land use intensity gradient (primary forest, secondary forest, agroforestry, annual cr ops) have also been observed in Indonesia (Waltert et al. 2004). In Pennsylvani a long distance migrants and forest dependant species were more common and edge species less common in forest

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40 remnants surrounded by silviculture versus those surrounded by agriculture (Rodewald and Yahner 2001). Finally, a summary of 34 stud ies in tropical fore sts indicated species richness of birds, ants, and lepidoptera decrea sed in areas converted to agriculture but not in areas that were selec tively logged (Dunn 2004). Rodewald and Yahner (2001) compared fo rested landscapes within either an agriculture or silviculture matrix and concl uded that the type of disturbance strongly influenced the breeding bird community. In general, more forest species declined and edge species increased in forested lands capes disturbed by agriculture relative to silviculture disturbances. Cavity nesters, resident species, downy woodpecker, American crow, indigo bunting, and brown-headed cowb ird were significantly more abundant in forest with adjacent agricultural land uses. Long-distance migrants, forest-canopy nesters, forest-understory nesters, yellow-billed cuckoo, and hooded warbler were more abundant in forests adjacent to silvicultu ral treatments (Rodewald and Yahner 2001). Norris et al. (2003) explored the avian community composition of forests along a land use intensity gradient from mature, slightly disturbed, su ccessional, and highly disturbed (i.e., selectively logged and grazed) forests in Iowa. A vegetation disturbance gradient was used to rank sites and was esse ntially an index of land use intensity. Bird species richness and abundan ce were not good predictors of the forest disturbance gradient. However, species groups responde d to the forest disturbance gradient. Neotropical migrants, species of management concern, and area-sensitive species were most common in forests of low land use in tensity while permanent residents, short distance migrants, and ground nesters were more abundant in forests that had higher land use intensity. Although only forests were st udied, habitat heterogeneity between forest

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41 types was implicated as the driving mechanis m for avian community response (Norris et al. 2003). The results of these studies indicate that there is a response of amphibians and bird communities to general land uses and in most landscapes there appears to be a gradient from natural to silviculture to agriculture to urban. Habitat Fragmentation In many landscapes, isolated patches of natural habitat remain in a mosaic of human land uses. Yet, our understanding of the effects on wildlife populations in fragments is incomplete. There has been limited success applying the Equilibrium Theory of Island Biogeography to natura l habitats surrounded by human dominated landscapes (Wiens 1995). Similarly, studi es of fragmentation have had difficulty establishing trends between habitat pattern and biodiver sity (e.g., McGargial and McComb 1995, Lehtinen et al. 1999, Hostetler and Holling 2001). These approaches appear to be inadequate in th eir ability to accurately describe observed trends in species composition, abundance, and persistence in frag mented habitat patches. There are two main factors influencing wild life populations in fragmented habitat patches not accounted for by the island biogeography theory or tr aditional fragmentation studies. The surrounding landscape type influences both specie s-specific dispersal capabilities and the ecological quality of the habitat patch. The focus cannot only be on habitat pattern, but needs to include the dynamics of the lands cape mosaic (Wiens 1995, Rodewald 2003). The species that respond (i.e., increase or decrease) to habitat fragmentation may be evidence that environmental change and species dispersal capabilities are important determinants of wildlife occupation of ha bitat patches within human influenced landscapes. In general, species with poor disp ersal capabilities, sedentary species (Wiens

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42 1995, Walter et al. 1999), and habitat specialis ts (Canaday 1996, Ford et al. 2001, Craig 2002) tend to be the most sensitive to fragme ntation. Similarly, wide ranging species and habitat generalists tend to thrive in fragmente d landscapes (Miller et al. 1997, Ford et al. 2001, Craig 2002). From a species perspective habitat qual ity and dispersal capab ilities are important factors influencing species composition in isolated habitat patches or fragmented landscapes. In turn, landscape composition and configuration contribut e to the degree at which a fragment is isolated and suitable. The composition or intensity of human land use surrounding habitat fragments can be highly variable. Lands capes that receive periodic or low intensity anthropogenic inputs (e.g., selective logging, na tive rangeland) appear to be the least restrictive to wildlife movement (McGargial and McComb 1995, Wigley and Roberts 1997). Not only are some species more likely to disperse th rough low intensity land us es there is often a subset of the original species persisting in the “degraded” habitat. These residual and peripheral species have the ab ility to influence species recruitment, predator-prey dynamics, and resource competition within natu ral habitat patches (Ford et al. 2001). Even high intensity anthropogenic landscapes are not barriers to dispersal for some species (e.g., Mazerolle 2001, Bulger et al. 2003) Again, it is a combination of a species behavior and the composition of the surrounding landscape th at determines dispersal permeability. Inhibiting dispersal is not the only in fluence adjacent landscapes can have on the biota of habitat patches. Juxt aposed habitats interchange ab iotic and biotic components. Habitat patches are often influenced by th e neighboring landscape reducing habitat

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43 quality for some species (e.g., Debinski a nd Holt 2000, Davidson et al. 2002). There may be a gradient of environmental change with increasing human use of the surrounding landscape (e.g., Blair 1996, Findl ay and Houlahan 1997). However, quantifying human land use and environmental degradation is di fficult. Common trends in habitat patches adjacent to human landscapes are increased nut rient inputs, reduced fire periodicity, and “edge effects”. The intensity of these eff ects is highly variable and dependent on the landscape mosaic. Wildlife populations are influenced by th e synergistic effect s of habitat loss, habitat fragmentation, and habitat degradation (Ford et al. 2001). Fragmentation studies that do not account for the intensity of human disturbance may lead to an oversimplification or oversta ting of observed trends. Esta blishing disturbance trends with human land use types and land use in tensities may increase the accuracy of predicting the effects of habitat fragmentation on wildlife populations. Amphibian and Avian Assemblages as Indicators of Land Use Intensity Amphibian and bird assemblages have been used as an index of human habitat disturbance. Species richness and diversity vary inconsisten tly with habitat disturbance (e.g., Brooks et al. 1998), however, species co mposition and guild assemblages appear to offer more potential for quantifying the biotic effect of land use intensity and habitat degradation on wildlife communities (Brooks et al. 1998, O’Connell 2000). An evaluation of the amphibian, bird, fi sh, invertebrate, and plant community response to the proportion agriculture, urba n, grassland, forest, and water circumjacent 116 Minnesota wetlands revealed 79 metric s significantly correlated to land use (Galatowitsch et al. 1999). Metrics potent ially applicable to other wetland systems included: proportion of wetland birds, wetland bird richness, proporti on of insectivorous

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44 birds, importance of Carex and importance of invasive perennials. Only 2 amphibian and 23 bird metrics were signifi cantly correlated to land use and possibly an artifact of their overall diversity. Of all the taxa studied birds were c onsidered the most useful and amphibians the least useful for monitoring changes in wetlands related to land use (Galatowitsch et al. 1999). Another north temperate study also f ound amphibians could not be used as the only indicator of biot ic integrity for ephemeral ponds and wetlands. Amphibians, crayfish and fish assembla ges were needed to accurately detect anthropogenic disturbance gradient in Indiana (Simon et al. 2000). O’Connell et al. (1998, 2000) developed an index of biological integrity for the uplands of the Mid-Atlantic States using songbird response guilds. From literature searches songbirds were typed either specia list or generalist for eight guild categories; trophic status, foraging substrate, nest placem ent, primary habitat, patch size, number of broods, migratory status, and nest predators (O’Connell et al. 1998). The proportion of the songbird community that was omnivore, inse ctivore, single-brood species, and forest area-sensitive species was significantly diffe rent for six levels of habitat integrity (O’Connell et al. 1998). Sites that were typed as being of poor quality by the songbird IBI were dominated by agricultu re and urban land uses (O’Connell et al. 2000). Finally, the proportion of forest, landscapelevel diversity, and canopy he ight were selected as the best predictors of the songbi rd IBI (O’Connell et al. 2000). Similarly, Croonquist and Brooks (1991) ranke d species into avian response guilds and compared a protected watershed with a highly developed watershed. Response guild scores were significantly different between the two waters heds. As the proportion of area developed increased from the headwaters to the mouth in the impacted watershed the

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45 response guild scores decreased. For exampl e, the proportion of Neotropical migrants decreased and the proportion of edge and exotic species in creased with increasing land use intensity. At the same time, mammalia n guilds did not respond to habitat disturbance in a predictable manner (Croonquist and Brooks 1991). It is increasingly apparent that multiple taxa assemblages are better at quantifying human habitat disturbance. Se veral studies have indicated th at there is often not crosstaxon congruence with response to habitat dist urbance, especially when species richness is compared (e.g., Lawton et al. 1998, Lund and Rahbek 2002, Provencher et al. 2003, Su 2004). Landscape Development Intensity (LDI) Index Although a comparison of biodiversity betw een discrete land use types within a region may reduce variability and increase in ference on the effect of land use type on biodiversity, the typical lands cape is often a patchwork of individual landowners, land use types, and land use intensities. Thus the wildlife assemblage of a region is often influenced by a variety of land use intensi ties, and studies that reduce the inherent variability of landscapes may be reducing realism and applicab ility of study results. The question then becomes how to study the respon se of wildlife composition in the “real” world but yet have a quantifiable land use intensity to make comparisons with other studies. The Landscape Devel opment Intensity (LDI) index offers a potential mechanism for quantifying the human influence to landscapes (Brown and Vivas 2005). The LDI index is a measure of development intensity based on the energy use per unit area (Brown and Vivas 2005). Accounting for all of the anthropogenically derived materials and energies within land uses may not seem per tinent to studies of wildlife composition. However, with increasing human energy use there appears to be a gradient

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46 of direct, indirect, and cumula tive impacts, which have the ability to influence wildlife abundance, composition, behavior, distribution, and survival. The LDI index quantifies the human land use gradient which is corre lated with increasing noise pollution, light pollution, human presence, traffic noise, and hum an structures all of which are examples of factors that have the abili ty to affect wildlife (e.g., Bau tista et al. 2004, Bird et al. 2004, Crawford and Engstrom 2001, Klem et al. 2004, Longcore and Rich 2004, Riffel et al. 1996, Slabbekoorn and Peet 2003). Tradit ional studies of w ildlife response to forest/nonforest situations that do not account for the intensity of human land use within the matrix may be missing an aspect importa nt in shaping wildlife assemblages.

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47 CHAPTER 2 METHODS Plan of Study The methods employed in this study of wildlife response to land use intensity included three years of field da ta collection of a number of biotic and abiotic parameters in 111 isolated forested wetlands and data anal ysis using several sta tistical techniques. The following methods section is organized in to three subsections: site selection, field data collection, and statistical analysis methods. Site Selection This study concentrated on small isolate d, depressional, forested wetlands in Florida. Wetlands varied in size from 0.1 to 2.1ha (Table 2-1). Site selection was conducted to insure that approximately one quarter of the total 111 study sites were sampled within each of Florida’s four wetla nd ecoregions (Lane 2000)(Figure 2-1). In addition, wetlands were selected th roughout Florida to represent four a priori landscape settings; reference, silviculture, agriculture and urban/residential. Because of the difficulty of obtaining access to wetlands and th e fact that depressional forested wetlands are not randomly distributed throughout Florida, selection of sample wetlands was not conducted randomly.

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48 Table 2-1. Mean wetland size (ha) and the minimum and maximum wetlands sampled per land-use category and region. M eans were not significantly different between land use types or Florida re gion (one way ANOVA, Tukey multiple comparisons, land use types (p=0.266) and between regions (p=0.066)). mean (s.d.) min max Reference 0.66 (0.41) 0.20 1.75 Silviculture 0.49 (0.24) 0.22 0.96 Agriculture 0.81 (0.46) 0.12 2.11 Urban 0.74 (0.48) 0.10 2.03 Panhandle 0.60 (0.38) 0.10 1.75 North 0.73 (0.38) 0.26 1.85 Central 0.88 (0.52) 0.12 2.03 South 0.61 (0.44) 0.18 2.11 Figure 2-1. The distribution of the surv ey sites by land use and Florida region.

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49 Table 2-2 lists the number of wetlands sampled in each of the regions by a priori landscape matrix. Twenty-seven wetlands were sampled in the panhandle, thirty in the northern peninsula, twenty-nine in the central region, and twenty-five in southern Florida. Twenty-seven study wetlands were located with in predominately agri cultural landscapes, thirty-three were in reference landscapes, eleven wetlands were in or adjacent to pine plantations, and forty wetlands were located within an urban/residential matrix. Table 2-2. Count of 2003 sites by land-use and Florida region. Reference Silviculture Agriculture Urban Total Panhandle 7 4 6 10 27 North 7 7 6 10 30 Central 11 0 8 10 29 South 8 0 7 10 25 Total 33 11 27 40 111 Access to wetlands was gained by cont acting landowners and land managers directly and with the aid of ag ricultural extension agents. Si tes were grouped into blocks of 4 to 6 sites incorporating at least one low intensity site (e.g., reference or silviculture) with higher intensity sites (e.g., agriculture and urban). Each wetland within a site block (four to six sites) was sampled twice within a four-day period (genera lly two to four sites were sampled per day) for a tota l of 111 wetlands sampled between 30 May and 27 August 2003. Wetlands within th e site blocks were sampled to minimize driving time regardless of land use type. Site order was reversed for the second site visits when possible. An attempt was made to sample site blocks alternately among the four Florida eco-regions throughout the 2003 sampling period. Figure 2-2 shows the mean proportion of land cover type within 200 meters of each focal wetland for the four a priori land use categories. The mean proportion of undeveloped land comprised greater than 20% of the land area within 200 meters of the

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50 agriculture, silviculture and urban land use categories, and greater than 97% of the area for reference wetlands. Agriculture, silvicul ture, and urban sites all had means greater than 70% of the land cover w ithin 200 meters that define s each category. 0.0 0.2 0.4 0.6 0.8 1.0 RSAU Sites by land useProportion of 200m buffer Undeveloped Pine Plantation Agriculture Urban Figure 2-2. The mean proportion of land cover type within 200 mete rs of each focal wetland by land use categories. R=refe rence sites, S=silviculture sites A=agriculture sites, and U=Ur ban/residential sites. The dominant land use surrounding 13 of the 27 agriculture sites was cattle pasture. Five sites had row crops or spray fields a nd three sites had citrus plantations as the dominant surrounding agricultural land use. Six a priori agriculture sites consisted of a combination of pasture and row crops. Sixteen of the 19 sites that had cattle pasture had improved pasture (e.g., evidence of planted non -native grasses, drainage ditches and canals, clear-cut forests etc.) and the rema ining three were within native rangeland. Among the reference sites two wetlands were in city parks, four were in county parks and seven were in state parks, preser ves or reserves. Two sites were on military bases, two were in national fore sts and three were in national w ildlife refuges or reserves. Finally, six sites were within state forest, three were on private c onservation tracts, and four were on water management district properties. Approximately 20 of the 33

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51 reference sites allowed game hunting and sele ctive timber harvest operations within the landscape surrounding the study sites. Five of the 11 sites categorized as silv iculture sites were adjacent to pine plantations within state forests and one was in a national forest. The remaining five silviculture sites were on private timber company land. The dominant urban land use surrounding twelve of the 40 a priori urban sites was categorized as industrial, comme rcial, or institutional and four of these also had a large proportion of the surrounding landscape as arbo retum or silviculture. Twenty-two urban sites were adjacent to primarily residential land uses of varying human densities, and 6 sites were a combination of residential and commercial or institutional. Four of the 22 residential sites also had gol f courses as a dominant surr ounding land use, and finally, two of the residential sites had a large proportion of city park or maintained open space within 100 meters. Field Sampling Sampling of amphibian and avian species was conducted within the limits of each focal wetland and the circumjacent landscape within 200 m. Habitats adjacent to the focal wetland were included in the survey because many amphibians are dependent on uplands during part of their life cycle and only use wetlands seasonally. Furthermore, the relatively small size of the focal wetlands in this study may only be a component of some avian species home range (Schoener 1968), and re stricting species detections to just the wetland during a two-visit synoptic sample c ould underestimate the wildlife use of the habitat. Two hundred meters was chosen becau se it encompasses the area of most avian territory sizes (Schoe ner 1968), and is the maximum radius of audible detectability for most eastern forest bird species (Whitcom b et al. 1981). Two hundred meters from the

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52 wetland edge also includes the mean minimu m “core” terrestrial habitat area for most amphibians (Semlitsch and Bodie 2003). All site surveys were conducted by the same observer. Survey start and end times were recorded for each survey and surveys were started anytime during the daytime hours. The air temperature, cloud cover (weather bureau sky condition codes), and estimated wind speed (Beaufort land based scale) were recorded at the beginning of every survey. Surveys were not started during heavy rains. Visual and Auditory Encounter Surveys Each survey began by walking a 15-mi nute transect through the wetland that started five meters from the wetland edge at the closest physical access point to the wetland. The surveyor slowly walked an appr oximate straight line through the center of each wetland stopping periodically to listen and visually scan the surroundings. One minute was spent surveying at opposite side s of the wetland while standing in the wetland/upland ecotone. Species flying above the sites that were suspected to be foraging (e.g., swallows, vultures, etc.) or flee ing from the site were included. After the completion of the 15-minute transect, sp ecies observed opportunistically while conducting other onsite measurements and sa mpling were also recorded. The time elapsed from the survey inception was recorded for the first observation of each species detected in the wetland and within the landscape, as well as how the species was first detected (song, call, or visual, and for birds if it was flying). Logs, trash, and other cover objects within the wetland bounda ry were turned when possi ble to detect sheltering species.

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53 Automated Recorder Surveys An automated recording system (Bedford Technical model ARS-o6X001) was deployed at 96 wetlands during the first site visit. To protect the microphones from inclement weather they were placed inside an inverted plastic water bottle about 2.5cm back from the cut off bottom end (Dodd 2003) The microphone and plastic cover were tied to a branch with fishing line, at a slig ht downward angle, suspended 1.5 to 2m above ground level, and directed at the deepest area in the center of each wetland. Leaves and small branches within one meter of the microphone were removed. The timer on each automated recorder was set to record for one minute every hour for a total time of approximately 45 hours. The same observer listened to all of the cassette tape samples in the lab with a headset. All species heard singing or calling and the time of day were recorded for each one-minute acoustic sample and the maximu m number of anurans heard calling was coded by species; 1= one individual heard, 2 = more then one heard but individuals distinct, 3 = multiple calls but individuals ca nnot be distinguished. Finally, notes were made of all identifiable noises (e.g., airplanes, automobile traffic, fireworks, rain, wind, etc.) and whether the noises obscured the de tection of calling vert ebrates during each one-minute acoustic sample. Combined Sampling Effort Two 15-minute transect surveys were conduc ted at all 111 sites in 2003 for a total of 3,330 standardized minutes. An additiona l 7,753 minutes were spent at the sites documenting species observations beyond the 15-minute time period along with 3,762 minutes of audible cassette tape recordings, for all sites combined. The mean amount of total minutes spent recording species compos ition, including audio re corder minutes, at

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54 each agriculture, reference, silviculture, and urban site was 142.3, 145.6, 122.2, and 121.3 minutes respectively. The amount of time spent sampling reference sites was significantly greater than at ur ban sites. There was not a significant difference in the amount of time spent surveying at all other la nd use types. Audio loggers were used at 24 agriculture sites, 25 referen ce sites, 9 silviculture sites and 34 urban sites, 89%, 76%, 82%, and 85% of the sites resp ectively. The sites at which cassette recordings were taken, and the amount of audible minutes for each site, are given in Appendix A. Dip-net Sampling At sites that had standing water twenty dip-net sweeps were used to sample for larval anurans and salamanders during one of th e two visits to each si te. The dip-net was a U.S. standard 30 mesh D-frame net and each sweep was approximately 1.0m long. The twenty dip-net sweeps were stratified acr oss each wetland, perpendicular to the 15minute survey transect, targeting representa tive microhabitats. Caudates and larval anurans (tadpoles) were identi fied, tallied, and released af ter each sweep (Conant and Collins 1991, Altig et al. 1999). Some specime ns could not be reliably identified to species in the field (e.g., Rana sphenocephala, R. capito, Rana grylio, Hyla squirella, Bufo terrestris and some very small individual s) and were given a unique unknown grouping (e.g., unknown a, unknown b etc.), tallie d and released. All fish captured during the sweeps were identified to genus and species when possible, tallied, and released. The presence of know n macroinvertebrate tadpole pred ators were also recorded but not tallied (e.g., Belostomatidae, Decapods Dolomedes sp., Dytiscidae, Gyrinidae, Nepidae, Notonectidae, Odonata).

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55 Terrestrial Insect Sampling A 38 cm diameter sailcloth sweep net was used to sample terrestrial macroinvertebrates during one of the site vis its at 82 of the site s by walking across each wetland and sweeping the net in an approxi mate180 degree arc across shrubs and forbs 100 times. The collected insects and vegetation were placed in a site -specific plastic bag then put in an ethyl acetate killjar with th e bag left open. Each site bag was removed from the killjar after 20 minutes, placed on i ce and/or refrigerated until they could be processed. Processing was conducted in the lab and involved separating all macroinvertebrates visible without the aid of microscopes or hand lenses from pieces of vegetation. Macroinvertebrat es were identified to order, grouped by species, and counted. Each macroinvertebrate order and collected vegetation was weighed separately to the nearest milligram. Other Biotic Sampling Prior to the sampling for amphibian and av ian fauna in 2003 the sites were sampled in either 2001 or 2002 for macrophytes, aqua tic macroinvertebrates, diatoms, tree composition, percent canopy cover and basal area (Reiss 2004). As part of the Reiss (2004) study several metrics using macr ophytes were calculated for the wetlands including macrophytes sensitive or tolera nt to development, the proportion of macrophytes native or exotic to Florida, th e proportion of macrophyte species that were wetland dependent, and an average coeffi cient of conservatism score based on macrophyte presence were calculated for each site (Reiss 2004). Reiss (2004) summarized these metrics along w ith macroinvertebrate and diat om metrics into a Florida Wetland Condition Index (FWCI) score for each s ite. In that same study each site was

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56 given an average Wetland Rapid Assessment Procedure (WRAP) score using methods outlined by Miller and Boyd (1999). Environmental Variables Water Sampling A grab water sample was taken from an undisturbed area of each wetland that had standing water at least 10cm deep following the standard operating procedure of the Florida Department of Environmental Protec tion (FDEP) for water quality measurements. A hand held YSI-55 Dissolved Oxygen meter was used to measure dissolved oxygen and water temperature in the same general locati on immediately prior to each water sample. Water samples were sent to the FDEP Cent ral Chemistry Laboratory, and analyzed for ammonia-nitrogen (EPA 350.1), color (EPA 110.2), nitrate/nitritenitrogen (EPA 353.2), pH (150.1), specific conductance (EPA 120.1), total Kjeldahl nitrogen (EPA 351.2), total phosphorus (EPA 365.4), and turbidity (EPA 180.1). The maximum water depth at each site was measured by walking across each wetland, and recording the water depth in th e deepest area encountered to the nearest centimeter. For the vast majority of the site s the deepest area was near the center of the wetland. Maximum depth measurements were not taken in areas that were obviously animal burrows, tree tip-ups, ditches, canals, or excavated pools. Fire History During the site visits evidence of any fi re in the surrounding landscape as well as within the sample wetland was recorded. Fire evidence included charred, stumps, snags, and logs, or burn marks on the boles of livi ng trees. An estimate was made about whether the fire had occurred more or less th an 6 years ago. This was a gestalt estimate based on vegetation height and composition as well as character istics of the charring. In

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57 some instances landowners or land managers were able to date the last fire in the area. A sample wetland and surrounding landscape were sc ored separately as follows, zero for no evidence, one for old evidence (>6 years ago), and two for recent fire evidence (<6 years ago). Landscape Variables Several landscape indices were calculat ed from Land Use/Land Classification (LU/LC) coverages, aerial photos, and site vi sits as follows: 1) Landscape Development Intensity (LDI) index using 100m buffer ar ound each sample wetland and current LU/LC, 2) a historical 200m LDI using 1974 LU/LC data, 3) the percent of the 200m buffer consisting of the following land types; undeve loped land, pine plantation, agriculture, urban/residential, roadway, wetland, conti guous forest, and canal/retention pond, 4) distance to nearest paved road and two closest wetlands. Landscape Development Intensity Index A 100m buffer LDI score was calculated w ith ArcView software (ESRI 1995) for all of the sample wetlands (Brown and Vivas 2005). The LDI index is intended to represent the amount of human impact on sy stems. The assumption is the amount of human energy use in the landscape is associated with the amount of change to adjacent systems through direct, secondary, and cumulativ e affects. The LDI score for each site was calculated as follows. LDITotal= %LUi LDIi LDITotal = LDI ranking for 100 meter buffer around each forested wetland site %LUi = Proportion of the 100 meter buffer in land cover/use i LDIi = Landscape development intensity coefficient for land use i

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58 The LDI coefficients for each land cover/use (FDOT 1999, FDEP 2002) are derived from previous studies of ener gy use within common land use systems and represent the amount of nonrenewable energy (e .g., electricity, fuels, fertilizers, and pesticides) used per unit area (Brown and Vi vas 2005). The range of coefficients has been normalized on a scale from 1 to 10. A land use coefficient of 1 equals no nonrenewable energy use (i.e., natural system) and land use coefficient of 10 is the most intensive land use (i.e., central business district that has a av erage height of 4 stories). For comparative purposes other typically enc ountered land cover/uses encountered in this study and their LDI coefficients are pine plantation (1.58), na tive rangeland with livestock (2.02), improved pasture with live stock (3.74), row crops (4.54), low density single family residential (6.90), high density single family resident ial (7.55), two lane highway (7.81), four lane highway (8.28), a nd commercial (9.18)(a more complete list can be found in Brown and Vivas 2005). Digital Orthophoto Quads (DOQ) from 1999 (LABINS 2002) and site visits in 2001 and 2002 were used to classify land cover/ use within 200 meters of each wetland. An LDI score based on the 1974 land use c overage (FDEP 2002) within 200m of each site was also calculated. Landscape indices In addition to the LDI calc ulations several other indices of landscape context were tabulated within the 200m buffer of each wetlands. These included the proportion of land area comprised of the buffer and the sample wetland that was agriculture, undeveloped, silviculture, urban, roadway, canal/retenti on pond, and wetland. The proportion of each land use was calculated using ArcView software (ESRI 1995, FDEP 2002), 1999 DOQs, and site visits. A 200 m buffe r was used because this was the area of biotic sampling,

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59 and it was shown by Houlahan and Findlay (20 03) that land uses within 200m have the strongest effect on amphibian assemblages. Land areas with very little to no annu al human derived energy inputs were classified as undeveloped (e.g., pine flatwoods, dry prairies, hardwood hammocks, wooded vacant lots). Land areas categorized as agriculture were primarily used for growing crops and or raising domesticated anim als. It was often difficult to determine the livestock grazing intensity in forests a nd wetlands, especially from aerial photos, and these areas were designated undeveloped, as were, woodlots in urban and agricultural areas. Silviculture consisted of clear-cuts and pine planta tions, however selectively cut forests were included with the proportion undeve loped. Golf courses, recreational fields, cleared lots, buildings, and road s along with their associated “green space” (i.e., yards and right-of-ways) were classified as urban. Many of the classifications were not mutually exclusive. For example, the proportion of the 200 m buffer considered road way was also included in the proportion classified as urban. However, single track “jeep trails” or forest roads that were on the original soil substrate and were not graded were included with the land use on which they occurred (e.g., agriculture or reference). The proportion of the total area classified as wetland was calculated, and included in th e proportion undevelope d. Canal/retention pond area was calculated alone but also included within th e land use in which they occurred (e.g., silviculture, agriculture, urban). The variable “% contiguous forest (200m)” was the area of the sample forested wetland plus the forest adjacent to the focal wetland as a proportion of the combined wetland and 200 m buffer area. For example only the forested portion of the focal

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60 wetland would be contiguous forest at sites surrounded by wet or dry prairie. Citrus groves, forested fence lines, and pine plantatio ns greater than 5 year s from planting were included as contiguous forest area if adj acent or continuous with the focal wetland. Forests were considered areas with greater then 25% canopy cover and breaks in the forest canopy greater then 25m were cons idered discontinuities of the forest canopy. The distances from the focal wetland to the nearest wetland, the second nearest wetland, paved road, major river system (FDE P 2002), and forest cover were measured using ArcView software (ESRI 1995). Dist ance to the two closest wetlands was measured from the focal wetland perimeter to the closest edge of any type of wetland delineated in FLUCCS code. Distance to paved road was measured from the focal wetland edge to the closest paved road usi ng 1999 DOQs. There is some indication that major floodplain forests may be wildlife corrid ors, thus the distance to the nearest river system from the focal wetland edge was meas ured using an FDEP (2002) GIS coverage of Florida rivers. Finally, the distance to th e nearest forest patch from the focal wetland edge was measured for sights w ithout contiguous upland forest. Data Analysis First, comparisons were made betwee n land use types using the measured environmental and landscape variables. S econd, associations between LDI scores and landscape and environmental variables are pr esented. Then relationships between the species assemblages and landscape content a nd environmental variables were explored looking for possible explanati ons for observed trends in amphibian and avian species composition. Finally, characterist ics of the species life history traits were compared with environmental variables, land us e type, and land use intensity.

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61 Water Chemistry and Landscape Variables Environmental variables were measured and then compared for significant differences between land use types and LDI qua rtiles (sites arranged into approximately four equal sized groups and ranked based on LDI scores; LDI quartile one consisted of the sites with the lowest LDI scores and LDI quartile four consisted of the sites with the highest LDI scores). A ll variables were tested for normality, Kolmogorov-Smirnov Normality Test, and transformed when necessary, (power, logarithmic, and arcsine squareroot transformations). If normality was still not met, non-parametric tests were conducted, and parametric tests on normally di stributed data. Univariate correlations were used to compare the site LDI scores with all other environmental and site sampling variables. Species Richness Presence or absence of amphibian and avian species were recorded for each site. The mean species richness, total species richness, and the minimum and maximum number of species observed at the sites fo r each taxon were compared by land use type, LDI quartile, and Florida region. Visual estimations of species accumulation curves (species richness by time) suggested that the majority of the avian and amphibian species pool was sampled at each site. Amphibian frequency of site occurrence wa s calculated by land use type and for all sites. Amphibian species frequency of occurre nce is equal to the num ber of sites species X was found by land use divided by the number of sites in Florida counties within species X’s range (ARMI 2004) by land use type.

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62 Species Composition To test for differences in the species composition of the a priori land use types and LDI quartiles multi-response permutation pr ocedure (MRPP) was used in PC-ORD (McCune and Mefford 1999). MRPP is a nonpara metric test of differences between groups (Mielke 1984). The Sorensen (Bray-Cu rtis) distance measure was selected to calculate the distance matrix used in the MRPP analyses (McCune and Grace 2002). Amphibian and avian community struct ure were ordinated using nonmetric multidimensional scaling (NMS)( Kruskal 1964, Mather 1976) with PC-ORD software (McCune and Medfford 1999). Ordination is used to reveal dominant patterns in data by arranging items (sites or species) along an axis or multiple axes (McCune and Grace 2002). NMS is an ordination technique th at does not assume normality and uses a repeated search to maximize the rank correspondence between the original multidimensional space and distances in the re duced dimension of the ordination space (Peterson and McCune 2001). For the amphibian and avian NMS ordina tions 40 runs with a random starting configuration were performed with the data using the Sorensen (B ray Curtis) distance measure. A run consisted of a series of so lutions that started w ith 6 axes and stepped down in dimensionality to 1 axis. Then th e data was randomized (species presence shuffled by site) and 50 runs were complete d. Stress (an inverse measure of fit) is calculated for the runs with th e real data and the randomized data. The best ordination solution is then selected based on the real r un with the lowest stress, and dimensionality was determined when adding another dimensi on did not reduce the stress by 5 or less. Finally, the selected dimensionality had to ha ve a final stress lower than 95% of the randomized runs.

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63 Rarely encountered species can create noise in the NMS ordination results, and it is recommended that species that occur at fewe r than 5% of the sites should be deleted (McCune and Grace 2002). Thus, for the amphibian NMS only species that occurred at 2 or more sites were included and for the av ian NMS only species that occurred at 3 or more sites were included. Including amphibian s that occurred at tw o or more sites and birds that occurred at th ree or more sites in the NMS ordi nations resulted in models with the lowest stress. Correlations were conduct ed between NMS ordinati on axis scores and the measured site environmental variables. The correlation between axes scores and the LDI and FWCI were of interest and if not already the strongest correlation of the environmental variables was displayed along wi th the strongest correlation coefficients with the remaining ordination axis in joint pl ots of the data. With joint plots the angle and length of the vector indicates the st rength of the relationship between the environmental variable and ordination axes scores (McCune and Grace 2002). Stepwise multiple regression with forw ard and backward selection (alpha=0.05) was used to explore the relationship betw een the NMS axes scores with landscape variables, wetland variables, and sampling attr ibutes. The variable s listed in Appendix B were screened for covaria tion and a choice was made be tween highly co rrelated (>0.7) variables. The resulting independent vari ables were used in the stepwise multiple regressions can be found in Table 2-3. Expl oring the influence of LDI on the NMS axes scores was of interest; however, LDI was also strongly correlated (> 0.7) with the % urban and the % undeveloped. The percent ur ban land and percent undeveloped land are considered important in influencing sp ecies composition, so all (LDI, % urban, %

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64 undeveloped) were included for selection in the stepwise regression analyses. Similarly, latitude and longitude of study sites were strongly correlated (=-0.71), however both variables were included for selection in the stepwise multiple regression models. Total phosphorus was replaced separately by all of the nitrogen based water chemistry variables and separate regre ssion analyses were conducted to explore the relationship between the amphibian NMS axes scores and nutrients. However, nitrogen has been associated with causing stress in amphibian s, and all of the nitrogen based water chemistry measures were individually included in separate runs of the multiple regression analyses. Estimating habitat condition can be done at three levels of assessment: remotely, rapidly with site visits, or by more thorough m eans, this is a goal of the EPA and others wishing to evaluate landscapes (Brooks et al. 1999). Thus amphibian and avian based NMS axes scores were first only compared with 14 landscape variables that can be calculated remotely (without a site visit), then again with the landscape variables along with variables measured within the focal we tland and attributes of sampling (e.g., survey length, starting time, Julian date of first site sample). Finally, stepwise multiple regression models were create d for amphibian based NMS axes scores using landscape, focal wetland, sampling attributes, and water chemistry parameters (Table 2-3). Water chemistry parameters were not included in avian stepwise multiple regression models.

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65 Table 2-3. Independent vari ables included in stepwise regressions predicting the amphibian based NMS axes scores. Landscape Focal Wetland and Sampling Attributes Water Chemistry LDI % Broad Leaved Trees Dissolved Oxygen 1974 LDI % Evergreen Trees Total Phosphorus % Agriculture (200m) % Exotic Trees pH % Canal (200m) Basal Area Conductivity % Roadway (200m) Fire Evidence in Wetland % Silviculture (200m) Maximum Depth (cm) % Undeveloped (200m) FWCI % Urban (200m) Amphibian Species Richness % Wetland (200m) Fish Species Richness Distance to Nearest Wetland (m ) Plant Species Richness Distance to Paved Road (m) Julian Date Focal Wetland Area Starting Time (survey 1) Latitude Starting Time (survey 2) Longitude Survey Length (min) Indicator Species The amphibian NMS ordination was used to calculate species scor es with weighted averaging in PC-ORD (McCune and Mefford 1999) Site ordination scores were used to calculate the average position for each speci es along the ordination axis. The large sample size of sampled birds made it difficult to construct an observa ble graph thus avian species scores were not displayed. Indicator Species Analysis (Dufrene and Legendre 1997) was used to select amphibian and avian indicators of land use and LDI quartile in PC-ORD (McCune and Mefford 1999). Indicator Species Analysis s upplies indicator values to species based on their faithfulness of occurrence to a priori groups from a range between 100, perfect indication, and zero, no indication. Indicator values are calculated using a species

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66 relative abundance and relative frequenc y (McCune and Grace 2002). To test for statistical significance 1000 Monte Carlo randomized runs were conducted for each Indicator Species Analysis. Avian Indicator values were conducte d for land use and LDI quartile. The indicator values for amphibian s were strongest when LDI quartiles 3 and 4 were combined and when reference and silvic ulture, and agriculture and urban sites were combined. Community Characteristics Amphibian and avian species have life hi story traits that may influence their response to anthropogenic land uses or land use intensity. Within each class species share some of these life history traits (i.e., guilds) and may react in a similar manner to land use change. Identification of life history tr aits that respond to habitat characteristics of land uses or land use intensity may make it possible to apply resu lts to other regions and to the exploration of driving environmental mechanisms of species composition change with habitat alterations. Amphibian community The hydrology of wetlands is often altered with increasing land use intensity. To explore trends in amphibian community ch ange with increasing land use intensity, amphibians were grouped into either obligate or facultative users of ephemeral wetlands for reproduction (Moler and Franz 1987), a nd one amphibian species sampled was not dependent on wetlands for reproduction (Append ix E). Comparisons were made between the proportion of the amphibian communities’ ephemeral wetland dependency between land uses and LDI quartiles. Amphibian eggs attached to vegetation in wetlands that have highly variable water levels may be prone to sedimentation or experience periods of anoxia (Azous and Horner 1997). To expl ore the relationship between egg placement

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67 strategy and land use intensity amphibian spec ies were first grouped into one of four strategies (Appendix E)(Bishop 1943, Conant and Collins 1991, Bartlett and Bartlett 1999). Comparisons were made by strategy between the proportion of the amphibian community that attaches their eggs, has free fl oating eggs, is capable of attaching or has free floating eggs, and has terrestrial development, by land use category and LDI quartile. Avian community Avian species were grouped by, trophic gui ld (carnivore, herbivore, insectivore, and omnivore) (adapted from Martin et al. 1951, De graaf et al 1985, Poole and Gill 200X), foraging substrate/technique (aerial screener, bark gleaner, canopy gleaner, ground gleaner, hawker, sallier, and water ba sed)(adapted from De graaf et al. 1985), wetland dependency (Poole and Gill 200X, De Graaf et al. 1985), native/exotic, Neotropical migrant, establishes feeding te rritories (Poole and Gill 200X), nest location (canopy, cavity, ground, human structure, parasiti c, and shrub)(Ehrlich et al. 1988), and Florida population trend (significant increasi ng and decreasing species (p<0.1) between 1966-2002)(Saur et al. 2003). Species weights derived from Sibley (2000) (Appendix F).

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68 CHAPTER 3 RESULTS The organization of the results section begi ns with comparisons of attributes of the sample wetlands and circumjacent lands cape by land use. Then the site LDI index scores are compared with phys ical, biological, landscape, and water chemistry attributes of the study sites. Finally, amphibian and avian richness and composition results are presented and possible explanatory mechanis ms for the observed trends are explored. Wetland Characteristics Wetland Vegetation Generally, there were few differences in the macrophyte species richness, composition of the wetland overstory, canopy co ver, and mean basal area of the wetland vegetation between wetlands grouped by a priori land uses (Table 3-1). The dominant overstory woody vegetation of the sampled wetlands was Taxodium ascendens followed by Nyssa biflora (Table 3-2). Taxodium ascendens with N. biflora comprised more than 40% of the basal area at 108 of the 111 sites. Broad-leaved trees averaged less than 24% of the basal area at all sites combined a nd were generally more common at urban and silviculture wetlands. More of the basal ar ea at agriculture and ur ban wetlands consisted of exotic woody species and no exotic woody sp ecies were detected within the basal area measurements at the reference a nd silviculture wetlands.

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69 Table 3-1. The mean and standard deviati on of the wetland macr ophyte species richness, percent canopy cover, and tree basa l area (>10.2 dbh) by land use. Macrophyte* species richness % canopy cover** Basal area** m2/ha Reference 34.7 (10.5)a 81.1 (7.6)a 29.2 (8.8)a Silviculture 29.2 (14.8)a 81.7 (7.7)ab 30.5 (8.7)a Agriculture 37.9 (13.5)a 82.4 (8.6)ab 35.4 (13.1)a Urban 36.6 (10.4)a 85.5 (6.0)b 32.9 (13.1)a Land uses that share letter s uperscripts do not have signi ficantly different means (*= Oneway ANOVA, Tukeys multiple comparisons, **= Mann-Whitney U-test, p<0.05). Table 3-2. The mean proportion and standard deviation of the basa l area of the wetland overstory vegetation by land use. evergreen broad-leaved shrub* species exotic species Taxodium** ascendens Nyssa*** biflora R 0.12 (0.15)a 0.11 (0.16)a 0.05 (0.12)a 0.00 (0.00)a 0.83 (0.18)a 0.05 (0.10)a S 0.12 (0.05)a 0.37 (0.29)b 0.06 (0.06)a 0.00 (0.00)a 0.57 (0.28)b 0.21 (0.19)b A 0.08 (0.11)a 0.14 (0.17)a 0.02 (0.03)a 0.01 (0.03)b 0.80 (0.20)ac 0.07 (0.11)acU 0.08 (0.12)a 0.26 (0.26)b 0.02 (0.03)a 0.04 (0.13)b 0.70 (0.25)bc 0.11 (0.16)bc *= Shrub species were woody species nor mally classified as shrubs (e.g., Cyrilla racemiflora Myrica cerifera ), and were sampled in the wetland basal area measurements (stems >10.2cm dbh). **= The most dominant and ***= the second most dominant tree species. Land uses that share letter superscr ipts do not have signifi cantly different means (Mann-Whitney U-test, p<0.05). Wetland Water Depth The mean maximum water depth for all sites that had standing water at the time of the first visit in 2003 was 44.7 cm (Table 3-3) There was not a si gnificant difference between the mean maximum depth of agri culture, reference, silviculture and urban/residential land use categories (one-way ANOVA, Tukey multiple comparisons, (p=0.401)). However, there were large differenc es in the percentage of sites that did not have standing water; 35% of the urban sites were dry at the time of the first visit followed by silviculture (27%), agriculture (15%) and reference sites (9%)(Table 3-3).

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70 Table 3-3. The mean and standard deviat ion of the maximum wetland water depth at sites with standing water by land use, a nd the count of site s without standing water at the time of sampling by land use. water depth dry sites Reference 47.8 (21.4)a 3 Silviculture 53.9 (22.0)a 3 Agriculture 41.8 (25.8)a 4 Urban 41.0 (24.7)a 14 Shared letter superscripts i ndicate no significant differen ce (one-way ANOVA, p<0.05). Wetland Water Chemistry Water samples were analyzed for chemical c onstituents at 87 of the 111 sites. Sites where water samples were not ta ken were dry or had water de pths that were less than 10cm during the first site visit. There was not a significant difference in water chemistry parameters between land-uses for temperature, color, and nitrite-nitrate-nitrogen (Table 3-4). Reference sites had significantly hi gher dissolved oxygen levels, and lower pH, turbidity, conductivity, ammonia, TKN and total phosphorus than both agriculture and urban sites. Reference and silviculture sites had significantly lower nutrient levels (ammonia and total phosphorus) than both agricu lture and urban sites. Agriculture and urban sites did not have stat istically significant differences for the measured water chemistry parameters. Wetland Terrestrial Insects Mean insect species richness, mean tota l individuals captured, and mean biomass were significantly greater at agriculture ( 42, 146, 1.3g) sites than at reference (27, 59, 0.5g) and urban (33, 84, 0.5g) sites (One-way ANOVA, Tukeys multiple comparisons, p<0.05). Silviculture sites had the lowest sample size (n=9) and mean insect species richness, mean total individuals captu red, and mean insect biomass (32, 87, 0.4g respectively) were not significantly differe nt from the other land use categories.

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71 Table 3-4. Mean and standard deviation of water chemistry parameters by land use. Land uses that share lette r superscripts do not have significantly different means (Mann-Whitney U-test, p<0.05). Reference Silviculture Agriculture Urban Sites sampled (n) 30 8 23 26 Dissolved oxygen (mg/L) 3.4 (1.9)b 2.0 (1.3)ab 1.7 (1.5)a 1.5 (1.5)a Temperature (C) 26.2 (1.6)a 25.9 (1.5)a 25.9 (1.5)a 26.0 (2.0)a Color (PCU) 243 (146)a 351 (234)a 390 (425)a 228 (191)a pH 5.2 (1.2)b 4.4 (0.4)b 6.5 (0.7)a 6.2 (1.3)a Turbidity (NTU) 2.2 (3.0)b 1.9 (1.7)ab 17.4 (48.7)a 10.3 (31.5)a Conductivity (umhos/cm) 93 (171)b 46 (18)b 174 (203)a 173 (238)a Ammonia (mg N/L) 0.02 (0.02)b 0.02 (0.01)b 0.74 (1.92)a 0.11 (0.13)a NO2-NO3-N (mg N/L) 0.01 (0.02)a 0.01 (0.01)a 0.01 (0.03)a 0.01 (0.02)a TKN (mg N/L) 1.13 (0.37)b 1.28 (0.38)ab 4.13 (6.23)a 1.58 (0.54)a Total phosphorus (mg P/L) 0.04 (0.03)b 0.04 (0.02)b 1.59 (3.53)a 0.18 (0.17)a Circumjacent Land Use Characteristics Table 3-5 displays the mean distance to the nearest and second nearest wetland, distance to the closest pa ved road, and the mean proportion of wetland, road, and canal/retention pond within 200 meters of each s ite. The most significant differences in these measures were between reference and urba n sites. For example, the mean distances to the two closest wetlands at urban sites, 213 and 423 m, we re more than double that of reference sites. The mean distance to a paved road was two orders of magnitude lower at urban sites than the distance for agriculture, si lviculture, and reference sites. Reference sites have the highest mean proportion of wetland within 200 meters (p<0.01), while agriculture and urban sites ha d significantly more canals, ditches and retention ponds within 200 meters than referen ce and silviculture sites.

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72 Table 3-5. The mean and standard deviati on (in parentheses) of attributes of the landscape surrounding the sample wetlands by land use type. Shared letter superscripts indicate means that are not significantly different (p>0.05). distance to (m) Reference Silviculture Agriculture Urban nearest wetland 96a (79) 158ab (136) 136ab (103) 213b (196) 2nd nearest wetland 170a (115) 299ab (212) 252a (145) 423b (259) paved road 1600a (1410) 1948a (2201) 1248a (1172) 57 (152) proportion of 200m buffer wetland 0.25 (0.01) 0.09ab (0.07) 0.13a (0.09) 0.07b (0.07) road 0.02a (0.02) 0.02b (0.02) 0.01b (0.01) 0.08 (0.09) canal/retention pond 0.00a (0.00) 0.00a (0.00) 0.02b (0.02) 0.03b (0.06) Landscape Development Intensity The mean and median 100 meter buffer LD I score for all sites sampled in 2003 combined was 3.49 and 3.64 respectively. Four teen sites received the lowest possible LDI score (LDI=1.0), and the highest site LDI score was 7.8 (appendix d). Table 3-6 shows the 111 sites grouped by 100 meter buffe r LDI score quartiles. Quartile 1 is comprised entirely of, the a priori reference category sites an d has the least amount of variation in LDI scores. Quarti le 2 contains all of the silviculture sites along with sites from the other three major land use categories. Quartile 3 was a mix of agriculture and urban sites. Quartile 4 consisted almost entirely of urban s ites with only one high intensity agriculture site. Table 3-6. The range, mean, and standard de viation of LDI scores by LDI quartile, and the number of sites by land use type for each LDI quartile. LDI quartile range mean stdev Sites n ag(n) ref(n)siv(n) urb(n) 1 1.00-1.12 1.03 0.04 29 0 29 0 0 2 1.20-3.64 2.25 0.86 27 9 4 11 3 3 3.66-5.21 4.40 0.51 28 17 0 0 11 4 5.24-7.78 6.44 0.82 27 1 0 0 26 The mean LDI score for the 27 agricultu re sites was 3.9 with a 0.7 standard deviation. The lowest LDI score for agricu lture sites was 2.4 and the highest score was

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73 5.3. The mean LDI score for the 33 referen ce sites was 1.1 and the scores ranged from 1.0 to 1.8. The eleven silviculture sites ha d the smallest range of LDI scores, 1.5-1.8, with a mean score of 1.6. Finally, the 40 urba n sites had the largest range of LDI scores, 2.2-7.8, with a mean score of 5.7. The 100m LDI scores were significantly correlated to many features of the landscape that have the potential to influe nce amphibian and avian composition (Table 37). Examples of the strongest correlati ons include the proportion of forest area contiguous with, and including, the sample we tland within 200m (r=0.71), evidence of fire in the landscape (r=-0.75), and the am ount of undeveloped land (excludes pine plantation) within 200m of the focal wetland (r=-0.75). Significant but weaker correlations included the proportion of road way within 200m (r=0.48), distance to paved road (-0.45), the distance to the neares t and second closest wetland (r=0.39, r=048), wetland area within 200m (r=-0.40), fire ev idence in the dome (r=-0.53), dissolved oxygen (r=-0.41) and pH (r=0.45). Amphibians Amphibian Species Richness A total of 23 amphibian species were de tected during the 2003 field season giving a total of 588 species by site detections. Four species, Ambystoma talpoideum Bufo woodhousii fowleri B. marinus Notophthalmus perstriatus were encountered at only one site, and two species, Rana capito aesopus and Scaphiopus holbrookii were each detected at only two sites.

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74 Table 3-7. The Pearson correla tion coefficients between the site’s 100m buffer LDI score and selected environm ental variables. Environmental Variable Pearsons p-value N % Undeveloped Land (200m) -0.71 0.00 111 % Roadway (200m) 0.48 0.00 111 % Canal/Retention Pond (200m) 0.34 0.00 111 % Wetland (200m) -0.40 0.00 111 % Contiguous Forest (200m) -0.75 0.00 111 Nearest Wetland (m) 0.39 0.00 111 2nd Nearest Wetland (m) 0.48 0.00 111 Distance to Paved Road (m) -0.45 0.00 111 Focal Wetland (ha) 0.19 0.05 111 Fire Evidence in Dome (1/0) -0.53 0.00 111 Fire Evidence in Landscape (1/0) -0.75 0.00 111 Latitude 0.02 0.83 111 Longitude -0.04 0.65 111 Julian Date 0.07 0.48 111 Water Dissolved Oxygen -0.41 0.00 87 Water Temperature -0.11 0.30 87 Water Ammonia 0.10 0.37 87 Water NO2NO3-N 0.13 0.22 87 Water TKN 0.11 0.30 87 Water Total Phosphorus 0.09 0.40 87 Water Turbidity 0.11 0.32 87 Water Color -0.04 0.74 87 Water pH 0.45 0.00 87 Water Conductivity 0.20 0.06 87 Maximum Water Depth -0.20 0.04 111 Fish Species Richness -0.13 0.21 90 Amphibian Species Richness -0.32 0.00 111 Insect Species Richness 0.09 0.40 82 Insect Count 0.11 0.33 82 Insect Biomass 0.03 0.76 82 Plant Richness 0.09 0.37 109 Basal Area 0.15 0.12 111 % Canopy Cover 0.23 0.07 63 % Evergreen Tree Species -0.07 0.48 108 % Broad Leaved Trees 0.23 0.02 108 % Exotic Trees 0.27 0.00 108

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75 Table 3-8 shows the amphibian species ri chness by land use type, LDI quartile, and Florida region. Species detected using di p nets, cassette recordi ngs and site visits were included. Among four land use types, th e mean species richness at reference sites was significantly different than the mean sp ecies richness at urban sites. The minimum and maximum number of amphibian species sa mpled at reference or silviculture sites were three and ten, and at urban and agri cultural sites the minimum was one and the maximum was nine. Mean species richness fo r LDI quartile 1 was significantly greater than the other three quartiles. However, LDI quartile 2 had the highest total species richness. There was not a significant differe nce in the mean species richness by region. The panhandle sites had the highest total spec ies richness of the four sample regions. Table 3-8. Amphibian species richness by land use, 100 meter LDI quartile, and Florida region. Species richness was determined by using a composite richness of all sampling methods (e.g., duration of the en tire site visit, dip net sweeps, and cassette recordings). sites (n) mean (s.d.) total range Reference 33 6.5 (2.0)a 20 (3-10) Silviculture 11 5.3 (2.4)ab 16 (3-10) Agriculture 27 5.3 (2.2)ab 18 (1-9) Urban 40 4.2 (2.2)b 19 (1-9) LDI Quartile 1 29 6.8 (2.0) 19 (3-10) LDI Quartile 2 27 5.1 (2.2)a 20 (2-10) LDI Quartile 3 28 4.7 (2.1)a 17 (1-8) LDI Quartile 4 27 4.5 (2.4)a 18 (1-9) Panhandle 27 6.0 (2.4)a 19 (3-10) North 30 5.3 (2.3)a 18 (1-9) Central 29 5.3 (2.4)a 15 (1-10) South 25 4.5 (2.1)a 15 (1-8) all sites 111 5.3 (2.3) 23 (1-10) Note: Land uses with similar letter superscr ipts do not have signifi cantly different means (ANOVA, Tukey multiple comparisons, p<0.05).

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76 Figures 3-1 and 3-2 show the mean amphibian richness by land use and LDI quartile as detected by only using the two 15 minute surveys for each site. The pattern of only the reference sites being significantly differe nt than the urban site s is repeated as in Table 3-8. The amphibian species richness decreases with increasing LDI quartile; however, the relationship is only signif icant for LDI quartiles 1 and 4. Amphibian Species Frequency of Occurrence Table 3-9 shows the frequency of occurre nce for all species by land use and all sites combined corrected fo r species-specific ranges (i.e., known county occurrences, (ARMI 2004)). Rana sphenocephala and Hyla cinerea were encountered most frequently (64% of the sites) followed closely by H. squirella (60% of the sites). Silviculture sites were only sampled in the panhandle and north Florida regions thus Bufo marinus and Osteopilus septentrionalis south Florida species, were not included in the frequency of occurrence for this land use. The species’ w ith the highest frequency of occurrence at silviculture sites were H. femoralis R. sphenocephala and Acris gryllus (91%, 82%, and 73% of the sites respectively). Hyla femoralis and A. gryllus also had the highest frequency of occurrence at refere nce sites (85% of the sites). Hyla cinerea H. squirella and R. sphenocephala had the highest frequency of o ccurrence at both agriculture and urban sites.

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77 ab b 0 1 2 3 4 5 6 7 rsauamphibian species richness inside the wetland inside the wetland and the landscape a a a a a ab Figure 3-1. Mean and standard deviation of amphibian species richness within the wetland and within the wetland and th e 200m buffer combined by land use type. Species richness was determined during two 15-minute surveys for each site. Means that share letters were no t significantly different between the land use types (One-way ANOVA, Tuke y multiple comparisons, p<0.05). b a 0 1 2 3 4 5 6 Q1Q2Q3Q4amphibian species richnes s inside the wetland inside the wetland and the landscape aaa a abab Figure 3-2. Mean and standard deviation of amphibian species richness within the wetland, and within the wetland and the 200m buffer combined by LDI quartile. Species richness was determ ined during two 15-minute surveys for each site. Means that share letters were not significantly different between the LDI quartiles (p<0.05, One-way ANOVA, Tukey multiple comparisons).

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78 Amphibian Breeding Effort The mean number of tadpoles sampled at a ll sites with standing water during 20 dip net sweeps was 9.6. However, the standard deviation was 15.1 even after excluding an outlier that had approximately 1266 Scaphiopus holbrookii tadpoles collected during 20 sweeps (also excluded in Figures 3-3 and 3-4). The mean number of tadpoles and the mean number of tadpole species was significantly different fo r only reference sites versus urban sites (Figure 3-3). The mean number of species calling and their mean maximum calling intensity were not significantly different between land uses. At all sites combined the mean number of tadpole species, mean number of species calling, and their mean maximum calling intensity were 1.7, 3.9, and 1.8 respectively. LDI quartile one had the highest mean numb er of tadpoles, tadpole species, species calling and mean maximum calling intensity (Figure 3-4). However, only the mean number of tadpoles and the mean number of tadpole species found at quartile 1 versus quartile 4 sites were significantl y different (p<0.05). Amphibian Species Composition Table 3-10 shows the results of the MR PP test for significant differences in amphibian species composition by dominant la nd use and LDI quartile. The amphibian species composition was not signi ficantly different between reference and silviculture sites, and between agriculture and urban sites. All other between land use type comparisons were significantly different at th e p=0.01 level. The eff ect size or chancecorrected within-group agreemen t (A) statistic show s the biggest difference in amphibian composition between land use types was the reference and urban sites (A=0.122) followed by reference versus agriculture si tes (A=0.101). Results of the MRPP by LDI quartile show that only the third and fourth qua rtiles did not have a significant difference

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79 Table 3-9. Amphibian frequency of occurren ce by land use type and all sites combined.a Reference Silviculture Agriculture Urban All Sites Acris gryllus 0.85 0.73 0.41 0.15 0.48 Ambystoma talpoideum 0.00 0.00 0.07 0.00 0.01 Bufo marinus 0.00 NA 0.00 0.11 0.06 Bufo quercicus 0.52 0.36 0.19 0.05 0.25 Bufo terrestris 0.06 0.09 0.22 0.28 0.18 Bufo woodhousii fowleri 0.14 0.00 0.00 0.00 0.04 Eleutherodactylus planirostris 0.06 0.09 0.30 0.35 0.23 Gastrophryne carolinensis 0.27 0.18 0.52 0.35 0.35 Hyla chrysoscelis 0.07 0.09 0.08 0.16 0.11 Hyla cinerea 0.58 0.27 0.81 0.68 0.64 Hyla femoralis 0.85 0.91 0.11 0.10 0.41 Hyla gratiosa 0.39 0.18 0.11 0.00 0.16 Hyla squirella 0.52 0.36 0.74 0.65 0.60 Notophthalmus perstriatus 0.08 0.00 0.00 0.00 0.02 Notophthalmus viridescens 0.09 0.00 0.07 0.05 0.06 Osteopilus septentrionalis 0.13 NA 0.33 0.24 0.23 Pseudacris ocularis 0.69 0.50 0.16 0.18 0.35 Rana capito aesopus 0.03 0.00 0.00 0.03 0.02 Rana catesbeiana 0.22 0.09 0.41 0.26 0.27 Rana clamitans 0.39 0.36 0.47 0.28 0.36 Rana grylio 0.52 0.18 0.30 0.10 0.29 Rana sphenocephala 0.58 0.82 0.67 0.63 0.64 Scaphiopus holbrookii 0.00 0.09 0.00 0.03 0.02 a=species frequency of occurrence was calcula ted using only sites th at were located in counties with species-speci fic records of historic al occurrence (ARMI 2004).

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80 ab ab b b 0 2 4 6 8 10 12 14 tadpole counttadpole speciesspecies callingmax intensity r s a u a ab a ab aaaa aaa a Figure 3-3. The mean tadpole count, mean nu mber of tadpole species, mean number of anuran species calling and mean maximum intensity of anuran species calling by land use. Land uses that share lett ers do not have significantly different means (Mann-Whitney U-test, p<0.05). (Only sites with standing water and where 20 dip-net sweeps were conducted were included)(Species calling and mean maximum intensity based on s ites with audio logger data only) a a b b 0 2 4 6 8 10 12 14 16 tadpole counttadpole speciesspecies callingmax intensity Q1 Q2 Q3 Q4 ab ab abab a aaa a a aa Figure 3-4. The mean tadpole count, mean nu mber of tadpole species, mean number of anuran species calling on cassette tape s, and mean maximum intensity of anuran species calling by LDI quartile. Qu artiles that share letters do not have significantly different means (Mann-Whitn ey test p<0.05). (Only sites with standing water and where 20 dipnet sweeps were conducted were included)(Species calling and mean ma ximum intensity based on sites with audio logger data only)

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81 in amphibian species composition, and the ne gative A value indicates there was less agreement within these groups than exp ected by chance. The largest significant difference in species composition occurred between LDI quartile 1 and 4 followed by LDI quartile 1 and 3, (A=0.150 and 0.135). Figure 3-5 shows axis one and three of a nonmetric multidimensional scaling (NMS) ordination using presence/absence data for all amphibian species found at two or more sites in 2003. The final stress on the ordination was 17.5 and was stable after 183 iterations. The sites in Figure 3-5 are coded by their a priori land use classification. Approximately 78% of the variance was re presented with three NMS axes, and the greatest proportion of variation was explai ned by axis three ( 35%)(Table 3-11). The angle and length of the arrows in Figur e 3-5 represent the relationship with the environmental variables and the ordination sc ores. The environmental variables LDI and FWCI were significantly correla ted to the y-axis (axis 3) and the maximum site depth was significantly correlated with the x-axis (axis 1)(Table 3-12). The correlation between axis three and the independent vari ables FWCI (Reiss 2004) and the percentage of sensitive plants were the strongest corre lations, 0.77 and 0.78 respectively. Axis three ordination scores had the strongest and the most si gnificant (p<0.01) Pearson’s correlations with the site-specific independe nt variables (Appendix b). Axis one was significantly correlated (Pears on, p<0.01) with maximum water depth, percent canal/ditch within 200 meters, Julian date, survey length an d latitude. The correlation between axis one and water depth had the only coefficient gr eater than 0.5. Axis 2 (not shown) was significantly correlated (Pear son, p<0.01) with amphibian richness and maximum water depth, -0.37 and –0.25 respectively.

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82 Table 3-10. MRPP test for significant diffe rences in amphibian species composition between land use types and LDI quartiles. Variable A* T-statistic p-value Land use All land uses 0.109 -17.66 p<0.000 A vs R 0.101 -14.99 p<0.000 A vs S 0.094 -8.63 p<0.000 A vs U 0.001 -0.12 p=0.393 R vs S 0.008 -0.81 p=0.189 R vs U 0.122 -22.85 p<0.000 S vs U 0.076 -9.76 p<0.000 LDI Quartiles All quartiles 0.096 -15.74 p<0.000 1 vs 2 0.024 -3.34 p<0.006 1 vs 3 0.135 -18.72 p<0.000 1 vs 4 0.150 -21.02 p<0.000 2 vs 3 0.038 -5.39 p<0.000 2 vs 4 0.047 -6.71 p<0.000 3 vs 4 -0.004 -0.53 p=0.658 Land use abbreviations: A=agri culture, R=reference, S=silv iculture, and U=urban. = Agreement statistic, chance corrected within group agreement. The reference site with the lowest site sc ore on the y-axis (NMS axis 3) had a rural highway cutting through its so uthern boundary and was located in a relatively small (336 ha), isolated county park. Four of the five agricultural sites with positive y-axis scores (NMS axis 3) were wetlands located on pub lic conservation lands that were lightly grazed. The urban sites with positive y-axis scores were wetlands adjacent to either undeveloped land or very recent urban developments.

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83 -1.5 0 1.5 -1.501.5 agriculture reference silviculture urban LDI Maximum Depth FWCI A xis 1 A xis 3 Figure 3-5. Unrotated joint pl ot of the amphibian based NM S ordination site scores for axis one and axis three by land use with the strength and re lationship of three environmental variables (FWCI, LDI, and maximum depth). Axis 2 is not shown. Table 3-11. Coefficients of determination (R2) between amphibian based NMS ordination distances and distances in th e original space. Proportion of the variance represented by each axis is listed incrementally and cumulatively. Axis Increment Cumulative 1 0.247 0.247 2 0.187 0.434 3 0.348 0.782

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84 Table 3-12. Pearson’s correlati on coefficients (r) and Kendall’s Tau comparisons with NMS ordination axes based on amphibian species composition and LDI, FWCI, and maximum water depth. LDI FWCI Maximum Depth axis 1 r 0.028 0.005 0.521 tau 0.039 -0.013 0.379 axis 2 r 0.026 0.128 -0.246 tau 0.030 0.072 -0.178 axis 3 r -0.683 0.776 0.280 tau -0.478 0.570 0.223 Table 3-13. Independent vari ables included in stepwise regressions predicting the amphibian based NMS axes scores. Landscape Focal Wetland and Sampling Attributes Water Chemistry LDI % Broad Leaved Trees Dissolved Oxygen 1974 LDI % Evergreen Trees Total Phosphorus % Agriculture (200m) % Exotic Trees pH % Canal (200m) Basal Area Conductivity % Roadway (200m) Fire Evidence in Wetland % Silviculture (200m) Maximum Depth (cm) % Undeveloped (200m) FWCI % Urban (200m) Amphibian Species Richness % Wetland (200m) Fish Species Richness Distance to Nearest Wetland (m ) Plant Species Richness Distance to Paved Road (m) Julian Date Focal Wetland Area Starting Time (survey 1) Latitude Starting Time (survey 2) Longitude Survey Length (min) Table 3-14 shows the results of a stepwise regression using the landscape variables (Table 3-13) to represent the NMS ordinati on scores based on amphibian composition at the 111 sites. Axis 3 had the highest adjusted R2 (59.5) among the three axes. LDI, percent urban, percent agriculture, distance to the neares t wetland, latitude and longitude were all selected as Landscape predictors of axis 3. Axis 1 had a relatively low adjusted R2 (16.3) and two variables, th e area of the focal wetland a nd latitude, were selected. There were no landscape variables that were co rrelated with axis 2 at the 0.05 level.

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85 Table 3-14. The variables and regression coefficients select ed (p<0.05) from a stepwise regression using the Landscape variables listed in Table 313 to represent the amphibian based NMS ordination axes. Selected Variables axis 1 axis 2 axis 3 LDI -0.10 Focal Wetland Area 0.30 % Urban (200m) -0.79 % Agriculture (200m) -0.70 Distance to Nearest Wetland (m) 7.40E-04 Longitude 0.09 Latitude 0.14 0.15 Adjusted R-square 16.32 59.54 Table 3-15 shows the results of a step wise regression using Landscape, Focal Wetland and Sampling variables (Table 3-13) to represent the NMS ordination scores based on amphibian composition at the 111 site s. Axis 3 had the highest adjusted R2 (70.9) followed by axis 1 (37.1) than axis 2 (27.3). When comparing Table 3-14 with Table 3-15 the adjusted R2 of axis three improved by more then 10% with the addition of focal wetland specific measures. None of the Sampling Effort variable s were selected as predictors of axis 3. The selected variab les for axis 3 included; LDI, FWCI, proportion agriculture (200m), distance to the nearest wetland, latitude and longitude. Maximum depth, Julian date and latitude we re selected as predictors of axis 1. Finally, predictors of axis two were primarily wetland or sampling effort variables. Table 3-16 shows the results of a stepwise regression using a total of 32 landscape, focal wetland, sampling, and water chemistry variables (Table 3-13) to represent the NMS ordination scores based on amphibian com position at 87 sites. Only the adjusted R2 for axis 3 (73.6) improved from Tables 314 and 3-15 with the a ddition of the water chemistry variables. Water conductivity and pH along with LDI and FWCI were selected

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86 as predictors of axis 3. Axis 1 and 2 di d not have any water chemistry parameters selected as predictors. Table 3-15. The variables and regression coefficients select ed (p<0.05) from a stepwise regression using the Landscape, Focal Wetland, and Sampling variables listed in Table 3-13 to represent the amphi bian based NMS ordination axes. Selected Variables axis 1 axis 2 axis 3 LDI -0.09 FWCI 5.7 E-03 0.03 % Agriculture (200m) 0.34 Distance to Nearest Wetland (m) 7.0E-04 Maximum Depth (cm) 8.1E-03 -4.1 E-03 % Exotic Trees (focal wetland) -1.42 Amphibian Species Richness -0.09 Julian Date 6.6E-03 Longitude 0.06 Latitude 0.08 0.09 0.12 Adjusted R-square 37.08 27.34 70.87 Table 3-16. The variables and regression coefficients select ed (p<0.05) from a stepwise regression using the Landscape, Focal Wetland, Sampling, Water Chemistry variables listed in Table 3-13 to represent the amphibian based NMS ordination axes. Selected Variables axis 1 axis 2 axis 3 Amphibian Species Richness -0.09 -0.08 Survey Length (min) 3.1E-03 Maximum Depth (cm) 5.7E-03 Latitude 0.09 % Agriculture (200m) -0.29 LDI -0.09 pH -0.14 Conductivity -5.0E-04 FWCI 0.01 Adjusted R-square 29.96 18.32 73.59

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87 Amphibian Indicator Species Figure 3-6 shows the same ordination as Fi gure 3-5, but with the species ordination scores plotted on axes one and three. The sp ecies most associated with sites with high FWCI scores and low LDI scores were Bufo quercicus Hyla femoralis H. gratiosa Pseudacris ocularis and Acris gryllus The two exotic species Eleutherodactylus planirostris and Osteopilus septentrionalis along with native B. terrestris were most often associated with sites that had low FWCI sc ores and had high LDI scores. Ranid species along with Hyla chrysoscelis and Notophthalmus viridescens presence was associated with deeper sites. Eleutherodactylus planirostris Osteopilus septentrionalis Hyla squirella and B. quercicus were associated with shallow or dry sites. The weighted score of Gastrophryne carolinensis was the closest to the origin of axis 1 and 3. This suggests G. carolinensis was not strongly associated with land use intensity or water depths. Results of indicator species analyses reflect the species pattern observed in Figure 3-6. The majority of the significant indicator s of LDI quartiles (Table 3-17) and land use (Tables 3-18 and 3-19) tend to be indicators of reference sites or lo w intensity sites (e.g., Acris gryllus Bufo quercicus, H. femoralis, H. gratiosa, Pseudacris ocularis and Rana grylio ). When LDI quartiles thr ee and four were lumped or agriculture and urban sites were lumped (Tables 3-17 and 3-19) Eleutherodactylus planirostris B. terrestris Gastrophryne carolinensis and H. squirella had their highest indicator value at these higher land use intensity sites. The exotic Osteopilus septentrionalis had increasing indicator values with incr easing land use intensity, how ever, the pattern was not significant (p=0.15). An additional 10 amphibi an species detected in 2003 were not significant indicators of land use or development intensity.

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88 OSSE ELPL HYSQ HYCH RACA NOVI RASP RACL RAGR RAAE ACGR PSOC BUQU HYFE HYGR GACA HYCISCHO BUTE -0.8 0 0.8 -0.800.8 FWCI LDI Maximum DepthAxis 3 Axis 1 Figure 3-6. Unrotated joint plot of the amphibian based NMS ordination species scores for axes one and three with the st rength and relationship of three environmental variables (FWCI, LD I, and maximum depth). (ACGR= Acris gryllus BUQU= Bufo quercicus BUTE= Bufo terrestris ELPL= Eleutherodactylus planirostris GACA= Gastrophryne carolinensis HYCH= Hyla chrysoscelis HYCI= Hyla cinerea HYFE= Hyla femoralis HYGR= Hyla gratiosa HYSQ= Hyla squirella NOVI= Notophthalmus viridescens OSSE= Osteopilus septentrionalis PSOC= Pseudacris ocularis RAAE= Rana capito aesopus RACA= Rana catesbeiana RACL= Rana clamitans RAGR= Rana grylio RASP= Rana sphenocephala SCHO= Scaphiopus holbrookii ). The species not selected as indi cators were either ubiquitous (e.g., Rana sphenocephala ), evenly distributed throughout land us e intensities or had low frequency of occurrence (e.g. Rana capito and Scaphiopus holbrookii ).

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89 Table 3-17. Significant (p<0.10) amphibian indicator species values by LDI quartile with quartiles three and four combined. Q1 Q2 Q3 and Q4 p value Acris gryllus 44 23 2 0.0002 Bufo quercicus 37 7 1 0.0002 Bufo terrestris 0 3 19 0.0204 Eleutherodactylus planirostris 1 6 18 0.0637 Gastrophryne carolinensis 10 4 22 0.0892 Hyla femoralis 53 18 1 0.0001 Hyla gratiosa 34 2 0 0.0001 Hyla squirella 14 14 31 0.0502 Pseudacris ocularis 44 9 2 0.0001 Rana grylio 27 11 2 0.0046 Two species, Gastrophryne carolinensis and Hyla cinerea had the highest significant indicator values for the agricultu ral land use category (Table 3-18). Only Eleutherodactylus planirostris was selected as an indicator of urban sites (Table 3-18). Hyla femoralis had a slightly higher i ndicator value for silviculture sites than reference sites, 42 and 37 respectively. The indicator values for Acris gryllus Bufo quercicus Hyla gratiosa and Pseudacris ocularis decrease with the increasing land use intensity represented by the four land us e categories (i.e., Reference < Silviculture < Agriculture < Urban). Only one species selected as a significant indicator of reference sites, Rana grylio did not have indicator values that corre spond with the land use intensity gradient where indicator values for agriculture were slightly higher than silviculture. Grouping the two higher intensity land use categorie s, agriculture and urban sites, and the lower intensity sites, reference and silviculture, resulted in the highest indicator values (Table 3-19). Four species had signifi cantly higher indicator values for agriculture and urban ( B. terrestris, Eleutherodactylu s planirostris, Hyla cinerea and H. squirella ).

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90 Six species had significantly higher indicator values for re ference sites combined with silviculture sites. Table 3-18. Significant (p< 0.10) amphibian indicator sp ecies values by land use. r s a u p-value Acris gryllus 34 25 8 1 0.0028 Bufo quercicus 24 12 3 0 0.0342 Eleutherodactylus planirostris 0 1 11 15 0.0912 Gastrophryne carolinensis 6 2 20 9 0.0694 Hyla cinerea 14 3 28 19 0.0479 Hyla femoralis 37 42 1 1 0.0007 Hyla gratiosa 23 5 2 0 0.0161 Pseudacris ocularis 31 14 2 2 0.0091 Rana grylio 24 3 8 1 0.0136 Table 3-19. Significant (p< 0.10) amphibian indicator sp ecies values by land use (agriculture and urban versus reference and silviculture). ref and siv ag and urb p-value Acris gryllus 62 6 0.0001 Bufo quercicus 39 2 0.0001 Bufo terrestris 1 20 0.0208 Eleutherodactylus planirostris 1 27 0.0020 Hyla cinerea 20 43 0.0280 Hyla femoralis 77 1 0.0001 Hyla gratiosa 30 1 0.0001 Hyla squirella 20 41 0.0515 Pseudacris ocularis 48 3 0.0001 Rana grylio 31 5 0.0033 Amphibian Community Characteri stics and Fish Composition Isolated depressional wetla nds are important breeding ar eas for a subset of the Florida amphibian community (Moler an Franz 1987). Ten of 14 species found in Florida that are dependent on ephemeral we tlands and 11 of 21 facultative users were encountered. Reference and LDI quartile 1 sites had a signif icantly higher mean

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91 proportion of amphibians depende nt on ephemeral wetlands than urban and LDI quartile 4 sites respectively (Table 3-20). Only one species, Eleutherodactylus planirostris encountered is independent of we tlands for breeding in Florida. E. planirostris comprised a significantly greater proportion of the amphibian species richness at urban and LDI quartile 4 sites than reference a nd LDI quartile 1 sites respectively. Although the proportion of amphibian species that facultatively utilize ephemeral wetlands increased with increasing land use intensity or high intensity land uses the trend was not significant at the p<0.05 level. Table 3-20. The mean proportion of amphibi an species independent of wetlands for breeding, dependent on ephemeral wetlands, and will facultatively use ephemeral wetlands by land use and LDI quartile. Independent Dependent Facultative r 0.01a 0.50a 0.48a s 0.01ab 0.49ab 0.49a a 0.05ab 0.40ab 0.55a u 0.10b 0.30b 0.60a Q1 0.01a 0.52a 0.47a Q2 0.04ab 0.40ab 0.55a Q3 0.06ab 0.41ab 0.53a Q4 0.11b 0.26b 0.63a Shared letters indicate no significant diffe rence. (One Way ANOVA, Tukey multiple comparisons, p<0.05, Independent arcsin squareroot transformed) Isolated ephemeral wetlands are impo rtant to a subset of the amphibian community because they are often free of predat ory fish. However, fish were sampled at 53% of the sites that had sta nding water (Figure 3-7). Out of the 47 wetlands with fish 44 had at least one amphibian species dependent on isolated ephemeral wetlands. Fifteen fish species were sampled and the most common were Gambusia affinis Elassoma sp., Heterandria formosa and Lepomis sp. (Table 3-21). Elassoma sp. was only found at the low intensity land uses, 55% of the wet refere nce sites and 63% of the wet silviculture

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92 sites. Gambusia affinis was found at approximately half of the wet sites regardless of land use type. Perhaps the most pred atory group relative to amphibians, Lepomis sp., was only recorded at two reference a nd two agricultural sites. Amphibian eggs that are attached to substrat es may have lower hatch success in wetlands that experience widely fluctuating water levels (Figures 3-8 and 3-9) Species that have free floating eggs make up a significantly higher proportion of the amphibian community at sites that have higher intens ity land uses. Meanwhile, species that attach their eggs to substrates make up a higher proportion of th e amphibian community at sites that have lower intensity surrounding land uses. However, there was considerab le variation in the egg laying strategy of the amphibian community at the sampled sites within each land use category and LDI quartile, and the relationshi p between egg laying strategy and land use intensity was significantly diffe rent primarily for sites at th e opposite ends of the land use intensity gradient. 0.00 0.10 0.20 0.30 0.40 0.50 0.60 0.70 refsivagurb % of sites with fish % of wet sites with fish Figure 3-7. The proportion of all sites and only sites that had standing water in 2003 where fish were encountered by land use.

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93 Table 3-21. The number of sites where fish sp ecies occurred by land use. There were 24, 31, 8, and 26 agriculture, refere nce, silviculture, and ur ban sites, respectively, with standing water at the time of sampling. Ref Siv Ag Urb Total Ameiurus sp. 1 1 Aphredoderus sayanus 1 1 Astronotus ocellatus 1 1 Elassoma sp. 10 4 14 Etheostoma sp. 1 1 Fundulus chrysotus 1 1 Gambusia affinis 13 3 12 14 42 Heterandria formosa 7 2 1 10 Hoplosternum littorale 2 2 Jordanella floridana 1 1 Lepomis sp. 2 2 4 Leptolucania ommata 1 1 Lucanai goodie 1 1 ? Notopterus chitala 1 1 Poecilia latipinna 1 1 2 *=exotic b a bc a ab a b a a b a b a b ac b 0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 % attached% both% float% terrestrial r s a u Figure 3-8. The proportion of the amphibian comm unity that attaches eggs to substrates, has free floating eggs, attaches or has floati ng eggs, and has terrestrial egg development by land use. Land uses that share letters we re not significantly different (Mann-Whitney U-test, p<0.05)

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94 a a a a ab ab ab a b b b b b b ab b 0 0.1 0.2 0.3 0.4 0.5 0.6 % attached% both% float% terrestrial Q1 Q2 Q3 Q4 Figure 3-9. The proportion of the amphibian comm unity that attaches eggs to substrates, has free floating eggs, attaches or has floati ng eggs, and has terrestrial egg development by LDI quartile. Land uses that share lett ers were not significantly different (MannWhitney U-test, p<0.05) Birds Bird Species Richness One hundred and four bird species were detected at 111 site s during the 2003 field season for a total of 1,557 species by site de tections. Eighteen of the species were detected at only one site, and an additional nine species were detected at only two sites. The survey period, May 30th to August 27th, primarily encompassed the period summer residents were present throughout Florida. However, during the surveys that occurred in the later part of the year so me of the species detected we re most likely migrating (e.g., belted kingfisher, Louisiana waterthrush, Kentucky warbler, and yellow Warbler). Northern cardinal, Carolina wren, red-bel lied woodpecker, and mourning dove were the most frequently recorded avian species 95, 94, 85, and 66 sites respectively.

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95 Table 3-22 shows richness measures for bi rd species by land use, LDI quartile, and Florida region. The mean number of species detected in the wetland and surrounding area was significantly greater at agricultural s ites than both silviculture and urban sites. There was no significant difference in bird sp ecies richness between the LDI quartiles or Florida regions. Total species richness was lo west at silviculture sites (45), which may partially be an artifact of th e relatively smaller number of sampled sites of this land use type (n=11). However, quartile 2, which encomp asses all of the silviculture sites, also had the lowest total species richness (68) re lative to the three other quartiles. The Panhandle had the highest total species richne ss (78) among Florida regions and the south had the lowest (69). The most bird species detected during a combination of the two surveys was 32 at a reference site in the Pa nhandle. The lowest number of bird species detected was 4, at an urban site in the Panhandle. The highest land use intensity quartile (Q4) had almost twice as many bird species (5.0 vs 2.7) detected within the wetland duri ng both 15 min surveys than did the lowest land use intensity quartil e (Q1) (Table 3-23). The ratio of bird species detected within LDI quartile 1 wetlands to those detected in the landscape was 1 to 3.4 while the ratio in LDI quartile 4 wetlands to the landscape dete ctions was 1 to 1.5. The mean number of bird species detected in the wetland increased with increasing land use intensity quartile. Mean bird species richness, as determ ined by two 15 min surveys per site, was significantly greater in the surrounding landscap e of agriculture sites than the landscape surrounding urban sites. There was not a sign ificant difference in mean bird species richness between all other land uses, LDI qua rtiles, and Florida re gions. Finally, there

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96 was not a significant difference in bird speci es richness among the Florida regions during the 15 min surveys. Table 3-22. Avian species richness by land use, LDI quartile, and Florida region. Species richness was determined by us ing the composite richness of all sampling methods (e.g., the duration of bot h site visits, cassette recordings, and species detected inside and with in 200 meters of the focal wetland). sites (n) mean (s.d.) total range Reference 33 14.5 (5.1)ab 74 (7-32) Silviculture 11 12.2 (3.9)b 45 (5-17) Agriculture 27 16.7 (5.3)a 76 (9-28) Urban 40 12.3 (3.9)b 76 (4-20) LDI Quartile 1 29 14.6 (5.2)a 72 (7-32) LDI Quartile 2 27 13.4 (4.4)a 68 (5-22) LDI Quartile 3 28 15.5 (5.0)a 75 (6-26) LDI Quartile 4 27 12.5 (4.8)a 71 (4-28) Central 27 13.3 (4.6)a 71 (6-22) North 30 14.4 (4.7)a 70 (5-26) Panhandle 29 15.0 (6.1)a 78 (4-32) South 25 13.3 (4.3)a 69 (6-21) all sites 111 14.0 (4.9) 104 (4-32) Shared letter superscripts denote means that were not significantly different (One way ANOVA, Tukey multiple comparisons (p<0.05)).

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97 Table 3-23. Mean number and standard devia tion of bird species detected within the wetlands, landscape, total, and landscap e to wetland ratio during two 15 min surveys by land use, LDI quartile, Fl orida region and all sites combined. Wetland Landscape Total Landscape/ Wetland Reference 2.8 (2.7)a 9.0 (2.9)ab 10.1 (3.5)ab 3.2 Silviculture 2.5 (2.0)a 7.9 (3.7)ab 8.7 (3.8)ab 3.2 Agriculture 4.6 (3.4)a 9.7 (3.9)a 12.0 (4.5)a 2.1 Urban 4.4 (2.4)a 7.3 (3.3)b 9.0 (3.0)b 1.7 Q1 2.7 (2.5)a 9.3 (2.9)a 10.3 (3.3)a 3.4 Q2 3.6 (3.2)ab 7.8 (3.2)a 9.4 (4.1)a 2.2 Q3 4.0 (2.3)ab 9.2 (4.2)a 10.9 (4.1)a 2.3 Q4 5.0 (2.9)b 7.3 (3.3)a 9.4 (3.5)a 1.5 Panhandle 3.4 (2.8)a 8.9 (3.7)a 10.6 (4.0)a 2.6 North 4.5 (2.7)a 8.9 (3.9)a 10.3 (4.2)a 2.0 Central 3.3 (2.5)a 7.6 (3.6)a 9.4 (3.8)a 2.3 South 3.9 (3.2)a 8.4 (2.4)a 9.8 (3.2)a 2.2 All Sites 3.8 (2.8) 8.4 (3.5) 10.1 (3.8) 2.2 Shared letter superscripts denote means that were not significantly different (p<0.05)(One way ANOVA, Tukey multiple comparisons). Bird Species Composition Bird species composition was significantly different between land use types and LDI quartiles for all comparisons (p<0.013) exce pt for reference versus silviculture sites (p>0.05)(Table 3-24). The biggest differen ce in avian species composition among land use types was between reference and urban sites, A equal to 0.09, and LDI quartile 1 and LDI quartile 4 showed the bigge st difference among quartiles, A equal to 0.11. There was a gradient in avian composition that follo ws the LDI quartile gradient. That is, the quartiles adjacent to each other have avian compositions that are more similar (i.e., smaller A values) than quartiles farther away.

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98 Table 3-24. MRPP test for signi ficant differences in bird species composition between sites by land use and LDI score quartiles Variable A T-statisticp-value Land use All 0.085 -23.97 p<0.000 A vs R 0.054 -15.12 p<0.000 A vs S 0.034 -5.90 p<0.000 A vs U 0.047 -14.37 p<0.000 R vs S 0.001 -0.27 p=0.355 R vs U 0.090 -26.87 p<0.000 S vs U 0.059 -12.50 p<0.000 LDI Quartile All 0.071 -20.00 p<0.000 1 vs 2 0.014 -3.72 p<0.002 1 vs 3 0.057 -15.11 p<0.000 1 vs 4 0.113 -24.42 p<0.000 2 vs 3 0.010 -2.71 p<0.012 2 vs 4 0.067 -15.07 p<0.000 3 vs 4 0.025 -6.45 p<0.000 Figures 3-10 and 3-11 show the results of an NMS ordination using bird species composition at each site coded by land use. A three dimensional NMS model (i.e., three axes) was selected after the maximum allott ed 400 iterations, and the final stress was 21.0. Approximately 66% of the variance was explained with three dimensions. Axis one explained 27% of the variation, while ax es two and three expl ained 22% and 17% of the variation respectively (Table 3-25). In both figures 3-10 and 3-11 a general pattern of higher LDI sites gradating into lower LDI sites appears along axis 1. Axis one had the strongest Pearson’s correlation with LDI followed by FWCI (r = -0.74 and 0.58 respectively)(Table 3-26). Axis three was also significantly correlated (p<0.01) to LDI and FWCI, 0.48 and –0.53 respectiv ely (Table 3-26 and Appendix C). The strongest significant Pearson’s correlati on (p<0.01) with axis 2 (r = -0.51) was the proportion of

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99 land use within in 200 meters of the focal we tland that was categorized as agriculture (Table 3-26 and Appendix C). 1. 5 0 1. 5 1. 5 0 1. 5 agriculture reference silviculture urban FWCI LDI A xis 1 A xis 3 Figure 3-10. Unrotated joint plot of the avia n based NMS ordination site scores for axes one and three with the strength and re lationship of the landscape variables LDI and FWCI.

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100 2 0 2 2 0 2 Agriculture Reference Silviculture Urban A xis 2 A xis 1 LDI % agriculture within 200m FWCI Figure 3-11. Unrotated joint plot of the avia n based NMS ordination site scores for axes one and two with the strength and rela tionship of three landscape variables LDI, FWCI, and proportion of the landscape within 200 meters of the wetland that was used for agriculture. Table 3-25. Coefficients of determination (r2) between avian based NMS ordination distances and distances in the original space. Proportion of the variance represented is listed incrementally and cumulatively. Axis Increment Cumulative 1 0.266 0.266 2 0.215 0.481 3 0.174 0.655

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101 Table 3-26. Pearson’s correlation coefficients (r) and Kendall’s Tau comparisons with NMS ordination axes based on av ian species composition and the environmental variables with the strongest correlation. axis 1 axis 2 axis 3 LDI r -0.74 -0.04 0.48 tau -0.54 -0.02 0.32 FWCI r 0.58 0.33 -0.53 tau 0.39 0.22 -0.37 % ag (200m) r -0.09 -0.51 0.21 tau -0.08 -0.37 0.14 Table 3-27. Independent variab les included in stepwise regr essions to predict the avian based NMS axes scores. Landscape Focal Wetland and Sampling Attributes LDI % Broad Leaved Trees 1974 LDI % Evergreen Trees % Agriculture (200m) % Exotic Trees % Canal (200m) Basal Area % Roadway (200m) Fire Evidence in Wetland % Silviculture (200m) Maximum Depth (cm) % Undeveloped (200m) FWCI % Urban (200m) Amphibian Species Richness % Wetland (200m) Fish Species Richness Distance to Nearest Wetland (m) Plant Species Richness Distance to Paved Road (m) Julian Date Focal Wetland Area Starting Time (survey 1) Latitude Starting Time (survey 2) Longitude Survey Length (min) Table 3-28 shows the results of a stepwi se regression using landscape variables (Table 3-27) to represent the NMS ordinati on scores based on avian composition at the 111 sites. Axis 1 had the highest adjusted R2 (65.1) among the three axes. LDI, percent canal/retention pond, longitude, and latitude were se lected as predictors of axis 1. Axis 2 was best represented by the percent agricultu re within 200 meters of the focal wetland (R2=25.0). Likewise, axis 3 only had one pr edictor, LDI, with an adjusted R2 of 22.3.

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102 Table 3-29 shows the results of stepwi se regression using 28 independent variables (Table 3-27) to represent the th ree avian composition based NMS ordination axes. The comparison of Table 3-28 with Table 3-29 shows that none of the additional 14 Focal Wetland or Sampling Effort variable s (Table 3-27) improved the adjusted R2 of axis 1. The adjusted R2 of axis 2 and 3 improved 9% and 12% respectively, with the addition of the focal wetland and attributes of sampling independent variables. Axis 2 retained the Landscape variable percent agri culture and gained percent evergreen tree species in the focal wetland and the total amount of time spent surveying. Axis 3 had two predictors selected, FWCI and tota l amount of time spent surveying. Table 3-28. The selected vari ables and regression coefficien ts selected (p<0.05) from a stepwise regression using the Landscape variables listed in Table 3-27 to represent the avian based NMS ordination axes. Selected Variables axis 1 axis 2 axis 3 LDI -0.20 0.12 % Agriculture (200m) -0.81 % Canal (200m) -3.2 Longitude 0.12 Latitude 0.14 Adjusted R-square 65.06 25.01 22.28 Table 3-29. The selected vari ables and regression coefficien ts selected (p<0.05) from a stepwise regression using th e Landscape, Focal Wetland, and Sampling variables listed in Table 3-27 to represent the avia n based NMS ordination axes. Selected Variables axis 1 axis 2 axis 3 LDI -0.20 FWCI -0.02 % Agriculture (200m) -0.73 % Canal (200m) -3.2 % Evergreen Trees (focal wetland) 0.84 Longitude -0.12 Latitude 0.14 Survey Length (min) -3.4E-03 -3.5E-03 Adjusted R-square 65.06 34.22 33.82

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103 Bird Indicator Species Table 3-30 shows the signifi cant (p<0.1) indicator values for avian species by land use. Eleven species had their highest indi cator values for agriculture sites and cattle egrets had the highest value. Eight species had significant indicator values for reference sites with Bachman’s sparrow and common ye llowthroat sharing the highest indicator values. Bachman’s sparrow showed strong fi delity to reference sites while the common yellowthroat also had a high signi ficant indicator value at silv iculture sites. Only two other species had significant indicator values at silviculture sites. Seven species had their highest indicator values at urban sites. The northern mockingbird had the highest indicator value among all species for all land uses at urban sites. Table 3-31 shows the signifi cant (p<0.1) indicator values for avian species by LDI quartiles. LDI quartile one had nine avian sp ecies selected as indi cators (p<0.1). The common yellowthroat and the rufous-sided towh ee had the highest indicator values for LDI quartile one. The common yellowthroat and the rufous-sided towhee also had relatively high values for quartile 2. The rema ining seven species selected as indicators of LDI quartile one had relatively low scores for the remaining quartiles. Three species were selected as significant indicators of LDI quartile 2 and four species for LDI quartile 3. The highest indicator values for quart ile two and three were well below that of quartiles one and four. Eight species were se lected as significan t indicators of LDI quartile 4.

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104 Table 3-30. The significant (p<0.10) av ian indicator values of land use. Species R S A U p-value Common Yellowthroat 29 29 1 0 0.0085 Bachman's Sparrow 29 2 0 0 0.0024 Pine Warbler 22 4 2 1 0.0124 Common Nighthawk 22 5 0 1 0.0276 Brown-headed Nuthatch 17 2 0 0 0.0208 Barred Owl 14 0 5 0 0.0423 Prothonotary Warbler 12 0 0 0 0.0581 Red-cockaded Woodpecker 9 0 0 0 0.0651 Rufous-sided Towhee 28 34 6 4 0.005 American Swallow-tailed Kite 0 12 2 0 0.0893 Cattle Egret 0 0 54 2 0.0001 Red-shouldered Hawk 12 8 24 3 0.0372 Turkey Vulture 5 7 21 1 0.0393 American Crow 3 14 21 5 0.0567 Red-winged Blackbird 5 0 20 2 0.0393 White Ibis 0 0 20 1 0.0097 Blue Grosbeak 1 6 16 0 0.0254 Eastern Meadowlark 4 0 14 0 0.0765 Black Vulture 3 2 14 0 0.052 Wild Turkey 1 0 12 0 0.0487 Barn Swallow 0 3 12 0 0.0218 Northern Mockingbird 0 0 12 61 0.0001 Blue Jay 1 7 17 33 0.0046 Fish Crow 2 0 5 27 0.0035 Common Grackle 1 0 19 26 0.0261 House Finch 0 0 1 17 0.0139 Eurasian Collared-Dove 0 0 2 15 0.0778 House Sparrow 0 0 2 10 0.0798

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105 Table 3-31. The significant (p<0.10) avia n indicator values of LDI quartiles. Species Q1 Q2 Q3 Q4 p value Common Yellowthroat 39 13 0 0 0.0001 Rufous-sided Towhee 35 21 10 1 0.0001 Bachman's Sparrow 32 1 0 0 0.0001 Common Nighthawk 29 2 1 1 0.0001 Pine Warbler 29 2 3 1 0.0001 Brown-headed Nuthatch 23 1 0 0 0.0004 Barred Owl 17 0 3 0 0.0120 Prothonotary Warbler 14 0 0 0 0.0125 Red-cockaded Woodpecker 10 0 0 0 0.0583 White-eyed Vireo 8 20 11 0 0.0285 American Swallow-tailed Kite 0 12 1 0 0.0208 Barn Swallow 0 10 2 0 0.0396 Cattle Egret 0 4 25 4 0.0011 American Crow 4 8 21 5 0.0326 Turkey Vulture 4 11 18 1 0.0513 Wild Turkey 0 1 11 0 0.0390 Northern Mockingbird 0 3 19 49 0.0001 Common Grackle 0 3 12 34 0.0002 Blue Jay 1 8 23 30 0.0022 Fish Crow 1 1 6 27 0.0010 Eurasian Collared-Dove 0 0 3 18 0.0025 Boat-tailed Grackle 0 0 6 14 0.0181 House Sparrow 0 0 4 12 0.0287 Muscovy Duck 0 0 0 11 0.0262 Bird Community Characteristics Figure 3-12 compares the proportion of species found at each site belonging to four a priori feeding guilds among the four LDI quart iles. There is a general trend as LDI increases the proportion of the avian community that was primarily insectivorous decreases (pearson correlation coefficient r=-0.57) while the proportion that were omnivorous or herbivorous increases (r=0.45 and 0.44 respectively). Insectivores decrease from approximately 51% in LDI quartile 1 to 29% at LDI quartile 4. Omnivores

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106 increase from 31% of the avian community at LDI quartile 1 sites to 49% of the species at LDI quartile 4 sites. There was no signi ficant trend for the pr oportion of carnivorous species by LDI quartile (r=-0.12). However, LDI quartile four had the lowest proportion of carnivorous species. Approximately 89% of the insectivorous sp ecies encountered maintain intraspecific feeding territories. The insectivorous species that were not territo rial were primarily aerial screeners (e.g., swallows ) (DeGraaf et al. 1985). The proportion of omnivores, carnivores, and herbivores that maintain in traspecific feeding territories was 52%, 38%, and 18% respectively. The relationship between territoriality a nd land use is shown in Figure 3-13. The mean proportion of birds that maintain feedi ng territories declines with increasing LDI quartile. There was a 20% decrease from appr oximately 80% to 60% territorial species from LDI quartile 1 to LDI quartile 4. The proportion of territorial bird species by land use shows there was no significant difference between urban and agriculture sites nor reference and silviculture sites. The foraging strategy of the bird comm unity changes with land use and along the LDI gradient (Table 3-32). The proportions of bird species that bark or canopy glean decreases, and ground gleaners in crease with increasing land use intensity. However, there is not a significant di fference between the higher inte nsity land uses (agriculture and urban) and lower intensity land uses (silviculture and refe rence). Aerial screeners are insectivores that do not re ly directly on tree cover fo r foraging, but there was significantly greater propor tion at reference sites than agricu lture sites. Despite a wetland

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107 Quartile 1herbivore 5% omnivore 31% carnivore 13% insectivore 51% a c e h Quartile 2insectivore 46% herbivore 6% omnivore 35% carnivore 13% a cd ef h Quartile 3insectivore 37% herbivore 10% omnivore 39% carnivore 14% b d efg h Quartile 4insectivore 29% herbivore 14% omnivore 49% carnivore 8% b g h Figure 3-12. The mean proportion of avian sp ecies within each feeding guild by LDI quartile. Shared letters within feeding gu ilds were not significantly different (one-way ANOVA Tukey multiple comparisons (p<0.05)).

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108 0 0.1 0.2 0.3 0.4 0.5 0.6 0.7 0.8 0.9 1 Q1Q2Q3Q4rsauProportion of Territorial Birds aabbab ab Figure 3-13. The mean and standard deviat ion of the proportion of avian species that maintain feeding territories at each site by LDI Quartile and Land Use. Shared letters were not significantl y different (One-way ANOVA, Tukey multiple comparisons, (p<0.05)). Table 3-32. Mean proportion of sampled bi rd species predominate foraging strategy by land use and LDI quartile. Shared letter superscripts indicate no significant difference at the 95% confidence interv al within a guild by land use or LDI quartile (Mann Whitney U-Test). bark gleaner canopy gleaner ground gleaner hawker sallier scavenger screener water forager R 0.14b 0.31b 0.30b 0.06ab 0.04a 0.03a 0.07b 0.05a S 0.13ab 0.37b 0.30b 0.05ab 0.06a 0.03a 0.05ab 0.01b A 0.10a 0.24a 0.41a 0.06a 0.04a 0.06a 0.03a 0.06a U 0.10a 0.23a 0.45a 0.04b 0.03a 0.05a 0.05ab 0.05ab Q1 0.15a 0.32a 0.29a 0.05a 0.03a 0.03a 0.07a 0.06a Q2 0.13ab 0.30ab 0.33a 0.06a 0.04a 0.04a 0.06ab 0.04a Q3 0.10b 0.24bc 0.43b 0.05a 0.04a 0.06a 0.03b 0.05a Q4 0.10b 0.22c 0.48b 0.04a 0.03a 0.05a 0.04ab 0.05a

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109 being the focal point of every study site the mean proportion of sampled bird species that were water based foragers was less then 6% for all sites combined. The proportion of species sampled that we re cavity or ground nesters decreased with increasing land use intensity (Table 3-33). Species that nest in the upper or lower canopy increased proportionally with increasi ng land use intensity. The relative amount of shrub nesters had a weak trend with increasing land use in tensity and was only significant between LDI quartiles one and four or between refe rence and urban sites. Table 3-33. Mean proportion of sampled bi rd species predominate nesting strategy by land use and LDI quartile. Shared letter superscripts indicate no significant difference at the 95% confidence interval within a nesting strategy by land use or LDI quartile (Mann Whitney U-Test). ground shrub canopy cavity human structure paracite R 0.21a 0.14a 0.26b 0.36b 0.03a 0.00a S 0.16a 0.20ab 0.28b 0.32bc 0.03a 0.01a A 0.16a 0.16a 0.40a 0.24a 0.03a 0.01a U 0.09 0.19b 0.41a 0.25ac 0.06a 0.01a Q1 0.21a 0.14a 0.26a 0.36 0.02a 0.00a Q2 0.17ab 0.18ab 0.30a 0.30a 0.04a 0.00a Q3 0.14b 0.17ab 0.39b 0.25ab 0.04a 0.01a Q4 0.08 0.19b 0.44b 0.23b 0.06a 0.01a The mean proportion of the bird species that were exotic increases from zero for LDI quartiles 1 and 2 to three percent for LD I quartile 3 and nine percent for LDI quartile 4. Seven of the eight exotic species encountered were he rbivorous and the remaining species, European Starling, was omnivorous. Urban sites had the highest proportion of exotic species of the four land uses. Only the proportion of Neotropical migr ant bird species at LDI quartile 2 was significantly different than LD I quartile 4 (Figure 3-14). Similarly, amongst the land

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110 uses only the silviculture sites had a si gnificantly higher propor tion of Neotropical migrants than urban sites. Approximately 81% of the Neotropical migrants were insectivorous, while, 9% we re either carnivorous or omnivorous, and there were no herbivorous Neotropical migrants sampled. Two of the three carnivorous Neotropical migrant bird species rely heavily on large insects (Mississippi Kite and American Swallow-tailed Kite). Finall y, 79% of the Neotropical migr ant bird species encountered maintain feeding territories. 0 0.05 0.1 0.15 0.2 0.25 Q1Q2Q3Q4rsau % exotic % neotropical migrant aa aaa abababaabb ab Figure 3-14. The mean proportion and standa rd error of the exotic and Neotropical migrant birds recorded by LDI Quartile and Land Use. Shared letters were not significantly different (arcsin square root transformation of the of % exotic species)(One-way ANOVA, Tukey multiple comparisons, (p<0.05)). Figure 3-15 shows the mean proportion of bi rd species detected at the sites that have had significant population increases or decreases in the Breedi ng Bird Survey data since 1966 in Florida. The mean proportion of declining specie s decreases with increasing LDI quartile, and the mean proporti on of increasing speci es increases with increasing LDI quartile. The differences in the means were only significant between the

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111 ends of the LDI quartiles. Similarly, the mean proportion of decr easing species for the low intensity (silviculture and reference site s) and high intensity (a griculture and urban sites) land uses were not significantly di fferent. Only the reference sites were significantly different than the ur ban sites for the mean proportion of increasing species. 0 0.1 0.2 0.3 0.4 0.5 Q1Q2Q3Q4rsau % Decreasing Species % Increasing Species aabbbbb a a aabbccaababb Figure 3-15. The mean proportion and standard e rror of bird species th at have significant (p<0.1) population trends in the Breedi ng Bird Survey data for Florida since 1966. (arcsin squareroot transformation of % increasing species)(One-way ANOVA, Tukey multiple comparisons, (p<0.05)). Figure 3-16 shows the mean weight of the sp ecies detected at the four LDI quartiles and land uses. There was not a significant difference in the mean weight of species detected at the sites when separated by LDI qua rtiles. However, the mean weight of the species detected at agricultural sites was si gnificantly greater than the remaining three land uses. Furthermore, seven of the eleven avian indicator species of agricultural sites weigh more than 300 grams (Table 3-30). In contrast urban sites had no avian indicator over 300 grams, and reference and silv iculture sites each had only one.

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112 0 100 200 300 400 Q1Q2Q3Q4rsauMean Avian Species Weight (g) a aa a a a a Figure 3-16. The mean weight (g) and standa rd error of avian species detected at the study sites by LDI quartile and Land use. Shared letters are not significantly different (Natural log of mean av ian species weight by site, One-way ANOVA, Tukey multilpe comparisons, (p<0.05)). Generally, larger species maintain larger territories or have larger home ranges than smaller species (Figure 3-17). The largest species home range sizes are several orders of magnitude greater th en the size of the sample sites. Meanwhile, the smallest species territories are an order of ma gnitude smaller then the sampled areas. r2 = 0.5341 0 0.5 1 1.5 2 2.5 3 3.5 4 -202468 Territory or Home Range log10(ha)Species Mass log10 (g) Figure 3-17. Log10 of avian mass versus log10 of avian territory or home range size for 55 species sampled (Schoener 1968).

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113 CHAPTER 4 DISCUSSION This study of avian and amphibian speci es composition in small depressional wetlands of Florida was conducted to eval uate species compositional response to development intensity. Wetlands embedde d in landscape matrices of varying development intensity have amphibian and avian species compositions that correspond with the degree of anthropogenic land use in tensity in the circumjacent landscape. In overview, the following are the ma jor findings of this study: 1) Species richness was weakly associated with land use intensity and land use type, although amphibian richness wa s highest at the lowest land use intensity sites. 2) Differences in richness within wetl ands compared to the circumjacent landscapes indicated there are effect s related to development intensity, where the wetlands suppor t proportionally more sp ecies at sites surrounded by the highest intensity land uses. 3) There were notable differences in sp ecies composition and guild structure of the amphibian and avian communities that followed an intensity gradient from reference to silvicu lture to agriculture to urban. a. Amphibians that were obligatory ephemeral pond breeders and those that have submerged attached eggs decrease with increasing land use intensity.

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114 b. Avian feeding, foraging, nesting, an d population status guilds responded to land use intensity. 4) Variables other than land use intens ity were found to be significantly associated with amphibian a nd avian species composition a. Amphibian species composition was a ssociated with FWCI, distance to the nearest wetland, maximum wetland water depth, wetland water pH, wetland water specific conductivity, and geographical location of the study sites. b. Avian species composition was associ ated with proportion agricultural land within 200m, proportion of re tention/canals w ithin 200m, FWCI, and the geographical location of the study sites. 5) A set of amphibian and avian indicator species of land use type or land use intensity were identified and potenti al explanatory mechanisms were discussed. 6) The LDI index proved to be a, remote ly measured, predictor of species assemblages. Species Richness Species richness was weakly associated w ith land use intensity. Reference site amphibian species richness was significantly grea ter than at urban site s, and significantly greater at the lowest intensity sites (LDI Q1). Bird species richne ss did not correspond to land use intensity and only agricultural si tes had significantly gr eater species richness than urban and silviculture sites. Prev ious studies have a found a pattern between amphibian and avian species richness with anthropogenic land use (e.g., Delis et al. 1996, Findlay and Houlahan 1997, Rottenborn 1999) wh ile others were unable to detect a

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115 significant trend (e.g., Rodewald and Yahne r 2001, Norris et al. 2003, Gray et al. 2004, Knutson et al. 2004). Species richness woul d be an intuitive measure of anthropogenic induced habitat change; however, this study de monstrates species composition changes in a more predictable manner than does speci es richness. Genera lly, anthropogenically induced changes to habitats caused transi tions in species occupation resulting in imperceptible changes to species richness. Some species were excluded while others with congruent life history ch aracteristics persisted or coloni zed anthropologically altered habitats. Blair (1996) found higher avian species rich ness at urban sites with intermediate levels of human development and suggested this pattern followed the Intermediate Disturbance Hypothesis (IDH). Similarly, this study f ound significantly higher avian species richness at agricultural sites, which were intermedia te along the land use intensity gradient studied. Amphibian species richne ss on the other hand, was highest in areas receiving the least amount of anthropogenic disturbance (refe rence sites and LDI quartile 1) and did not appear to follow the IDH. Ev en sites with low to moderate levels of anthropogenic land use intensity (e.g., LD I Q2) had significantly lower amphibian species richness than the least impacted sites. Landscape Versus Wetland Species Richness Amphibian and avian species utilizati on of small forested wetlands is disproportionately higher in developed landscape s. There were large differences in the number of species recorded inside the samp le wetlands versus the surrounding landscape between lower and higher intensity land uses despite the weak relationship between total species (within the surrounding landscape and the sample wetland combined) detected and land use. Nearly two times as many amphi bian species were found in urban wetlands

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116 versus the circumjacent landscape while th e ratio was one to one for agriculture, reference, and silviculture sites. The hi gher wetland to landscape ratio of amphibian species richness in urban environments was mo st dramatic for the highest intensity urban sites. The mean amphibian species richness inside the wetland versus the landscape was nearly 6 to 1 at the nine sites with highest LDI scores (LDI range 7.0-7.8). Likewise, avian richness was 1.7 times greater in the surrounding landscape than in the sample wetland at urban sites, while the ratio was much higher at reference and silviculture (3.2) sites and slightly higher at agriculture sites (2.1). This evidence suggests at lower intensity sites species ar e using wetlands and th e surrounding landscape interchangeably whereas in highly developed landscapes small forested wetlands support disproportionately more biodiversity relative to the su rrounding landscape. It is common to have high avian abunda nce in the urban matrix; however, species richness is generally lower than more continuous reference habitats (e.g., Emlen 1974, Clergeau et al. 1998, Sorace 2002). In landscap es that are being actively fragmented, mobile species may move from the habita t disturbance in the surrounding landscape to the remaining forest temporarily (approxima tely 200 days) increasi ng the observed avian density of the remnant hab itat patch (Bierreg aard et al. 1992, Hagan et al. 1996). Amphibian and avian species richness within study wetlands was greater relative to the surrounding landscape in long established (>200 days) urban areas than wetlands embedded in landscapes of lower land use intensity. Crooks et al. (2004) also found higher avian species richness in forest fragme nts than more continuous reference habitats and the urban matrix.

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117 The increase in avian species richness in the forested wetlands within urban areas in this study was primarily because of the retention and/or addition of edge, generalist, and exotic species, but many of the original reference habitat specialists were absent (Murphy 1989, Croonquist and Brooks 1991, Rodewald and Yahner 2001, Blair 1996, Bennett et al. 2004). The “concentration eff ect” arises when edge species have few options but to nest in isolated fragments of developed landscapes (Bennett et al. 2004). Likewise, generalist and exotic amphibians (e.g., Bufo terrestris Eleutherodactylus planirostris and Osteopilus septentrionalis ) were also more prev alent at the higher land use intensity sites while many species sensitiv e to development were largely absent (e.g., Acris gryllus Bufo quercicus Hyla gratiosa and H. femoralis ). Wiens (1995) suggested that sedentary avia n species are particul arly sensitive to habitat fragmentation. Contrary to this, Gibbs (1998) and Carr and Fahrig (2001) found that some of the more sedentary amphibian species are less suscepti ble to human induced habitat fragmentation. Less va gile, often aquatic, amphibian species may persist in, or even benefit from, wetlands adjacent to high intensity land uses while many of the more terrestrial and arboreal species dependent on upl and habitats are significantly affected by alterations to the surrounding landscape. Ho wever, in this study the amphibian species most frequently encountered at the highest land use intensity s ites are generally not considered poor dispersers (e.g., Hyla cinerea H. squirella and Rana sphenocephala ). More likely, the landscape surrounding wetlands at the most devel oped sites is lacking the requisite resources for the species typi cally only found in less developed landscapes (e.g., Marnell 1998). From a landscape pers pective, the lower proportion of wetlands within 200m and a greater dist ance to the nearest neighbor wetlands at urban sites may

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118 have also influenced the number of amphibi an species detections in the circumjacent habitat relative to the reference sites (Table 3-5). Species Composition Land Use Type and Intensity Amphibian and avian species compositi on was significantly different between land uses and land use intensitie s. The strongest differences were between sites at the opposite ends of the land use intensity gradient For example the strongest differences as determined by the MRPP tests for both amphibians and birds were between reference and urban sites or LDI quartile one and LDI quartile four. However, both amphibian and avian speci es composition were not significantly different between reference and silviculture sites. Similar studies have reported no observable changes in species composition with some silvicultural practices (e.g., Enge 1984, Renken et al. 2003). There are many poten tial explanations for this observed trend, the most parsimonious being silviculture site s have the least amount of habitat change relative to reference sites (e.g., water chemistry, vegetati on structure and composition, etc.) amongst the anthropogenica lly influenced landscapes. Not only is the magnitude of nonrenewable energy use less in silvicultural landscapes, but also the temporal frequency of disturbance is often less than other human dominated la ndscapes. The disturbance in silviculture landscapes comes in severa l distinct pulses (e.g., tree-cutting, site preparation, herbicide applica tion, etc.) and in the interim these areas receive minimal human disturbance. Second, many of the sites classified as silviculture were embedded in large conservation areas, and several of which were in public ownership where self imposed best management practices (BMPs) are commonplace. In turn, some area

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119 sensitive species (e.g., American swallow-tail ed kite) may be able to find essential resources in expansive silvicultural landscapes (Robins et al. 1989). It should be noted that the relatively brief site visits conducted in this study may not be conducive to detecting the presence of many rare and cryptic species (e.g., Pellet and Schmidt 2005), and it is these species that have been suggested to be sensitive to common silvicultural practices (e.g., Means et al. 1996). Breeding phenology may also play a factor in the lack of detection of many rare amphibi an species in this study. For the most part none of the amphibians that are restricted to breeding in autumn, winter and spring were encountered at the sites during these summer surveys, and many species within in this group are cited as being ra re and particularly sensitive to habitat disturbance (e.g., Rana capito and Ambystoma cingulatum ) (e.g., Godley 1992, Palis 1997, Hipes et al. 2000). Amphibian species known to occur in Florida temporary wetlands and are primarily active (i.e., breed) from fall to early spring (Moler and Franz 1987, Bartlett and Bartlett 1999) that were not encountered at any sites during the 2003 sampling included; Pseudacris ornata P. crucifer P. nigrita Ambystoma cingulatum A. tigrinum and Eurycea quadridgitata or at a low number of sites Rana capito (2 sites) Ambystoma talpoideum (1 site). Species Composition Thresholds and Gradients Amphibian community response to land us e reached a threshol d while the avian community changed along a gradient of land us e intensity. The results of the MRPP tests revealed amphibian species composition wa s not significantly diffe rent between urban and agricultural sites as well as between LDI quartiles three and four. This was not the case for bird species assemblages, where, composition was distinct between agricultural and urban sites and between the hi gher intensity LDI quartiles.

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120 Defining the threshold amount of devel opment that species can withstand is important to those interested in incorpor ating wildlife conservation into landscapes utilized by humans (e.g., Environmental Law Institute 2003, Radford and Bennett 2004). Much of the discussion on thresholds is concerned with the minimum patch area, proportion of critical habitat in a landscape or the maximum distance to critical habitat that a species will disperse, in order for a population to persist in a landscape (With and Crist 1995, Radford and Bennett 2004). The re sponse of species assemblages in this study was dependent on land use type and in tensity implying that critical habitat thresholds for individual species may also vary depending on land use intensity of the matrix habitats. Individual avian species may indeed exhibi t a threshold response to development. Different degrees of development may enable some species to colonize while others do not persist, and thus the trend is for di stinct species assemb lages along the land use intensity gradient and between land use types. As a community, bird response to land use intensity or land use type did not reach a th reshold response, as defined as a marked response of a biotic community to human de velopment (Miller et al. 2003). Amphibian assemblages on the other hand were not distin ct between higher in tensity land uses and thus may have reached a threshold response to land use intensity. There are several potential explanatory mechanisms for the differences in the response to land use intensity between the amphibian and avian communities and these include, differences in species richness, ha bitat sensitivity, and metapopulation dynamics. There is a small amphibian species pool avai lable to respond to the higher intensity land uses, and in turn a limited number of edge and invasive amphibians capable of colonizing

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121 wetlands within high intensity land use types. Birds in Florida have a larger species pool and thus there is the potential for more sp ecies being indicative and adapted to more human influenced landscapes. Amphibians as a group, because of their uni que life history characteristics, may be more sensitive to habitat alteration (e.g., Dodd 1997, Sparling et al. 2001), and agriculture and urban environm ents may exert similar selection pressures on amphibian communities. For example many of the water chemistry parame ters and the proportion of canal/retention pond within 200 m of the sample wetlands were similar between agricultural and urban sites. The persistence of a wildlife population may be dependent on recruitment of individuals from nearby populations (Fahri g and Merriam 1985). Particularly for amphibians, the type and quality of neighbor ing wetlands could have an affect on the metapopulation dynamics and in turn the observed species composition within a landscape. In urban landscapes the nearest we tlands to the sample wetlands tended to be farther away and in many instances the quality and type of nearest neighbor wetlands in agricultural and urban landscap es was unknown. At the silvic ultural and reference sites the nearest neighbor wetlands were predominat ely of the same type, small depressional forested wetlands. Agricultural and urban landscapes may also be less permeable to dispersal by some amphibian species with ex ception of those capable of utilizing canals and retention ponds or other attributes of developed land uses (e.g., Ficetola and De Bernardi 2004). Many avian species utilize attributes specific to individual land use types (e.g., buildings for nesting, open fields for foraging, etc.) and as a result unique avian species assemblages were observed.

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122 Finally, there has been li ttle attention given to th e wildlife composition of landscapes at the low end of the human set tlement gradient (Mill er et al. 2001, Hermann et al. 2005). Likewise, simila r studies have found it difficult to differentiate sites at the low end of the disturbance gradient based on biotic composition (e.g., Bradford et al. 1998, O’Connell et al. 2000). Although not specif ically a human settlement gradient this study was organized by a human land use intens ity gradient and showed that amphibian and avian species composition was significantl y different between the low ends of the LDI index gradient. That is, the least impact ed sites had significant differences in avian and amphibian species composition compared to sites receiving moderate amounts of human influence (LDI quartile 1 versus LDI quartile 2). Regardless whether the community response is in the form of a gradient or exhibits a threshold response to human land use intensity th is study demonstrates that species composition varies by land use intens ity and suggests some land uses are more compatible with maintaining the species composition of reference condition forested wetlands. Individual sites within the develope d land use types had qualities that allowed the persistence of some of the species observed primarily in reference landscapes; however, when the sites were grouped by land use or land use intensity the trend was for distinct species assemblages between low and high human land use intensity. The binary comparisons of the MRPP tests reveal that ther e is an order of land use compatibility in which reference condition species composition is most similar to silviculture then agriculture and finally urban landscapes. Wh en ranked by the LDI index the relationship between species composition and land use intensity is even stronger.

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123 Correlates of Species Composition Amphibian and avian assemblages were not randomly distributed among the study sites. Ordination results indi cated there were distinct patterns within species assemblages that corresponded with ch aracteristics of the study sites. Several attributes of the wetland, surrounding landscape, and the regiona l distribution of the study sites were significantly correlated with the ordination axes. Correlates of amphibian species composition Stepwise regression results indicated am phibian assemblages were associated with estimates of land use intensity (L DI) and wetland quality (FWCI), area of agricultural land use, distance to the nearest wetland, maxi mum wetland water depth, pH, conductivity, proportion of the wetland basal area consisting of exotic trees, and geographical location of the study sites. Amphibian assemblages may be strongly in fluenced by characteristics of water bodies within the landscape because of their bi phasic life style. Hydroperiod, presence of aquatic predators, water quality, and wetland in solence have all been implicated as being important in determining amphibian species composition (Moler a nd Franz 1987, Kats et al. 1998, Houlahan and Findlay 2003, Ficetol a and De Bernardi 2004). Many of the variables selected in the stepwise regression analyses of the amphibian ordination axes in this study were related to characteristics of both the sample wetland and the surrounding landscape. The quality of the sample wetland as estim ated using the FWCI had the strongest univariate correlation with amphibian sp ecies composition. This suggests that composition of the amphibian community is strongly correlated with compositional changes in the diatom, macrophyte, and aqua tic macroinvertebrate communities, and the

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124 FWCI scores to a large degree represent th e quality of habitat for amphibians. An advantage of the FWCI is it may have the ability to represent wetland characteristics attributable to direct anthr opogenic effects and historical land use, manifested in the biotic community not represented by simply meas uring the current surroun ding land uses. The LDI index was also selected as a strong predictor of amphibian species composition. Houlahan and Findlay (2003) found a stronger associ ation between the proportion of the surrounding landscape that was undeveloped, forest or wetland, as a predictor of amphibian species assemblages than the amount of agricultural or urban land use surrounding sample wetlands. In this study LDI was selected as a better pred ictor of amphibian species assemblages than the amount of undeveloped land, including wetlands, within 200m. However, the proporti on of agricultural, and at times the proportion of urban, land use along with the LD I index were selected as predictors of amphibian species assemblages suggesting th e affect of these landscapes may not be completely represented by the LDI index. The maximum water depth was the strongest univariate co rrelation with axis 1 of the amphibian based NMS ordination results. Water depth is seldom selected as an indicator of amphibian species assemblages (f or an exception see, Knapp et al. 2002). Two factors may be responsible for the obser ved correlation. First, the maximum water depth variable included sites th at did not have standing water. The absence of standing water has a strong influence on the detection of amphibian species (Vic kers et al. 1985). Second, maximum water depth may be corr elated with wetland hydroperiod, which has been found to be an important factor dete rmining amphibian species composition (e.g., Kolozsvary and Swihart 1999, Snodgrass et al. 2000, Houlahan and Findlay 2003,

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125 Herrmann et al. 2005). Furthermore, many of the sample wetlands that were dry or had relatively shallow surface waters were adjace nt to retention ponds and/or canals with their concomitant assemblage of amphibian s contributing to the sites total species assemblage. Distance to the nearest wetland was another factor selected as being associated with amphibian species composition. Amphibi ans are the most wetland dependent and are generally the less vagile of the two taxa studied. Cri tical distance thresholds and barriers to dispersal can ha ve a profound effect on populat ion persistence (Gilpin and Hanski 1991), and in turn, the distance to the nearest wetland(s) and proportion of wetland area in the landscape have been ci ted as being important in determining amphibian species composition, richness, a nd abundance (Knutson et al. 1999, Lehtinen et al. 1999, Mensing et al. 1998, Russell 2002). Qualities with in the upland matrix may influence a species ability to disperse th rough it, where speciesspecific effective dispersal distances are variable and influen ced by attributes within the landscape (Laan and Verboom 1990, With and Crist 1995). For ex ample, many of the species found at the developed sites are species known to utilize human structures and other manmade habitat features (e.g., canals, buildings, hedgerows, gardens, etc.) Although nutrients have been linked to amphibian species composition (Mensing et al. 1998, Houlahan and Findlay 2003, Knuts on et al. 2004) total phosphorus, ammonia, TKN, and nitrite-nitrate-nitrogen of the we tland waters was not included as a significant predictor of amphibian species composition in the multiple stepwise regression models. At the same time, there were significant differences in many of the wetland water

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126 chemistry parameters between land use types, including TKN, ammonia, total phosphorus, pH, dissolved oxygen, turbid ity, and specific conductivity. The pH and conductivity of the sample we tland waters were selected as being significantly associated with amphibian spec ies composition in this study as well as by Bunnell and Zampella (1999). Especially in pineland habitats, the specific conductivity and pH of wetland waters increase with in creasing watershed development intensity (Zampella 1994, Dow and Zampella 2000). An increase in the pH of acidic wetland waters affects competition for resources among amphibian species and ultimately species composition (Warner et al. 1993, Warner and Dunson 1998). However, changes in wetland pH are often associated with changes to the upland vegetation compositi on. Pine forests su rrounding wetlands are usually removed during the development proce ss, and wetlands within pineland habitats are often acidic. It is difficult to dete rmine to what degree amphibian species are responding to the loss of the upland pine forest s and/or the associated changes to wetland water chemistry. Maintaining osmotic balance in ionic rich water may increase the energy demands of larval amphibians (Rowe et al. 2003). However, wetland specific conductivity not only increases with development intensity but also wetland connec tivity (Dunson et al. 1997, Dow and Zampella 2000). Again, it is di fficult to separate whether this water chemistry parameter is simply correlated with other aspects of th e environment shaping amphibian species assemblages, or the sel ected water chemistry parameters directly influenced species composition. Wetland c onnectivity increases the likelihood that a wetland will contain fish (Babbitt and Tanner 2000). Increased surface water

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127 connectivity as represented by increased specific conductiv ity could be a surrogate measure of wetland exposure to predatory fi sh (Dunson et al. 1997), and fish are known to have an influence on amphibian species assemblages (Kats et al. 1988, Babbitt and Tanner 2000, Reshetnikov 2004). Despite, fish being caught during the dip net sweeps this methodology may not have accurately sample d this group and their true affect on the amphibian composition in this study is largely unknown. More then 80% of the sample wetland s were observed without standing water prior to the 2003 field season. The sample we tlands in this study were primarily small isolated wetlands yet more then 50% cont ained fish in 2003, including the generally unadulterated reference set. This suggests that wetland water chemistry, fish occurrence, and amphibian composition are a temporally dynamic process of depressional wetlands of the southeastern coastal plain. Correlates of avian species composition Attributes of the sample wetland appeared to be less important than aspects of the surrounding landscape in influencing avian sp ecies composition. Significant variables selected in the multiple regression models of avian species composition included, LDI, proportion agriculture within 200m, propor tion retention pond/canals within 200m, FWCI, the proportion of evergreen trees with in the sample wetland, and the geographical location of the study sites. The LDI index was again selected as a significant predictor of species assemblages. Several studies have found a relationship between the proportion of undeveloped, or forested land, and avian species composition in human influenced landscapes (e.g., Estades and Temple 1999, D onnelly and Marzluff 2004). However, the LDI index was selected over the proportion of undeveloped land in the multiple stepwise

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128 regression analyses. Similarly, Rodewald and Yahner (2001) found that the landscape disturbance type (i.e., agricultu re or silviculture) was more important in influencing avian species composition than land use type extent Furthermore, previous studies have indicated that avian communities respond to a human land use intensity gradient within broad land use categories (e.g., silviculture agriculture, urban) (Harris et al. 1975, Cutright 1981, Blair 1996, Bradford 1998, Titt ler et al. 2001, Naidoo 2004, Waltert et al. 2004) and this study demonstrates a response gradient across land use types to land use intensity. At the same time, the proportion of agriculture was selected independently from LDI as explaining the variation in one of the NMS ordination axes indicating, there is a unique avian assemblage associated w ith agricultural land use in Florida. The proportion of retention pond and canal within 200m of the sample wetland was associated with the avian species compos ition. It appeared that often these water control devices were partia lly responsible for reducing the hydroperiod and standing water levels in the sample wetlands. The wate r may have been redirected and retained in the adjacent canals and retention ponds. Thes e manmade habitats generally had deeper water levels and marsh vegetation, which attrac ted a particular suite of avian species. As with the amphibians, the geographical coordinates of th e sites were selected as predictors of the observed avian species as semblages (Parris 2004). The most simple explanation for this trend is that many of the amphibian and avian species ranges lie predominantly in either north or south Florida, or are restricted to the panhandle (western portion of Florida). Indicator Species A set of amphibian and avian species were selected as being si gnificant indicators based on frequency of occurrence and fidelity to specific land use types or levels of land

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129 use intensity. Species not selected as indica tors but were frequen tly encountered at the study sites might be considered ub iquitous and tolerant to land use type and intensity. Amphibian Indicator Species There were ten significant amphibian indi cators of land use (p<0.1), six species were sensitive to land use intensity, and four species were associated with development. The results of the indicator species analyses revealed a nd mirrored the results of the MRPP results, in that particular amphibian species were indicative of primarily either high or low intensity land uses. The indica tor values for the species indicative of low intensity land uses were higher than the indi cator values for the species indicative of high intensity land uses. Three of the five species Delis (1996) listed as sensitive to development, Hyla femoralis, H. gratiosa and Bufo quercicus were also selected as indicators of low intensity land use in this study. However, Delis (1996) found greater abundances of ranids in urban environments, including Rana grylio which was indicative of low intensity land uses in this presence/absen ce study. Similarly, Mushinsky et al. (2001) reported Hyla femoralis and Bufo quercicus as reference conditi on indicators, however Acris gryllus was cited as tolerant to the modified habitat in their study. At the forested depressional wetlands in th is study, the presence of Acris gryllus was inversely correlated to land use intensity. Four of the six species indicative of reference si tes were included in a list of longleaf pine spec ialist by Means et al. (2004), Bufo quercicus Hyla femoralis H. gratiosa and Pseudacris ocularis Finally, Wilson and Porr as (1983) identified many of the same amphibian species, that are sensitive to land use intensity, Bufo quercicus, Hyla femoralis, Hyla gratiosa, Acris gryllus and Pseudacris ocularis and species found in association with human land us es as well as natural habitats, Eleutherodactylus

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130 planirostris, Osteopilus septen trionalis, Gastrophryne carolinensis, Rana sphenocephala, Hyla cinerea and Hyla squirella Amphibian populations appear to respond to features of the microenvironment and the landscape (e.g., Welsh and Lind 2002). Th e following sections de scribe life history characteristics of the indicator and ubiquit ous species that may lead to sensitivity or tolerance to land use. Partic ularly, feeding strategies, micr ohabitat selection, and species specific desiccation tolerances, may render spec ies sensitive or tolerant to changes to wetland landscapes associated with anthropogenic land use intensity. It should be noted that even when pars imonious explanations are available many different factors, including cu mulative affects, may be responsible for the observed trends in species occurrences. Potential explan atory mechanisms for the observed species occurrences and their relationship with land us e intensity beyond the scope of this study include, interspecific competition, diseases, a nd species specific pollution tolerances. Amphibian indicators of reference condi tion and low intensity land uses The highest indicator values of a ny of the amphibians were those for Hyla femoralis The pinewoods treefrog ( Hyla femoralis ) was significantly associated with the sites that had low intensity land uses su rrounding the forested wetlands, and almost exclusively with wetlands that had adjacent pine tree communities. As its common name implies the pinewoods treefrog is associated with pine trees. Perh aps an indication of coevolution with the pineland community, Hyla femoralis has the highest tolerance, out of 10 Florida anurans tested, to low pH conditions, which often occur in wetlands embedded in pineland communities (Warner a nd Dunson 1998). Furthermore, the bright red tails and copper bellies of the pinewood treefrog tadpoles may provide camouflage in the extremely tannin rich waters they frequently are found in (Altig 1972).

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131 The pinewoods treefrog response to land use intensity is more complex then simply a dependence on pine trees. For example, it wa s not recorded at 26 ur ban and agricultural sites that had pine tree communities in some portion of the 200 meter buffer. Changes to wetland water chemistry associated with development may subsequently influence Hyla femoralis populations. Of the six amphibian species selected as indicators of low intensity land uses only two are considered independent of epheme ral wetlands and capable of persisting in wetlands that contai n predatory fish, Acris gryllus and Rana grylio It is somewhat surprising that Rana grylio was found at more then 50% of the reference sites given their larvae take approximately one year to mature, and the reference set contained primarily ephe meral wetlands. The majority of R grylio’s diet consists of crayfish, almost to the poin t of being dietary specialists (Lamb 1984). Crayfish were abundant and more frequently sampled in reference wetlands, and were encountered significantly less at developed site s. During the dip net sweeps (i.e., sites with standing water) crayfish were reco rded at 86%, 17%, 50%, and 29% of the reference, silviculture, agricu lture, and urban sites respective ly, and in turn, the frequency of occurrence of Rana grylio was 52%, 17%, 21%, and 6%, respectively. This suggests that if not a direct dietary link between Rana grylio and crayfish then other wetland conditions (e.g., hydroperiod) c onducive to crayfish persistence are also positively associated with the presence of Rana grylio Alterations to the wetland edge of fo rested wetlands may adversely affect Acris gryllus Acris gryllus was selected as a significant i ndicator of reference condition, and frequency of occurrence decreased with incr easing land use intensity. The eggs of Acris

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132 gryllus are laid in “very shallow water” (Ca rr 1940). Similarly, tadpoles were only found in very shallow water at the extreme wetla nd edge, and adults were typically if not exclusively heard or seen resi ding within a few meters of the waters edge. The shallow water, high insolence, forb/graminoid do minated, ecotone often found around dome swamps was favored by Acris gryllus and this area in particular is frequently altered in developed landscapes (Kirkman et al. 1998). A host of anthropogenic activities appeared to have truncated many of the sample wetla nd edges (e.g., berms, road beds, plowing, mowing, water control systems etc.), perhap s reducing the preferred microhabitat of Acris gryllus in developed landscapes. Another reference indicator sp ecies possibly influenced by the characteristics of the wetland ecotone is the little grass frog ( Pseudacris ocularis ). The adults are typically found within thick wetland graminoid groundc over (Carr 1940, Bartlett and Bartlett 1999). Similarly, the habitat use of little gr ass frog tadpoles is correlated with emergent herbaceous cover, as well as, water temperat ure, and position within the wetland (Alford 1986). Experiments have shown increased vegetation based structural complexity increases tadpole survival due to decreased predation rates (Babbi tt 1996). Land uses that promote shading, frequent mowing, inte nse grazing, or removal, of wetland and the adjacent upland groundcover may negatively affect little grass frogs. The oak toad ( Bufo quercicus ) has been reported to be sensitive to land use changes (Wilson and Porras 1983, Delis et al. 1995, Mushin sky et al. 2001). Fe w explanations are given as to why oak toads are sensitive to la nd use intensity, other then an affinity for ephemeral wetlands (e.g., Mushinsky et al. 2001) It shares many characteristics with the more tolerant congeneric B. terrestris Some significant differences are the oak toads

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133 small adult size (25-30mm), diurnal activit y, small terrestrial home range, shorter breeding season, and an ant-dominated di et (Hamilton 1955, Punzo 1995). Development may strongly influence composition and abundan ce of terrestrial arthropod prey of oak toads. More specifically, imported fire ants ( Solenopis invicta ) increase in disturbed areas and decrease the numbers of native ants and possibly other ground dwelling organisms (Stiles and Jones 1998, Zettler et al. 2004). However, the relationship between the introduced fire ant and anuran s is poorly understood (Dodd 1997). Perhaps the most hydrologically sp ecific species encountered is Hyla gratiosa Breeding wetlands selected by Hyla gratiosa are typically free of fish, but at the same time, relatively deep (e.g., Neill 1958, Moler and Franz 1987). The mean maximum water depth for wetlands that contained H. gratiosa were 20cm deeper then all wetlands combined (58cm and 38cm respectively). Al terations to depressional wetlands and wetland complexes that influence hydrology an d fish occupation may strongly influence H. gratiosa populations. Furthermore, H. gratiosa and several of the other anuran species sensitive to land use intensity are fo ssorial and spend considerable portions of their lives in terrestrial habitats (e.g., Carr 1940, Neill 1952). Many of these species may be responding to alteration of upland habita ts, specifically, incr eases in impervious surface and soil compaction, removal of na tive vegetation, and loss of microhabitat refugia. Notophthalmus viridescens was not selected as an i ndicator of land use in this study. However, N. viridescens was not found in landscapes with less than 46% forest cover in the Northeastern United States (G ibbs 1998, Guerry and Hunter 2002), and is generally considered sensitive to land use in tensity (e.g., Lehtinen et al. 1999). Although

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134 N. viridescens was only found at 6% of the sites in th is study four of the seven sites were in highly developed agricultu ral (n=2) or urban landscapes (n=2), of which, none had greater then 45% forest cove r within 200m of the sample wetland and the average forest cover within 200m for th e four sites was 19%. N. viridescens louisianensis and N. viridescens piaropicola may be less susceptible to habitat loss in the Southeastern United States than N. viridescens viridescens in the northeast. Amphibian indicators of high intens ity land uses and ubiquitous species Persistence of amphibian species in la ndscapes of high human land use intensity may be strongly influenced by pollution, nutrient, and water quality to lerances (Berrill et al. 1995, Hecnar 1995, Houlahan and Findlay 20 03), dehydration tolerance, ability to utilize or disperse through human modified landscapes, util ization of a wide range of natural habitats, ability to breed throughout the year, and availability of prey. Determining which mechanisms were most influential in forming the observed species composition trends in this study is not possibl e. However, several lines of evidence suggest that species that were indicative of high intensity la nd uses have characteristics that aid in coping with wetland drying in developed landscapes. A second group of species is capable of producing offspring in more permanent waters that contain fish. The amphibian species indicative of highe r intensity land uses (agriculture and urban) were generally not found exclusively, but were found more frequently, at sites within developed landscapes. Urban sites ha d the highest proportion of sites that were dry. Several of the species indicative of de veloped sites may have greater tolerance to desiccation (Thorson and Svihla 1943) or able to withstand consecu tive years without water in the breeding wetlands (e.g., Bufo terrestris and Gastrophryne carolinensis )(Dodd 1994, 1995). Wetlands may also dry faster in developed landscapes

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135 and some of the indicator species have rapid maturation rates (e.g., Bufo terrestris ). Perhaps more evidence that wetland hydr ology may be an influencing mechanism shaping amphibian communities, Eleutherodactylus planirostris another species indicative of the more intensive land uses, is th e only species studied that does not require wetlands for breeding. Finally, in developed landscapes, Hyla squirella H. cinerea and Osteopilus septentrionalis are frequently found on buildings and other human structures (Goin 1958, Meshanka 2001). These structures may provide shelter from desiccation as well as predators and temperat ure extremes (Meshaka 1996). At the same time, many of the amphibian species indicative of low intensity landscapes are also capable of withstanding dry periods. However, some species may require specific refugia (e.g., grass clumps downed logs, loose sand, natural openings) (e.g., Neill 1952, Neill 1958, Stebbins and Cohen 1995, Jansen et al. 2001) that may not be available in sufficient quant ities in upland lands capes of the most developed sites. Many of the indicator species of reference sites are common in habitats that frequently burn (Means et al. 2004). Fire s decrease shrub density, increases the amount ground insolence, and increase edaphic evap orative water loss relative to unburned landscapes. Several of the species either in dicative of, or found more frequently in, developed landscapes may partially be res ponding to fire suppression. For example, Meshanka and Layne (2004) reported that Eleutherodactylus planirostris is particularly abundant in long-unburned sandhills with fu ll tree canopies, and Schurbon and Fauth (2003) reported the greatest numbers of Hyla chrysoscelis and H. cinerea at temporary wetlands that had the lowe st fire frequency.

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136 A second group of amphibians tolerant of anthropogenic la nd use are less adapted to dry environments and are generally mo re aquatic. The highest frequency of occurrence of Rana catesbeiana R. sphenocephala R. clamitans and Hyla chrysoscelis was at sites in developed lands capes, although none we re selected as an indicator several of these species have been identified as being more abundant in developed landscapes (Zampella and Bunnell 2000, Houlahan a nd Findlay 2003). These species and Hyla cinerea share a tolerance to fish presence and may have benefited from the many manmade water bodies (e.g., canals, cattl e ponds, retention ponds) in developed landscapes. Furthermore, many of the prim arily aquatic species aestivate and/or hibernate (e.g., Ranids)(Knutson et al. 1999) in or near water and are generally less dependent on upland habitats for survival. Finally, the increased pres ence of the batracophagous, Rana catesbeiana and Osteopilus septentrionalis in developed landscapes has th e ability to further influence sympatric anuran species populations (e.g., Adams 1999, Zampella and Bunnell 2000, Smith 2005). Additionally, a refere nce indicator species larva (e.g., Hyla gratiosa ) was more abundant in experimental wetlands that had greater numbers of caudate predators and when caudate predators were reduced se veral of the ubiquitous and developed land indicator species in this study (e.g., Rana sphenocephala Bufo terrestris Hyla chrysoscelis ) were competitively superior (Mor in 1983). This suggests a complex relationship between amphibian predators, co mpetitors, and land use that has the ability to further influence species assemblages. Avian Indicator Species More avian species were indicative of land use type and intensity than amphibians. There were eight, three, elev en, and seven species that had the highest

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137 significant (p>0.1) indicator values at referen ce, silviculture, agricu lture, and urban sites respectively. More species were selected as indicators of a speci fic land use type than were selected as significant indicators of LDI quartiles, 29 and 24 respectively. The values were greater for the more intens e land uses when grouped by land use type (agriculture and urban) than by LDI quartile (3 and 4). Howe ver, indicator values were greater at the LDI quartile 1 sites than the reference sites. A group of birds seems to sensitive to even relatively small amount s of anthropogenic land use intensity. Four of the species Woolfenden and Rohw er (1969) listed as sensitive to urban development in Florida, Bachman’s sparrow, pine warbler, brow n-headed nuthatch, and red-cockaded woodpecker, were selected as i ndicators of referen ce condition in this study. Bachman’s sparrow and common yellowthr oat had the highest in dicator values of the reference species. However, Bachman’s sparrow was rarely observed utilizing the forested wetland but was most frequently observed in the bu rned (approximately 3-5 yrs) pine savannah habitats adjacent to the re ference wetlands. The common yellowthroat and rufous-sided towhee were found within the fo rested wetlands as well as the adjacent reference habitat. However, both species were strongly associated with dense shrub cover and palmettos ( Serenoa repens ), which were also common in silvicultural landscapes. The prothonotary warbler was th e only species that might be considered dependent on forested wetlands selected as an indicator of reference sites. The relatively low indicator value for the prothonotary warbler was a result of it being detected at only four reference sites in the panhandle. Only two species had the highest indicat or values for silvicultural landscapes, rufous-sided towhee and American swallow tailed kite. However the swallow tailed kite

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138 was a weak indicator (silvicultu re Indicator value = 12, p=0.09) of silvicultural land uses given it was recorded at five sites, and onl y two of which were typed silviculture. The remaining sites were low intensity agricultural sites (n=2), or consisted of a mix of urban, agriculture, and reference ha bitats (n=1). The agricultural sites had the most avian species selected as indicators. However, few of the species were associated with the sample wetlands, but were more frequently detected in the open (i.e., nonforested) or sh rubby edge habitats w ithin the ag ricultural sites. Of the 11 agricultu ral indicators only the red shouldered hawk, red-winged blackbird, wild turkey and white ibis frequently utilize we tlands (FFWC 2003). By far the most common agricultural indicator was th e cattle egret. Although cattle egrets will nest in isolated wetlands, no nesting colonies were encount ered, and all observations occurred in the circumjacent habitats. As their common name implies they were frequently observed in association with ca ttle but were also frequently observed in adjacent hay and fallow fields. Hirth and Marion (1979) also found the highest frequency of occurrence of several of the same avian species in Florida agricultural lands capes that were selected as indicators of agriculture in this study, cattle egrets and easte rn meadowlarks in pastures, red-winged blackbird in row crops, and turk ey and black vultures in grazed forest. The highest indicator value for any of the avian species was the northern mockingbird at urban sites. Like many of the other avian indicators the mockingbird was not strictly associated with the forested we tland but more so with the adjacent land use. Only the common grackle and blue jay of the urban indicators and the boat-tailed grackle of the LDI category four sites were consistently observed ut ilizing the sample wetlands.

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139 The most common species Woolfenden and Rohwer (1969) found in Florida suburbs that were also indi cators of urban land use in this study included, northern mockingbird, house sparrow, and blue jay. A dditionally, blue jay (Smith and Schaeffer 1992, Kluza et al. 2000), northern mocki ng bird (Mills et al. 1989, Rottenborn 1999, Blair 2001), house sparrow (M ills et al. 1989, Clergeau et al. 1998, Rottenborn 1999, Blair 2001, Hostetler and Knowles-Yan ez 2003), and house finch (Rottenborn 1999, Blair 2001) have all been found to be indicativ e or most abundant in urban environments outside of Florida. Species Guilds and Land use The presence of a species is partially de termined by specific characteristics of the habitats within a landscape. In turn, inhere nt behavioral and ecologi cal attributes largely predispose the response of a species to as pects of a landscape. Grouping species by predominate life history characteristics reve aled trends in comm unity composition that suggest a functional response to habitat attributes. The fo llowing sections discuss the trends in amphibian and avian species gu ilds with changes in land use and land use intensity. Amphibian Guilds The amphibians encountered in this study have many overlapping life history strategies (e.g., all ad ults are predominately invertivores ) making it difficult to separate species into guilds. However, changes in the amphibian guilds, ephemeral wetland dependency and egg deposition strategy, were re lated to changes in land use intensity. The mean proportion of the amphibian community that was dependent on ephemeral wetlands declined from reference to urban sites. Three possible explanations for this trend are, urban wetlands had more permanent wetland waters, there were more

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140 fish in urban wetlands, and the amphibian species that are dependent on ephemeral wetlands for breeding are also the species that are incompatible with urbanized uplands. The hydroperiod (number of days inundated) of the urban wetlands appeared to be altered by two disparate mechanisms. H ydroperiod increased when the wetland was incorporated into a stormwater system or ge nerally decreased when adjacent to retention ponds and canals. Unfortunately the specific hyd roperiod of the wetlands in this study is largely unknown. The presence of fish, whic h are the reported driving mechanisms behind some species dependence on ephemeral wetlands, was not significantly higher at urban sites. Thus it is difficult to determ ine if the proportion of amphibians dependent on ephemeral wetlands for breeding decreases in urban areas because of their breeding strategy or a covariate. Azous and Horner (1997) suggested eggs at tached to vegetation or other substrates may be unable to survive in flashy urban wetla nds. In this study significantly greater proportion of the amphibians that have subm erged eggs attached to vegetation were present at reference and silvicultural wetlands However, amphibian species that attach eggs were not only proportionally lower at urban sites but also at the agricultural wetlands. Additionally, the egg stage for most amphibian s encountered in this study lasts only 1-6 days (Duelman and Trueb 1986, Stebbins an d Cohen 1995). Few if any of the sample wetland water levels appeared to drastically fluctuate over a 1-6 da y period. However, significantly higher ammonia levels and signifi cantly lower dissolved oxygen levels were recorded at urban and agricultural sites. Oxygen, carbon dioxide, and ammonia are passively transported across the membrane of amphibian ova (Stebbins an Cohen 1995).

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141 Egg development is dependent on oxygenation, and floating eggs are believed to be an adaptation for increasing oxygen availabil ity to developing embryos (Moore 1940). Species that have submerged or submerged attached eggs versus species that have floating eggs may receive less oxygen in wetla nd waters of developed landscapes. Furthermore, species specific embryo and larval sensitivity to environmentally relevant concentrations of ammonium nitr ate has been observed in several amphibians (Oritz et al. 2004). Extremely low dissolved oxygen levels, which often occur during the predawn hours in wetlands, in combination with a rich source of ammonium could potentially produce high levels of nitrite, a nd methemoglobninemia has been suggested as a mechanism afflicting tadpoles showing reduc ed activity (Hecnar 1995 ) and may also be a factor in reducing the proportion of species that have submerged eggs. Avian Guilds The regional species pool of birds exhibits a wide array of feeding, breeding, and migratory strategies, and signi ficant differences in the pr oportion of species utilizing different strategies were observed between la nd use types and land use intensity quartiles. To summarize the significant trends with increasing land use inte nsity, insectivorous birds decreased (e.g., Stratford and Stouffe r 1999, Ford et al. 2001, White et al. 2004), omnivorous and herbivorous (granivores and frugivores) birds increased (e.g., Smith and Schaefer 1992, Clergeau 1998, McKinney 2002) bark and canopy gleaners decreased (e.g., Rottenborn 1999), ground gleaners increase d, territorial species decreased (Emlen 1974, Mills et al. 1989), ground and cavity nesters decrease d (e.g., McComb et al. 1986, Cox 1987, Kluza et al. 2000, Rodewald and Yahne r 2001, Lohr et al. 2002, Hausner et al. 2003), canopy nesters increased, ex otic species increased (Bla ir 1996), species that have demonstrated population declines in Flor ida decreased, and species that have

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142 demonstrated population increases in Flor ida increased, with increasing land use intensity. Insectivores have been identified as bei ng particularly sensitive to anthropogenic land uses (e.g., Smith and Schaever 1992, Canaday 1996, Galatowitsch et al. 1999, Stratford and Stouffer 1999, Ford et al. 2001, Wh ite et al. 2004). Despite significantly more insect species, counts, and biomass sa mpled at the agricultu ral sites the proportion of the avian community that was insectivorous was significantly less then at reference sites. There were no significant differences in insect measures be tween LDI quartiles. A reduction in the vegetation structural co mplexity with increasing human land use intensity is often thought to be responsible for the declin e in insectivorous birds (e.g., Miller et al. 2003). The majority of the insectivorous species in this study maintain intraspecific feeding territories, and Mill s et al. (1989) reported decrea ses in avian species that maintain feeding territories with decreases in the volume of native vegetation. The habitat is often divided among insectivorous species based on species specific feeding strategies that often relate to specific regions of th e vegetation complex (e.g., bark gleaner, canopy forager etc.)( De Graaf 1985). The continuity of these vegetation microhabitats or strata can be severely disr upted in developed landscapes, as is forest habitat in general. Thus insectivorous avian species because they maintain feeding territories may be more susceptible to reducti ons in both, forest microhabitats and forest continuity, which commonly occu r in developed landscapes. Omnivorous and herbivorous avian species on the other hand are less likely to maintain feeding territories and may be more capable of exploiting patchy resources in

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143 developed landscapes. Furthermore, species th at utilize vegetation as a food source may require less land area of a specific habitat type It has been estimated that the minimum amount of habitat area needed to maintain a representativ e community of seed eating birds is 2 ha and to maintain an insecti vorous community is 40 ha (Forman et al. 1976, Galli et al. 1976, Environmental Law Institute 2003). Omnivores and herbivores may find a dditional food resources in developed landscapes (e.g., Lepczyk et al. 2004). Granivores may find more annual, wind dispersed, seed bearing plants in habitats that experience frequent vegetation and soil disturbance. Likewise, frugi vorous species may benefit from the greater abundance of exotic and ruderal fruit beari ng plants in developed landscape s, which in Florida are also more likely to have ripe fruit during the br eeding season than fruit bearing species in undisturbed habitats (Skeate 1987 ). Omnivores and herbivores may also benefit from the food crops, refuse, and feeding statio ns common in developed landscapes. The relationship between fragmentation, te rritoriality and disturbance can perhaps be seen with the finding of si gnificantly larger avian species occupying agricultural sites. Species that are large generally have large territories or home ranges and may be more adept at using multiple habitat patches (Hostetler and Holling 2000). The sample wetlands and the circumjacent landscape in th is study may only be a component of a larger area of utilization for species with larg e home ranges. In agricultural landscapes the spatial and temporal frequency of dist urbance is highly vari able. Additionally, agricultural lands generally ar e productive and seve ral large species (i.e., wading birds) are found more frequently in enriched wate rs (Crozier and Gawlik 2002). Thus large

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144 avian species in this study are capable of utilizing the, patchy but highly productive, resources of agri cultural landscapes. The proportion of avian species that fora ge in the forest canopy decreased and ground foragers increased with increasing la nd use intensity wher eas the proportion of the avian community that nested in the canopy increased and gr ound nesters decreased with increasing land use intensity. These tr ends demonstrate that habitat structure associated with land use intensity may have an opposite associati on with separate life history characteristics, and is likely the result of an interaction between the available habitat structure and associated food resour ces, and predation rates along the land use intensity gradient. Although many studies have indicated Neot ropical migrant birds are sensitive to increasing land use intensity (e.g., Kluza et al. 2000, McKinney 2002, Norris et al. 2003) this study failed to find a strong associat ion between the proportion of the avian community comprised of Neotropical migran ts and land use. A relatively small proportion of the avian community in Florida consists of Neotropical migrants making it difficult to find a significant association. At the same time, most Neotropical migrants are territorial insectivores suggesting one mechanism for the observed decreases with land use intensity in similar studies. The proportion of the avian community that is experiencing si gnificant population decreases in Florida was significantly hi gher, and species experiencing population increases significantly lower, at the lowest land use intensity sites. From a conservation perspective species experienci ng population decreases warran t attention and this study indicates an association between land use intensity and avian population trends. The

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145 relationship is slightly str onger when the land use intensity is quantified using the LDI index versus grouping the sites by land use type At the same time, approximately 28% of the avian species encountered at the ag ricultural and urban s ites are experiencing significant population de creases in Florida. This sugge sts isolated forested wetlands within developed landscapes provide habitat for declining avian sp ecies. Landscape Development Intensity (L DI) Index and Species Composition There are corresponding gr adients between land use intensity and avian and amphibian species composition. Comparisons between LDI and at least one of the NMS axes resulted in some of the strongest univari ate correlation coefficients out of more than 30 measured independent variables identified in the literature as potentially influencing avian and amphibian assemblages (Appendices B and C). The significance of these findings is the LDI index can be remotely es timated where as many of the other measures related to habitat quality re quire more intensive and time consuming efforts (e.g., site visits). Studies of land use intensity have st ructured the land use in tensity gradient using primarily subjective means and are often la nd use type specific (e.g., urban intensity, grazing intensity etc.). The LDI index is capable of incorporating land use intensity estimates from multiple land use types, and has the ability to standardize land use intensity estimates between land us e intensity gradient studies. The land use intensity gradient in ma ny studies is implied (e.g., Canaday 1996, Maestas et al. 2003), however, there are a fe w exceptions (e.g., Blai r 2001, Houlahan and Findlay 2003, Miller et al. 2003). Blair (2001) rated land uses using a Delphi Technique, which resulted in placing land uses along a gradient that corr esponds with the less subjective LDI coefficients (Brown and Vi vas 2004). There were many similarities between the results of Blair’s (1996) land use gradient study and this study. For example

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146 both studies identified species whose occurr ence decreases in landscapes with increasing land use intensity, and several of the same avian species indi cative of high intensity land use were common to both Ca lifornia and Florida. Miller et al. (2003) created an index of urbanization using principal components analysis that was strongly correlated to building density, –0.93 to -0.89, within 100 and 1500m respectively. However, this index is sp ecific to their study a nd could not be used for intensity estimates within agricultural and silvicultural lands capes. Houlahan and Findlay (2003) measured building density su rrounding wetlands and fo r agricultural areas used public data on estimates of chemical a pplication and the propor tion of cropland. However, these disparate measures, building density and estimated chemical application, of land use intensity cannot intuitively be used in a land use gradient analysis. There are several trends between increasing land use intensity and wildlife habitat. Generally the amount of impervious surface, bu ilding density, automobile traffic, exotic biota, chemical application, human presence and input of energy and matter increase with increasing land use intens ity, while the amount of fore sted or undeveloped habitat, amount of native vegetation decrease (e.g., Blai r 1996, Maestas et al. 2003, Miller et al. 2003). All of these factors have the po tential to individua lly, but more likely synergistically, affect verteb rate biodiversity, and the LD I index is one of the few methods capable of quantifying the intens ity of multiple land use types. Variability in LDI index There are some caveats to the LDI index in relation to wildlife species response to land use intensity. The strength of the re lationship between wildlife habitat and the energies accounted for in the index may be hi ghly variable. For example, although large amounts of energy are needed to use electricity, the effect of electricity utilization on

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147 wildlife habitat is presumably less then the effect of the less energy intensive act of clearing vegetation. Second, there is the poten tial for an acute response by some species to specific aspects of a landscape. For inst ance, one type of biocide may have a minimal effect on a species while another biocide coul d have a profound affect. Finally, the LDI index is calculated using the proportion of land of varying intensity surrounding the study site. A potential problem arises when a la rge proportion of the landscape surrounding a site is undeveloped and the remaining proporti on is a relatively high intensity land use resulting in a site with a middle range LDI score. For comparison, another site completely surrounded by a lower intensity la nd use could have an identical LDI site score. The undeveloped portion in the fo rmer example may be more conducive to wildlife movement and ultimately the likelihood of species persistence increases whereas in the habitat fragment completely surr ounded by anthropogenic land uses, habitat connectively may be reduced and some species persistence less likel y (deMaynadier and Hunter 1999, Ricketts 2001). Despite the inherent variability in wild life stressors and habi tat connectivity in landscapes the LDI index represented a large proportion of the differences in amphibian and avian species composition be tween sites in this study. Variability in the Landscape There are social, biological, temporal, and historical factors specific to each site that have the ability to influence the obs erved relationship between measured wildlife composition and a land use intensity estimate. Direct anthropogenic effects, time lags in species response to disturbance, land use history, and regional species pools all have the ability to create variabilit y in the relationship between land use intensity and species composition.

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148 There are a host of potential direct anthropogenic affects within a landscape that would not be captured by any remotely base d land use intensity estimate. The illegal dumping of toxins, release pe st species or exotic pets, and hunting are just a few examples. At the same time, landowner or land user individuality within the landscape is inherently variable and some actions will have a disproportionate effect on wildlife composition relative to the LDI coefficient for that land use (e.g., Best Management Practices, intensity of la ndscaping, etc.). Especially when presence or absence is concerned there may be substantial time lags in the response of a wildlife species to changes in land use intensity. Tilman et al. (1994) used the term “Extinction Debt” when the biotic communities ultimate response to a stressor is manifested over ti me and the true number of sp ecies that will be unable to persist after a habitat is mani pulated is not immediately evident. Likewise, the response of potential colonizers to resources made available by human influence to landscapes may also take time. Findlay and Bourdage s (2000) greatly increased the predictive abilities of their model of species richness and road density when the amount of time the roads were present was included. The length of time each land use was present surrounding the study sites was not incorporated into this study. An attempt was made to account for only sites that have been develope d for more than 30 years by using the land use data from 1974 however the coarseness of the land use delineation of the coverage may have inhibited the detection of a tem poral affect between species composition and land use intensity. Several of the most rece ntly developed urban si tes despite relatively high LDI scores maintained wildlife species characteristic of the sites with lower LDI scores.

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149 Land use history for each site was largel y unknown and there are several potential land use trajectories for each site. At some poi nt in the past a reference site may have formerly had homesteads, been grazed or ha d cropland, etc. Similarly a predominately urban site may have first been an agricultural site or alternatively b een developed directly from a reference habitat. All of these scen arios have the potential to influence what species were detected at the site in the current study. Conclusion Shifts in amphibian and avian spec ies composition of forested wetlands corresponded with land use intensity, and ther e was a gradient of in creasing dissimilarity from reference to silviculture, agricult ure, and urban land uses respectively. Furthermore, species sensitive and tolerant to anthropogenic land use were identified. This information may aid in habitat asse ssments, land use planning, wetland creation and restoration, and provide an ear ly warning for which species may be sensitive to future anthropogenic habitat manipulations. Biological integrity of isolated forested wetlands is dependent on the circumjacent terrestrial habitat and the dynamic ecologi cal processes that shape both. Wetland hydroperiod, water quality, regional wetland complexes, fire frequency, trophic interactions, vegetation structure and com position are all temporally and spatially variable and cumulatively affect biotic comm unities of isolated wetland systems. The number of processes affected by humans and the degree to which habitat alteration occurs strongly corresponds with anthr opogenic land use intensity. That being said, isolated wetlands in even the most developed landscapes perform many ecosystem functions including wildlife hab itat. Despite signifi cant differences in species composition, isolated wetlands in developed landscapes support many species

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150 and provide critical habitat on a regional basis. For example the only yellow crowned night heron rookery was found in a sample wetland within a highly developed urban landscape. Similarly, severa l declining migratory songbird s were observed utilizing urban sample wetlands and these habitats may provide crucial stopover habitat along the highly developed coasts. The exponentially growing human populati on in Florida demands continued land use intensification. Remnant forested wetla nds embedded in a human dominated matrix provide an opportunity to interface humans with nature as well as providing ecosystem services. However, the biota characteristic of Florida be fore large scale anthropogenic land subjugation will only be found in areas th at maintain vital eco logical processes and receive minimal human land use intensity.

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151 APPENDIX A SITE SURVEY LENGTHS Site Name both surveys (min) cassette tape (min) surveys and tape (min) AL43RD 50 48 98 ALACF2 73 48 121 ALBALU 66 48 114 ALBEEF 75 0 75 ALBUCK 90 41 131 ALHAG1 99 47 146 ALHAG2 68 48 116 ALJOEL 60 48 108 ALMORN 79 48 127 BACUT 84 46 130 BAFED 67 44 111 BAMEX 45 48 93 BAONF1 83 17 100 BAPINE 135 48 183 BASID 62 28 90 BRSSRR 111 46 157 CIMILL 177 14 191 CLBLAN 121 46 167 CLGUS 156 48 204 CLJEN2 136 48 184 CLJENN 135 47 182 COBODY 104 20 124 COBOGL 75 47 122 COCLAR 50 0 50 COFEDX 65 48 113 COHART 94 22 116 COMATO 150 48 198 COPRIS 70 45 115 COROB 94 24 118 DICALI 80 48 128 DUFROG 70 48 118 DUJACK 70 48 118 DUJAME 55 47 102 DUPUMP 100 48 148 ESGOD 122 44 166 ESPEN 135 43 178

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152 Site Name both surveys (min) cassette tape (min) surveys and tape (min) FLCANA 105 46 151 FLCSR 107 48 155 FRANF1 71 47 118 GIADDY 65 48 113 GLROAD 79 0 79 HAHUNT 127 44 171 HECHAS 172 49 221 HEDIDD 55 0 55 HEHIKE 55 0 55 HIBENI 124 15 139 HICLUB 113 15 128 HIFLAT 102 40 142 HILAND 124 0 124 HIPALM 93 41 134 HOBON 73 0 73 HOBON2 57 13 70 HOHOOD 78 49 127 IRSSRI 104 48 152 JEPICK 130 50 180 LE93 108 46 154 LEBIG 168 41 209 LECALU 105 43 148 LECOWS 105 0 105 LEDAWG 55 0 55 LEDGE 118 32 150 LEDONE 50 48 98 LELOST 90 22 112 LEMESS 111 39 150 LEROLL 121 46 167 LIANF2 102 0 102 MADEEP 93 47 140 MAPATA 138 21 159 OKAIR 193 46 239 OKJEB 129 38 167 OREOD 71 46 117 ORIFOR 115 0 115 ORUCF 116 43 159 ORVILA 106 46 152 ORXMAS 125 45 170 OSBILY 162 42 204 OSCRES 123 39 162 OSMIKY 87 44 131 OSMINI 78 42 120

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153 Site Name both surveys (min) cassette tape (min) surveys and tape (min) OSTNC1 161 0 161 PABULL 57 48 105 PBCAMP 117 0 117 PBGARL 79 46 125 PBHIGH 85 48 133 PBLOXR 86 0 86 PBROLL 89 47 136 PBSNAP 81 46 127 PBSTAR 81 40 121 PBTHAI 37 0 37 PBWOOD 103 29 132 PBZEP 94 47 141 POCOWS 187 45 232 POEGGS 60 48 108 POFECT 114 0 114 PUPLUM 133 34 167 PURICE 84 8 92 SAMYK2 91 0 91 SERENE 118 49 167 SJJULI 133 38 171 SLBOGR 115 41 156 SLLIME 83 36 119 SRGAR 118 46 164 SRGENE 113 0 113 SRJAY 143 44 187 SUMMER 125 30 155 SUWITH 93 0 93 WAGLOV 147 46 193 WAHELL 46 47 93 WAMARK 149 47 196 WASIV 51 0 51 WAWASH 101 47 148

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154 APPENDIX B CORRELATES OF AMPHIBIA N SPECIES COMPOSITION Correlations between site environmental vari ables and NMS ordination axes scores based on amphibian composition (Pearson correlation coefficients). Independent variables axis 1 axis 2 axis 3 count 2003 LDI 0.03 0.03 -0.68** 111 WRAP 0.06 0.00 -0.74** 109 FWCI -0.01 0.12 0.77** 109 1974 LDI 0.19 -0.10 -0.48** 111 mean LDI (2003 and 1974) 0.12 -0.04 -0.66** 111 % agriculture (200 m) -0.02 -0.12 -0.28** 111 % undeveloped (200 m) -0.06 -0.06 0.55** 111 % silviculture (200 m) 0.10 0.10 0.31** 111 % undeveloped + silviculture (200 m) 0.00 0.01 0.72** 111 % urban (200 m) 0.02 0.11 -0.50** 111 % highway/road (200 m) 0.09 0.11 -0.24* 111 % canal/ditch (200 m) -0.27** -0.10 -0.30** 111 % wetland (200 m) -0.01 -0.09 0.23* 111 nearest wetland (distance) 0.10 -0.04 -0.12 111 2nd wetland (distance) 0.20* 0.01 -0.31** 111 continuous forest area (ha) (200 m) 0.01 -0.01 0.63** 111 distance to forest cover (m) 0.05 -0.07 -0.29** 111 distance to major river (km) -0.19 -0.03 -0.15 111 paved road (distance) -0.04 -0.04 0.33** 111 focal wetland area (ha) 0.21* 0.06 -0.08 111 total area (200m buffer) 0.13 0.07 -0.11 111 fire history in dome 0.00 0.10 0.54** 111 fire history in landscape -0.05 0.01 0.69** 111 latitude 0.36** 0.16 0.19 111 longitude -0.25* 0.05 -0.01 111 Julian date 0.41** -0.10 0.08 111 both surveys (min) 0.32** -0.15 0.24* 111 surveys and tape (min) 0.28** -0.08 0.21* 111 water dissolved oxygen -0.06 -0.07 0.35** 87 water temperature 0.01 -0.01 0.06 87 water ammonia -0.17 0.04 -0.30** 87 water NO2NO3-N 0.20 0.09 -0.12 87 water TKN -0.16 0.06 -0.32** 87

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155 Independent variables axis 1 axis 2 axis 3 count water total phosphorus -0.12 0.05 -0.32** 87 water turbidity -0.11 0.12 -0.18 87 water color -0.11 0.21 -0.02 87 water pH 0.08 -0.15 -0.72** 87 water conductivity -0.08 0.10 -0.47** 87 maximum water depth (cm) 0.52** -0.25** 0.28** 111 wetland wet or dry (1/0) 0.51** -0.22* 0.26** 111 fish species richness (sweeps) -0.05 0.05 0.02 90 predatory Fish or not (1/0) 0.02 -0.03 -0.08 111 predatory fish (gambusia) or not (1/0) 0.24* -0.01 0.04 111 predatory fish excl uding dry sites (1/0) -0.09 0.01 -0.15 88 predatory fish (gamusia) excluding dry sites (1/0) 0.06 0.11 -0.07 88 amphibian species richness 0.13 -0.37** 0.31** 111 insect species richness -0.11 0.14 -0.22* 82 insect individual count -0.02 -0.07 -0.22* 82 insect shannon diversity -0.09 0.17 -0.15 82 insect biomass (g) 0.04 -0.12 -0.16 82 plant species richness 0.05 0.04 -0.13 109 basal area m2/ha -0.14 0.13 -0.25* 108 % canopy cover 0.01 0.18 -0.12 63 % evergreen (BA) -0.15 -0.16 0.16 108 % broad leaved (BA) 0.12 0.13 -0.07 108 % exotic trees (BA) -0.12 -0.18 -0.21* 108 % exotic plants -0.09 -0.19 -0.67** 109 % native perennial 0.06 0.19 0.64** 109 % sensitive plants -0.08 -0.04 0.78** 109 % tolerant plants 0.11 -0.03 -0.72** 109 % wetland plants 0.06 0.18 0.57** 109 air temp (mean) -0.04 0.20 -0.20* 111 = p<0.05, ** = p<0.01

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156 APPENDIX C CORRELATES OF AVIAN SPECIES COMPOSITION Correlations between site environmental vari ables and NMS ordination axes scores based on avian composition (Pearson correlation coefficient). Independent variables axis 1 axis 2 axis 3 Count 2003 LDI -0.74** -0.04 0.48** 111 WRAP -0.70** -0.21* 0.49** 109 FWCI 0.57** 0.34** -0.53** 109 1974 LDI -0.46** -0.15 0.33** 111 mean LDI (2003 and 1974) -0.68** -0.10 0.46** 111 % agriculture (200 m) -0.09 -0.51** 0.21* 111 % undeveloped (200 m) 0.58** 0.22* -0.35** 111 % silviculture (200 m) 0.21* 0.07 -0.20* 111 % undeveloped + silviculture (200 m) 0.68** 0.26** -0.46** 111 % urban (200 m) -0.65** 0.20* 0.30** 111 % highway/road (200 m) -0.29** 0.22* 0.23* 111 % canal/ditch (200 m) -0.43** -0.07 0.14 111 % wetland (200 m) 0.20* 0.02 -0.22* 111 nearest wetland (distance) -0.37** 0.04 0.05 111 2nd wetland (distance) -0.46** 0.12 0.22* 111 continuous forest area (ha) (200 m) 0.69** 0.17 -0.43** 111 distance to forest cover (m) -0.43** -0.29** 0.08 111 distance to major river (km) -0.04 0.02 0.24* 111 paved road (distance) 0.44** 0.00 -0.22* 111 focal wetland area (ha) 0.00 -0.06 0.08 111 total area (200m buffer) -0.05 -0.19 0.00 111 fire history in dome 0.38** 0.10 -0.35** 111 fire history in landscape 0.59** 0.17 -0.44** 111 latitude 0.12 0.01 -0.05 111 longitude 0.11 -0.07 0.03 111 Julian date -0.07 -0.14 -0.07 111 both surveys (min) 0.12 -0.25** -0.33** 111 surveys and tape (min) 0.16 -0.33** -0.31** 111 maximum water depth (cm) 0.18 -0.22* -0.29** 111 wetland wet or dry (1/0) 0.07 -0.13 -0.25* 111 fish species richness (sweeps) 0.11 -0.07 -0.05 90

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157 Independent variables axis 1 axis 2 axis 3 Count amphibian species richness 0.24* -0.10 -0.37** 111 insect species richness 0.03 -0.17 0.25* 82 insect individual count -0.05 -0.12 0.12 82 insect shannon diversity 0.00 -0.02 0.19 82 insect biomass (g) -0.04 -0.19 0.20 82 plant species richness -0.05 -0.17 0.07 109 basal area m2/ha -0.04 -0.06 0.19 108 % canopy cover 0.01 0.16 0.31* 63 % evergreen (BA) 0.05 0.26** -0.07 108 % broad leaved (BA) -0.14 0.12 0.08 108 % exotic trees (BA) -0.28** 0.17 0.21 108 % exotic plants -0.50** -0.24* 0.42** 109 % native perennial 0.48** 0.35** -0.40** 109 % sensitive plants 0.53** 0.27** -0.58** 109 % tolerant plants -0.45** -0.40** 0.47** 109 % wetland plants 0.48** 0.17 -0.42** 109 air temp (mean) -0.14 -0.02 0.16 111

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APPENDIX D SITE CHARACTERISTICS

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159 Site name County Region LDI LDI quartile WRAP FWCI Land use Latitude Longitude Julian date Maximum water depth AL43RD Alachua North 3.66 3 3.9 38 Urban 29.73240 -82.38854 154 5 ALACF2 Alachua North 1.59 2 2.1 50 Silviculture 29.76662 -82.20515 150 0 ALBALU Alachua North 1.58 2 3.3 49 Silviculture 29.72141 -82.16000 232 56 ALBEEF Alachua North 3.96 3 6.9 2 Agricultural 29.74165 -82.26923 232 51 ALBUCK Alachua North 6.14 4 6.6 29 Urban 29.72551 -82.35376 232 15 ALHAG1 Alachua North 3.92 3 4.3 40 Agricultural 29.79495 -82.41924 232 13 ALHAG2 Alachua North 4.28 3 5.1 12 Agricultural 29.80023 -82.41417 150 0 ALJOEL Alachua North 4.95 3 5.3 14 Urban 29.67127 -82.32561 151 20 ALMORN Alachua North 1.04 1 1.5 36 Reference 29.66101 -82.27516 150 13 BACUT Bay Panhandle 6.15 4 5.4 33 Urban 30.19108 -85.68608 181 41 BAFED Bay Panhandle 7.14 4 5.6 31 Urban 30.21191 -85.64720 181 64 BAMEX Bay Panhandle 4.27 3 3.9 47 Urban 29.93822 -85.39423 178 20 BAONF1 Baker North 1.66 2 2.9 57 Silviculture 30.30300 -82.41403 197 36 BAPINE Bay Panhandle 1.54 2 1.7 51 Silviculture 30.40339 -85.88346 229 99 BASID Bay Panhandle 5.70 4 5.1 19 Urban 30.19020 -85.77982 181 0 BRSSRR Brevard Central 1.00 1 1.5 43 Reference 27.84150 -80.56868 237 36 CIMILL Citrus North 4.96 3 4.4 25 Urban 28.73751 -82.54058 219 99 CLBLAN Clay North 1.00 1 1.4 53 Reference 30.01672 -82.01836 232 61 CLGUS Clay North 3.86 3 6.3 5 Agricultural 29.94500 -81.72512 208 53 CLJEN2 Clay North 1.02 1 2.5 59 Reference 30.17871 -81.93942 208 61 CLJENN Clay North 1.49 2 1.8 52 Silviculture 30.16982 -81.93639 208 58 COBODY Columbia North 7.00 4 5.9 41 Urban 30.20991 -82.64868 197 56

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160Site name County Region LDI LDI quartile WRAP FWCI Land use Latitude Longitude Julian date Maximum water depth COBOGL Collier South 1.00 1 1.0 52 Reference 25.98619 -81.24261 160 71 COCLAR Lee South 2.58 2 4.9 24 Urban 26.32488 -81.77125 160 0 COFEDX Collier South 5.21 3 5.5 34 Urban 26.31050 -81.78732 161 0 COHART Columbia North 4.47 3 5.4 38 Urban 30.17120 -82.66393 197 38 COMATO Collier South 4.81 3 7.5 12 Agricultural 26.26321 -81.41238 159 30 COPRIS Collier South 1.20 2 1.0 47 Reference 26.10602 -81.34489 158 18 COROB Hendry South 3.57 2 4.6 10 Agricultural 26.65347 -81.49650 202 0 DICALI Dixie North 1.70 2 4.2 40 Silviculture 29.47003 -83.09615 152 20 DUFROG Duval North 6.30 4 4.8 25 Urban 30.20305 -81.76350 163 20 DUJACK Duval North 7.11 4 6.4 21 Urban 30.40546 -81.72289 163 76 DUJAME Duval North 6.99 4 6.2 42 Urban 30.23957 -81.52941 163 5 DUPUMP Duval North 1.03 1 1.2 42 Reference 30.47229 -81.49923 163 71 ESGOD Escambia Panhandle 5.15 3 6.8 13 Agricultural 30.97707 -87.49655 187 56 ESPEN Escambia Panhandle 4.93 3 5.4 18 Urban 30.45497 -87.32888 186 41 FLCANA Flagler Central 4.78 3 5.8 12 Agricultural 29.48414 -81.44358 222 38 FLCSR Flagler Central 5.06 3 5.1 20 Urban 29.47656 -81.26731 222 61 FRANF1 Franklin Panhandle 1.04 1 1.6 58 Reference 29.95488 -84.99321 180 20 GIADDY Gilchrist North 3.97 3 7.5 4 Agricultural 29.59070 -82.72842 153 25 GLROAD Glades South 4.84 3 6.0 12 Urban 26.77605 -81.35576 201 0 HAHUNT Hamilton Panhandle 2.42 2 4.7 31 Agricultural 30.46537 -82.70119 197 8 HECHAS Hernando North 1.07 1 1.1 56 Reference 28.63427 -82.57047 219 71 HEDIDD Hendry South 3.12 2 4.5 25 Agricultural 26.28800 -81.22514 161 23 HEHIKE Hendry South 3.90 3 6.0 3 Agricultural 0.00000 0.00000 201 8 HIBENI Hillsborough Central 7.68 4 6.4 16 Urban 28.07078 -82.50734 194 41

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161Site name County Region LDI LDI quartile WRAP FWCI Land use Latitude Longitude Julian date Maximum water depth HICLUB Hillsborough Central 5.25 4 5.8 31 Urban 27.88246 -82.27471 194 36 HIFLAT Hillsborough Central 1.08 1 1.8 50 Reference 28.11667 -82.34595 194 41 HILAND Hillsborough Central 1.83 2 2.7 29 Reference 28.14343 -82.22698 195 71 HIPALM Hillsborough Central 6.46 4 6.4 23 Urban 28.10109 -82.38816 194 51 HOBON Holmes Panhandle 5.78 4 7.5 5 Urban 30.78597 -85.68024 228 8 HOBON2 Holmes Panhandle 4.89 3 6.4 19 Urban 30.76760 -85.68549 228 0 HOHOOD Holmes Panhandle 3.64 2 6.0 10 Agricultural 30.78991 -85.88736 228 99 IRSSRI Indian River Central 3.38 2 3.4 37 Agricultural 27.81016 -80.53548 237 41 JEPICK Jefferson Panhandle 3.86 3 6.6 17 Agricultural 30.58314 -83.72990 211 56 LE93 Lee South 4.88 3 6.9 3 Agricultural 26.45639 -81.62487 201 8 LEBIG Leon Panhandle 1.02 1 1.3 31 Reference 30.67379 -84.22337 211 99 LECALU Lee South 1.24 2 3.6 29 Reference 26.72535 -81.65383 200 56 LECOWS Lee South 3.74 3 3.2 35 Agricultural 26.69032 -81.58596 202 10 LEDAWG Lee South 5.91 4 6.3 23 Urban 0.00000 0.00000 159 0 LEDGE Leon Panhandle 7.09 4 8.6 11 Urban 30.48783 -84.27853 211 99 LEDONE Levy North 1.03 1 1.9 58 Reference 29.28337 -82.61729 154 13 LELOST Levy North 1.00 1 1.2 59 Reference 29.13109 -82.62290 218 43 LEMESS Leon Panhandle 6.26 4 8.3 6 Urban 30.44130 -84.32875 211 15 LEROLL Lee South 6.07 4 6.2 18 Urban 26.58418 -81.82126 200 20 LIANF2 Liberty Panhandle 1.29 2 1.3 57 Reference 30.26229 -84.82198 212 58 MADEEP Martin South 1.00 1 1.2 48 Reference 27.00919 -80.14564 175 58 MAPATA Marion Central 2.53 2 3.9 34 Agricultural 29.01557 -82.38937 218 56 OKAIR Okaloosa Panhandle 1.12 1 1.9 49 Reference 30.42565 -86.75117 184 56 OKJEB Okaloosa Panhandle 6.61 4 5.3 30 Urban 30.41486 -86.79694 186 48

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162Site name County Region LDI LDI quartile WRAP FWCI Land use Latitude Longitude Julian date Maximum water depth OREOD Orange Central 5.24 4 6.0 20 Urban 28.46612 -81.27323 191 0 ORIFOR Orange Central 7.57 4 7.2 29 Urban 28.42875 -81.47282 225 41 ORUCF Orange Central 3.08 2 2.4 38 Urban 28.60367 -81.19331 191 41 ORVILA Orange Central 5.36 4 6.0 7 Urban 28.45969 -81.26527 191 0 ORXMAS Orange Central 1.00 1 1.3 47 Reference 28.50238 -80.98923 191 0 OSBILY Osceola Central 1.00 1 1.1 45 Reference 28.03325 -81.01988 204 30 OSCRES Osceola Central 3.87 3 6.3 2 Agricultural 28.04368 -81.03569 204 30 OSMIKY Osceola Central 7.78 4 6.0 36 Urban 28.31473 -81.54881 225 41 OSMINI Osceola Central 7.20 4 6.1 24 Urban 28.32963 -81.53056 225 64 OSTNC1 Osceola Central 1.00 1 1.8 47 Reference 28.07775 -81.39177 204 58 PABULL Pasco Central 3.47 2 5.7 11 Agricultural 28.47059 -82.11763 169 56 PBCAMP Palm Beach South 1.00 1 1.4 46 Reference 26.86817 -80.41380 215 56 PBGARL Palm Beach South 1.00 1 1.7 43 Reference 26.86345 -80.23689 174 56 PBHIGH Palm Beach South 5.47 4 6.0 25 Urban 26.70849 -80.20811 215 0 PBLOXR Palm Beach South 1.00 1 1.0 53 Reference 26.95371 -80.18218 176 36 PBROLL Palm Beach South 3.55 2 5.5 24 Agricultural 26.87612 -80.21323 174 56 PBSNAP Palm Beach South 2.22 2 5.9 34 Urban 26.82646 -80.15216 174 43 PBSTAR Palm Beach South 5.79 4 7.1 17 Urban 26.65032 -80.21265 214 0 PBTHAI Palm Beach South 7.76 4 8.4 5 Urban 26.73417 -80.11917 174 0 PBWOOD Palm Beach South 6.92 4 4.6 22 Urban 26.62327 -80.20051 214 0 PBZEP Palm Beach South 1.06 1 1.8 49 Reference 26.73059 -80.25719 215 13 POCOWS Polk Central 4.01 3 5.5 28 Agricultural 28.06925 -81.42633 204 56 POEGGS Polk Central 3.79 3 8.1 8 Agricultural 28.24835 -82.09155 169 61 POFECT Polk Central 1.00 1 1.3 49 Reference 27.78911 -81.46244 205 46

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163Site name County Region LDI LDI quartile WRAP FWCI Land use Latitude Longitude Julian date Maximum water depth PUPLUM Putnam North 1.73 2 3.0 54 Silviculture 29.81470 -81.83806 222 56 PURICE Putnam North 1.58 2 3.1 46 Silviculture 29.67920 -81.73589 224 53 SAMYK2 Sarasota Central 1.01 1 1.5 52 Reference 27.23897 -82.34602 195 41 SERENE Seminole Central 1.00 1 1.0 40 Reference 28.77954 -81.39938 225 53 SJJULI St. Johns North 4.76 3 4.8 39 Urban 30.11208 -81.62379 208 41 SLBOGR St. Lucie Central 4.24 3 4.8 8 Agricultural 27.53780 -80.64056 236 13 SLLIME St. Lucie Central 4.23 3 6.3 14 Agricultural 27.43535 -80.64837 236 41 SRGAR Santa Rosa Panhandle 1.00 1 1.3 51 Reference 30.47204 -87.07998 186 20 SRGENE Santa Rosa Panhandle 1.59 2 4.9 40 Silviculture 30.83094 -86.96921 187 41 SRJAY Santa Rosa Panhandle 5.29 4 6.1 11 Agricultural 30.77000 -87.14000 187 56 SUMMER Sumter Central 1.12 1 1.4 45 Reference 28.39453 -81.97122 169 48 SUWITH Sumter Central 1.07 1 3.0 47 Reference 28.48059 -82.00019 169 20 WAGLOV Washington Panhandle 3.64 2 6.7 8 Agricultural 30.61916 -85.74185 229 102 WAHELL Franklin Panhandle 1.71 2 4.9 46 Silviculture 29.95437 -84.59852 178 0 WAMARK Wakula Panhandle 1.02 1 1.2 54 Reference 30.04041 -84.44025 178 64 WASIV Wakula Panhandle 1.79 2 3.3 49 Silviculture 30.17982 -84.16472 180 66 WAWASH Walton Panhandle 1.06 1 1.0 54 Reference 30.35190 -86.17112 183 71

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APPENDIX E SAMPLED AMPHIBIAN SPECIES ATTRIBUTES

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165Species Common Name Sitesa Ephemeralb Wetland Indigenous/c Exotic Eggd Placement Acris gryllus Southern Cricket Frog 53 Fac Indigenous attached Ambystoma talpoideum Mole Salamander 1 Obl Indigenous float Bufo marinus Giant Toad 1 Fac Exotic attach and float Bufo quercicus Oak Toad 28 Obl Indigenous attach and float Bufo terrestris Southern Toad 20 Fac Indigenous float Bufo woodhousii fowleri Fowler's Toad 1 Fac Indigenous float Eleutherodactylus planirostris Greenhouse Frog 25 Ind Exotic terrestrial Gastrophryne carolinensis Eastern Narrowmouth Toad 39 Obl Indigenous float Hyla chrysoscelis Gray Treefrog 6 Fac Indigenous attach and float Hyla cinerea Green Treefrog 71 Fac Indigenous attach and float Hyla femoralis Pinewoods Treefrog 45 Obl Indigenous attached Hyla gratiosa Barking Treefrog 18 Obl Indigenous attached Hyla squirella Squirrel Treefrog 67 Obl Indigenous attach and float Notophthalmus peristriatus Striped Newt 1 Obl Indigenous attached Notophthalmus viridescens Eastern Newt 7 Fac Indigenous attached Osteopilus septentrionalis Cuban Treefrog 11 Fac Exotic float Pseudacris ocularis Little Grass Frog 38 Obl Indigenous attached Rana capito Gopher Frog 2 Obl Indigenous attached Rana catesbeiana Bullfrog 25 Fac Indigenous float Rana clamitans Bronze Frog 25 Fac Indigenous attach and float Rana grylio Pig Frog 31 Fac Indigenous attach and float Rana sphenocephala Southern Leopard Frog 71 Fac Indigenous attached Scaphiopus holbrookii Eastern Spadefoot 2 Obl Indigenous ? a= Number of sites a species was present. b= Use of ephemeral wetlands (Fac= facultativ e, Obl= obligate, Ind= independent)(Moler and Franz 1987). c= Indigenous or exotic to Florida. d= Egg placement (attached = species atta ches eggs to wetland vegetation or substrate, attach and float = species is known to attach eggs to objects or eggs or free floating, float= species eggs are free floating on the water surface or within the water co lumn, and terrestrial= species eggs ar e deposited and mature on land).

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APPENDIX F SAMPLED AVIAN SPECIES ATTRIBUTES

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167Species sampled in 2003 Sitesa Dietb Foragingc Nest Locationd Wetland Dep.e Indigenous/ Exoticf MigrationgMass(g)h Territoriali Population Trendj Acadian Flycatcher 4 I S C I I N 13 yes -8.3 American Crow 38 O GG C I I R 450 yes NA Anhinga 3 C WB S D I R 2100 no 4.1 American Swallow-tailed Kite 5 C H C I I N 420 ? 3.9 Barred Owl 13 C H CV I I R 720 yes NA Bachman's Sparrow 13 O GG G I I R 19.5 yes -2.9 Barn Swallow 6 I AS HS I I N 19 no 5.3 Black-crowned Night-Heron 1 C WB C D I R 870 no NA Belted Kingfisher 1 C WB CV D I R 150 yes NA Bell's Vireo 1 I CG S I I N 8.5 yes NA Blue-gray Gnatcatcher 39 I CG C I I R 6 yes -2 Brown-headed Cowbird 13 O GG P I I R 44 no NA Brown-headed Nuthatch 12 O BG CV I I R 10 yes -6.9 Blue Grosbeak 12 O GG S I I N 28 yes NA Blue Jay 55 O CG C I I R 85 no NA Black Vulture 13 C SV G I I R 2000 no 3.4 Brown Thrasher 13 O GG S I I R 69 yes NA Boat-tailed Grackle 10 O GG C I I R 168 no 2.1 Carolina Chickadee 26 I CG CV I I R 10.5 yes NA Cattle Egret 24 I GG C I I R 340 no NA Carolina Wren 94 I CG CV I I R 21 yes 0.9 Chimney Swift 18 I AS HS I I N 23 no NA Eurasian Collared-Dove 10 H GG C I E R 200 yes 35.3 Common Ground-Dove 7 H GG G I I R 30 no -3.8 Common Grackle 36 O GG C I I R 115 no -1.3 Cooper's Hawk 3 C H C I I R 450 yes 24.2 Common Moorhen 5 O GG G D I R 315 ? NA Common Nighthawk 20 I AS G I I N 62 yes -4.2 Common Yellowthroat 31 I CG S I I R 10 yes -1.6

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168Species sampled in 2003 Sitesa Dietb Foragingc Nest Locationd Wetland Dep.e Indigenous/ Exoticf MigrationgMass(g)h Territoriali Population Trendj Chuck-will's-Widow 13 I AS G I I N 120 ? -1.1 Downy Woodpecker 54 I BG CV I I R 27 yes NA Eastern Bluebird 19 I GG CV I I R 31 yes NA Eastern Kingbird 12 I S C I I N 40 yes -4.5 Eastern Meadowlark 10 I GG G I I R 90 yes -4.7 Eastern Screech-Owl 1 C H CV I I R 180 yes NA Eastern Wood-Pewee 2 I S C I I N 14 yes NA European Starling 2 O GG HS I E R 82 no NA Fish Crow 27 O SV C I I R 280 no NA Great Blue Heron 6 C WB C D I R 2400 no NA Great Crested Flycatcher 42 I S CV I I N 34 yes NA Great horned Owl 3 C H C I I R 1400 yes NA Gray Catbird 1 O GG S I I R 37 yes NA Great Egret 13 C WB C D I R 870 no -3 Green Heron 16 C WB C D I R 210 no NA House Finch 9 H GG C I E R 21 no 72.2 House Sparrow 8 H GG HS I E R 28 no -3.4 Hooded warbler 1 I CG S I I N 10.5 yes -3.7 Indigo Bunting 8 O CG S I I N 14.5 yes -4.3 Kentucky Warbler 1 I GG G I I N 14 yes NA Killdeer 5 I GG G I I R 95 yes -2.3 Laughing Gull 1 O SV G I I R 320 no NA Little Blue Heron 16 C WB C D I R 340 no NA Least Bittern 1 C WB G D I R 80 ? NA Least Tern 2 C WB G D I N 42 ? NA Limpkin 1 C GG G D I R 1100 ? NA Loggerhead Shrike 11 I H C I I R 48 yes -3.6 Louisiana Waterthrush 3 I GG G D I N 20.5 yes NA Mississippi Kite 2 C H C I I N 280 no NA

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169Species sampled in 2003 Sitesa Dietb Foragingc Nest Locationd Wetland Dep.e Indigenous/ Exoticf MigrationgMass(g)h Territoriali Population Trendj Mourning Dove 66 H GG C I I R 120 yes 2.7 Mottled Duck 1 O WB G D I R 1000 ? NA Monk Parakeet 2 H CG C I E R 100 ? NA Muscovy Duck 3 H GG G D E R 2250 no NA Northern Bobwhite 15 H GG G I I R 170 no -3.6 Northern Cardinal 95 O GG S I I R 45 yes NA Northern Flicker 5 I GG CV I I R 130 yes -5.4 Northern Mockingbird 48 O GG S I I R 49 yes -1.2 Northern Parula 19 I CG C I I N 8.6 yes NA Northern Rough-winged Swallow 1 I AS CV I I R 16 ? NA Orchard Oriole 2 I CG C I I N 19 yes -1.6 Osprey 1 C WB C D I R 1600 no 3.6 Pied-billed Grebe 1 C WB G D I R 450 ? NA Pine Warbler 23 I CG C I I R 12 yes -2.1 Pileated Woodpecker 25 I BG CV I I R 290 yes NA Prothonotary Warbler 4 I CG CV D I N 16 yes -3.1 Purple Martin 20 I AS HS I I N 56 no -2.7 Red-bellied Woodpecker 85 O BG CV I I R 63 no NA Red-cockaded Woodpecker 3 I BG CV I I R 44 yes NA Red-eyed Vireo 3 I CG C I I N 17 yes -3.9 Red-headed Woodpecker 7 O S CV I I R 72 yes -4.5 Rock Dove 2 H GG HS I E R 270 no NA Red-shouldered Hawk 45 C H C I I R 630 yes NA Rufous-sided Towhee 59 O GG G I I R 40 yes -2.1 Red-tailed Hawk 4 C H C I I R 1080 yes NA Ruby-throated Hummingbird 3 O CG C I I N 3.2 no NA Red-winged Blackbird 21 O GG G I I R 52 yes -3.5 Sandhill Crane 6 O GG G I I R 4370 no 3.5 Snowy Egret 4 C WB C D I R 360 no NA

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170Species sampled in 2003 Sitesa Dietb Foragingc Nest Locationd Wetland Dep.e Indigenous/ Exoticf MigrationgMass(g)h Territoriali Population Trendj Summer Tanager 12 I CG C I I N 29 yes NA Tree Swallow 1 I AS CV I I R 20 ? NA Tricolored Heron 5 C WB C D I R 380 no NA Tufted Titmouse 42 I CG CV I I R 21.5 yes -1.2 Turkey Vulture 29 C SV G I I R 1830 no NA White-breasted Nuthatch 1 I BG CV I I R 21 yes NA White-eyed Vireo 33 I CG S I I R 11.5 yes NA White Ibis 10 C WB C D I R 900 no NA Wild Turkey 5 H GG G I I R 5800 no 7.7 Wood Duck 6 H GG CV D I R 600 no NA Wood Stork 3 C WB C D I R 2400 no NA White-winged Dove 1 H GG C I E R 150 no 15.3 Yellow-billed Cuckoo 10 I CG C I I N 65 ? -1.3 Yellow-crowned Night-Heron 4 C WB C D I R 690 no NA Yellow Warbler 2 I CG C I I N 9.5 yes NA Yellow-throated Vireo 2 I CG C I I N 18 yes NA Yellow-throated Warbler 1 I CG C I I N 9.4 yes -3.5 a= Number of sites a species was present. b= Species predominant diet (C=carnivore, H=he rbivore, I=insectivore, and O=omnivore). c= Foraging strata or technique (S = sallier, SV= scavenger, BG= bark gleaner, CG= canopy gleaner, GG= ground gleaner, H= hawker, WB= water based, and AS= air screener). d= Primary nest location or strategy (C= canopy, CV= cavity, G= ground, HS= human structure, P= parasite, and S= shrub). e= Wetland dependency (I= independent and D= dependent). f= Indigenous or exotic to Florida (I= indigenous and E= exotic). g= Florida residency (N= Neotropical migrant and R= resident). h= Approximate species mass. i= Maintains feeding territories. j= Breeding Bird Survey si gnificant (p<0.1) Florida population trends 1966-2002.

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194 BIOGRAPHICAL SKETCH James Anthony Surdick Jr. was born Fe bruary 21, 1971, in Mountain View, California. Soon thereafter his family moved back to the birthplace of his parents, James and Joyce, in Watertown, Wisconsin, where J im spent the next 16 years. It was during this time that a strong sense of biophilia es tablished in Jim via countless hours spent in “the marsh”, fishing trips with grandpa, and family vacations to local state parks. At best, high school and Jim were mutually to lerant of each other and a diploma was awarded in 1989 from Watertown High School. The study of nature continued, and a new sense of academic direction was found, at the University of Wisconsin Madison where he earned a Bachelor of Science degree in May of 1993, from the Department of Wildlif e Ecology. Summers during his studies at UW and the three years following graduation were spent working for wildlife agencies in Wisconsin, Minnesota, and Missouri. It wa s around this time that a minor case of wanderlust reared and several journeys were taken to wilderness locations within the United States. Then one cold December morning a call came from the University of Florida and the next two years were spent expl oring the Everglades. A Master of Science degree was received from the Department of Wildlife Ecology and Conservation in 1998. As part of Jim’s continued education outside of an institution he traveled through Central America, Ecuador, Alaska, southern Africa, a nd Mexico. The last five years were spent investigating the biotic comm unities of isolated wetlands throughout Florida and now he is working for the Florida Natural Areas Inventory.


Permanent Link: http://ufdc.ufl.edu/UFE0010745/00001

Material Information

Title: Amphibian and avian species composition of forested depressional wetlands and circumjacent habitat : the influence of land use type and intensity
Physical Description: Mixed Material
Language: English
Creator: Surdick, James A. 1971- ( Dissertant )
Brown, Mark T. ( Thesis advisor )
Cumming, Graeme ( Reviewer )
Franz, Richard ( Reviewer )
Montague, Clay ( Reviewer )
Publisher: University of Florida
Place of Publication: Gainesville, Fla.
Publication Date: 2005
Copyright Date: 2005

Subjects

Subjects / Keywords: Environmental Engineering Sciences thesis, Ph.D
Dissertations, Academic -- UF -- Environmental Engineering Sciences
Spatial Coverage: United States--Florida

Notes

Subject: Wetlands provide wildlife habitat; however, circumjacent anthropogenic land uses influence abiotic and biotic aspects of wildlife habitat and in turn possibly which species will persist in a landscape. The main objective of this study was to record the amphibian and avian species composition of depressional forested wetlands embedded in landscapes of varying human land use intensity and relate compositional differences to wetland and landscape characteristics. Presence/absence surveys for amphibian and avian species were conducted at 111 small (<2.5 ha) forested depressional wetlands throughout Florida within four common land use types, natural or reference, silviculture, agriculture, and urban/residential. Despite a modest relationship between land use and species richness, the results of nonmetric multidimensional scaling ordinations and multi-response permutation procedures indicated a strong association with amphibian and avian species composition and land use. Indicator species, species sensitive and tolerant of human land uses, were identified. The relevance of these results may broaden with the finding that not only were there species-specific responses, but also predominate life history traits of the avian and amphibian community varied with land use context. For example, amphibians that were obligatory ephemeral pond breeders decreased with increasing land use intensity. Within the avian community, insectivores, bark gleaners, canopy gleaners, territorial species, ground nesters, and cavity nesters decreased, while omnivores, herbivores, ground gleaners, canopy nesters, and exotic species increased, with increasing land use intensity. The Landscape Development Intensity index was a remotely measured predictor of amphibian and avian species assemblages. Amphibian species composition was also significantly associated with the Florida Wetland Condition Index (FWCI), distance to the nearest wetland, maximum wetland water depth, wetland water pH, wetland water specific conductivity, and geographical location of the study sites. Avian species composition was also significantly associated with the proportion of agricultural land within 200 meters, FWCI, proportion of retention ponds/canals within 200 meters, and geographical location of the study sites. The results of this study suggest that merely preserving wetland habitat in developed landscapes will not be enough to support a wildlife community indicative of natural settings. However, wetlands embedded in even the most intensive land uses provide valuable wildlife habitat.
Subject: agriculture, amphibian, avian, indicator, integrity, silviculture, urban, wetland
General Note: Title from title page of source document.
General Note: Document formatted into pages; contains 207 pages.
General Note: Includes vita.
Thesis: Thesis (Ph.D.)--University of Florida, 2005.
Bibliography: Includes bibliographical references.
General Note: Text (Electronic thesis) in PDF format.

Record Information

Source Institution: University of Florida
Holding Location: University of Florida
Rights Management: All rights reserved by the source institution and holding location.
System ID: UFE0010745:00001

Permanent Link: http://ufdc.ufl.edu/UFE0010745/00001

Material Information

Title: Amphibian and avian species composition of forested depressional wetlands and circumjacent habitat : the influence of land use type and intensity
Physical Description: Mixed Material
Language: English
Creator: Surdick, James A. 1971- ( Dissertant )
Brown, Mark T. ( Thesis advisor )
Cumming, Graeme ( Reviewer )
Franz, Richard ( Reviewer )
Montague, Clay ( Reviewer )
Publisher: University of Florida
Place of Publication: Gainesville, Fla.
Publication Date: 2005
Copyright Date: 2005

Subjects

Subjects / Keywords: Environmental Engineering Sciences thesis, Ph.D
Dissertations, Academic -- UF -- Environmental Engineering Sciences
Spatial Coverage: United States--Florida

Notes

Subject: Wetlands provide wildlife habitat; however, circumjacent anthropogenic land uses influence abiotic and biotic aspects of wildlife habitat and in turn possibly which species will persist in a landscape. The main objective of this study was to record the amphibian and avian species composition of depressional forested wetlands embedded in landscapes of varying human land use intensity and relate compositional differences to wetland and landscape characteristics. Presence/absence surveys for amphibian and avian species were conducted at 111 small (<2.5 ha) forested depressional wetlands throughout Florida within four common land use types, natural or reference, silviculture, agriculture, and urban/residential. Despite a modest relationship between land use and species richness, the results of nonmetric multidimensional scaling ordinations and multi-response permutation procedures indicated a strong association with amphibian and avian species composition and land use. Indicator species, species sensitive and tolerant of human land uses, were identified. The relevance of these results may broaden with the finding that not only were there species-specific responses, but also predominate life history traits of the avian and amphibian community varied with land use context. For example, amphibians that were obligatory ephemeral pond breeders decreased with increasing land use intensity. Within the avian community, insectivores, bark gleaners, canopy gleaners, territorial species, ground nesters, and cavity nesters decreased, while omnivores, herbivores, ground gleaners, canopy nesters, and exotic species increased, with increasing land use intensity. The Landscape Development Intensity index was a remotely measured predictor of amphibian and avian species assemblages. Amphibian species composition was also significantly associated with the Florida Wetland Condition Index (FWCI), distance to the nearest wetland, maximum wetland water depth, wetland water pH, wetland water specific conductivity, and geographical location of the study sites. Avian species composition was also significantly associated with the proportion of agricultural land within 200 meters, FWCI, proportion of retention ponds/canals within 200 meters, and geographical location of the study sites. The results of this study suggest that merely preserving wetland habitat in developed landscapes will not be enough to support a wildlife community indicative of natural settings. However, wetlands embedded in even the most intensive land uses provide valuable wildlife habitat.
Subject: agriculture, amphibian, avian, indicator, integrity, silviculture, urban, wetland
General Note: Title from title page of source document.
General Note: Document formatted into pages; contains 207 pages.
General Note: Includes vita.
Thesis: Thesis (Ph.D.)--University of Florida, 2005.
Bibliography: Includes bibliographical references.
General Note: Text (Electronic thesis) in PDF format.

Record Information

Source Institution: University of Florida
Holding Location: University of Florida
Rights Management: All rights reserved by the source institution and holding location.
System ID: UFE0010745:00001


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AMPHIBIAN AND AVIAN SPECIES COMPOSITION OF FORESTED
DEPRESSIONAL WETLANDS AND CIRCUMJACENT HABITAT:
THE INFLUENCE OF LAND USE TYPE AND INTENSITY















By

JAMES A. SURDICK JR.


A DISSERTATION PRESENTED TO THE GRADUATE SCHOOL
OF THE UNIVERSITY OF FLORIDA IN PARTIAL FULFILLMENT
OF THE REQUIREMENTS FOR THE DEGREE OF
DOCTOR OF PHILOSOPHY

UNIVERSITY OF FLORIDA


2005

































Copyright 2005

by

James A. Surdick Jr.















ACKNOWLEDGMENTS

I am not only grateful for the opportunity provided to me by my major advisor Dr.

Mark Brown, but also for his catalytic approach to learning which allowed me to pursue

my interests. I have gained high quality information, and an amazing experience

exploring the systems of Florida. I thank my committee members Graeme Cumming,

Richard Franz, and Clay Montague for their time, effort, and diligent input. I am forever

thankful to Kris Sullivan for her tireless energy, curiosity, and support. Friends essential

to this study include, but are not limited to, Matt Cohen, Tony Davanzo, Chuck Lane,

Kelly Reiss, and Ben Vivas. There are too many individual land owners, land managers,

and facilitators to mention, but without their effort and consent this work would not have

been possible. To the land owners who did not allow me on to their property you now

can see I did not have ulterior motives and really did just want to tally swamp critters.

The Florida Department of Environmental Protection provided the funding to Dr. Brown

and is the ultimate supporter of this research.
















TABLE OF CONTENTS

page

A C K N O W L E D G M E N T S ........... ......... ....................................................................... iii

LIST OF TABLES ........................................................ vii

LIST OF FIGURES ............................... ... ...... ... ................. .x

ABSTRACT ........ .............. ............. .. ...... .......... .......... xii

CHAPTER

1 IN TR OD U CTION ............................................... .. ......................... ..

State ent of the P problem .................................................................................. 1
P lan of Stu dy ...................................................... .................... 4
Stu dy Sy stem s ................................................................................................ ..... .. 5
W wildlife Response to Land Use Type................................ ......................... ....... 8
"R reference" H abitats ....................................... ......... .... .... .. ........ ..
Agriculture .................. ............................................... ........ 12
Amphibians and agricultural landscapes........................ ............... 13
Birds and agricultural landscapes...................................... ............... 16
Silv icu ltu re .................... ....................................................... 19
Amphibians and silvicultural landscapes ............................................. 19
Birds and silvicultural landscapes ............. ......... ..................................22
U rban/R esidential/R oadw ays ........................................ ......................... 24
Am phibians and urban landscapes ........................................ ............... 25
B irds and urban landscapes .................................. ..................................... 28
R o ad s .................. .................. ... ............ ............... .. .. ............... .. 34
Studies on the Response of Amphibians and Birds to Multiple Land Use Types36
H habitat Fragm entation ................. .......... .... .... ...... .... ... .... ........ .. ..............4 1
Amphibian and Avian Assemblages as Indicators of Land Use Intensity..................43
Landscape Development Intensity (LDI) Index ............................... ................45

2 M E T H O D S .................................................................................. .........................4 7

Plan of Study.........................................................................................47
Site Selection ................................................................... 47
F ield S am pling ..........................................................5 1
Visual and Auditory Encounter Surveys ........................................ .....52









A utom ated R ecorder Surveys....................................... ........................... 53
Com bined Sam pling Effort......................................................... ............... 53
D ip-net Sam pling ............ .......................................................... .......... .... 54
Terrestrial Insect Sam pling........................................... ........................... 55
O their B iotic Sam pling ............................................................ ............... 55
E nvironm ental V ariables .................................................. .............................. 56
W after Sam pling ............................. .... ..... ................... ........... ........ 56
Fire History ........................ ..................... 56
L landscape V ariables......................................................... .. ..........57
Landscape Development Intensity Index ............................................. 57
Landscape indices ............................................. ............ ............. 58
Data Analysis.............................. .......................60
W ater Chemistry and Landscape Variables ................................ ............... 61
S p e cie s R ich n e ss ........................................................................................... 6 1
Species C om position ..................... .. ........................ .. ...... ........... 62
In d ic ato r S p e cie s ........................................................................................... 6 5
Com m unity Characteristics ........................................ ........................... 66
A m phibian com m unity .................................................... .................. 66
Avian community .......................................... .........................67

3 R E S U L T S .............................................................................6 8

W etland Characteristics ..............................................................................68
W etland V egetation ............................................... ...... ..... ........ .... 68
W etland W ater D epth .......................................................... ............... 69
W etland W ater Chem istry ............................................................................70
W etland Terrestrial Insects........................................................ ............... 70
Circumjacent Land Use Characteristics................................................ 71
Landscape Development Intensity................................... ..................... 72
Amphibians .............. ............................. .................. 73
A m phibian Species R ichness ........................................ ......................... 73
Amphibian Species Frequency of Occurrence .......................................... 76
A m phibian B reeding Effort............................................................ ............... 78
Am phibian Species Com position ............................................. ............... 78
A m phibian Indicator Species ...................... ......... ......... ............... .... 87
Amphibian Community Characteristics and Fish Composition..........................90
B ird s ................................................................................. 94
B ird Species R ichness ................... .... .......... ............. .... ........94
Bird Species Com position ............................................................................97
B ird Indicator Species ............................................. ................ ............. 103
Bird Community Characteristics ........................................... ...............105

4 D ISCU SSION ...................................... ......................................... 113

Species R richness .................................................................. ................ 114
Landscape Versus Wetland Species Richness .................................................. 115
Species Com position .................. ............................... .. ...... .. ........ .... 118


v









Land U se Type and Intensity........................................ ...................... ......... 118
Species Composition Thresholds and Gradients..............................................119
Correlates of Species Com position ....................................... ............... 123
Correlates of amphibian species composition......................................123
Correlates of avian species com position ...................................................127
In dictator Sp ecies............................................................................. ............... 12 8
A m phibian Indicator Species ........................... .... .. .. ......... ...............129
Amphibian indicators of reference condition and low intensity land uses. 130
Amphibian indicators of high intensity land uses and ubiquitous species.134
A vian Indicator Species............................................. ............................ 136
Species G uilds and L and u se ........................................................................ ... ... 139
A m phibian G uilds ........................... ........ ............................ ...... ............139
A v ian G u ild s................................. .... ... .. ............. ... .. .................... 14 1
Landscape Development Intensity (LDI) Index and Species Composition............145
V ariability in LD I index ........................................................ ............. 146
V ariability in the L andscape......................................... .......... ............... 147
C onclu sion ..................................................................................................... 149

APPENDIX

A SITE SU R V E Y LEN G TH S ........................................................... .....................151

B CORRELATES OF AMPHIBIAN SPECIES COMPOSITION.............................154

C CORRELATES OF AVIAN SPECIES COMPOSITION.......................................156

D SITE CH A R A C TER ISTIC S ......................................................... .....................158

E SAMPLED AMPHIBIAN SPECIES ATTRIBUTES.............................................164

F SAMPLED AVIAN SPECIES ATTRIBUTES ............. ................................... 166

L IST O F R E F E R E N C E S ...................................................................... ..................... 17 1

BIOGRAPHICAL SKETCH ............................................................. ............... 194
















LIST OF TABLES


Table page

2-1 Mean wetland size (ha) and the minimum and maximum wetlands sampled
per land-use category and region. ............................................... ............... 48

2-3 Independent variables included in stepwise regressions predicting the
am phibian based NM S axes scores.................................. ......................... 65

3-1 The mean and standard deviation of the wetland macrophyte species richness,
percent canopy cover, and tree basal area (>10.2 dbh) by land use.....................69

3-2 The mean proportion and standard deviation of the basal area of the wetland
overstory vegetation by land use..................................... ......................... 69

3-3 The mean and standard deviation of the maximum wetland water depth at
sites with standing water by land use, and the count of sites without standing
water at the time of sampling by land use................................... ............... 70

3-4 Mean and standard deviation of water chemistry parameters by land use. ........71

3-5 The mean and standard deviation (in parentheses) of attributes of the
landscape surrounding the sample wetlands by land use type..............................72

3-6 The range, mean, and standard deviation of LDI scores by LDI quartile, and
the number of sites by land use type for each LDI quartile................................72

3-7 The Pearson correlation coefficients between the site's 100m buffer LDI score
and selected environmental variables. ...................................... ............... 74

3-8 Amphibian species richness by land use, 100 meter LDI quartile, and Florida
re g io n ........................................................................... 7 5

3-9 Amphibian frequency of occurrence by land use type and all sites combined. ....79

3-10 MRPP test for significant differences in amphibian species composition
between land use types and LDI quartiles. ................................. .................82

3-11 Coefficients of determination (R2) between amphibian based NMS ordination
distances and distances in the original space .......... .... ............. ..................... 83









3-12 Pearson's correlation coefficients (r) and Kendall's Tau comparisons with
NMS ordination axes based on amphibian species composition and LDI,
FW CI, and maximum water depth. .......................................... ................... 84

3-13 Independent variables included in stepwise regressions predicting the
am phibian based NM S axes scores.................................. ......................... 84

3-14 The variables and regression coefficients selected (p<0.05) from a stepwise
regression using the Landscape variables listed in Table 3-13 to represent the
amphibian based NM S ordination axes. ..................................... ............... 85

3-15 The variables and regression coefficients selected (p<0.05) from a stepwise
regression using the Landscape, Focal Wetland, and Sampling variables listed
in Table 3-13 to represent the amphibian based NMS ordination axes. ................86

3-16 The variables and regression coefficients selected (p<0.05) from a stepwise
regression using the Landscape, Focal Wetland, Sampling, Water Chemistry
variables listed in Table 3-13 to represent the amphibian based NMS ordination
ax e s. ........................................ .................... ................ 8 6

3-17 Significant (p<0.10) amphibian indicator species values by LDI quartile with
quartiles three and four com bined...................................... ........................ 89

3-18 Significant (p<0.10) amphibian indicator species values by land use .................90

3-19 Significant (p<0.10) amphibian indicator species values by land use
(agriculture and urban versus reference and silviculture)...............................90

3-20 The mean proportion of amphibian species independent of wetlands for
breeding, dependent on ephemeral wetlands, and will facultatively use
ephemeral wetlands by land use and LDI quartile...............................................91

3-21 The number of sites where fish species occurred by land use ............................93

3-22 Avian species richness by land use, LDI quartile, and Florida region.. ................96

3-23 Mean number and standard deviation of bird species detected within the
wetlands, landscape, total, and landscape to wetland ratio during two 15 min
surveys by land use, LDI quartile, Florida region and all sites combined.............97

3-24 MRPP test for significant differences in bird species composition between
sites by land use and LD I score quartiles.................................... ..................98

3-25 Coefficients of determination (r2) between avian based NMS ordination
distances and distances in the original space .............. ............. ..................... 100









3-26. Pearson's correlation coefficients (r) and Kendall's Tau comparisons with NMS
ordination axes based on avian species composition and the environmental
variables with the strongest correlation. ................................... ............... 101

3-27 Independent variables included in stepwise regressions to predict the avian
based N M S axes scores............................................... ............................. 101

3-28 The selected variables and regression coefficients selected (p<0.05) from a
stepwise regression using the Landscape variables listed in Table 3-27 to
represent the avian based NMS ordination axes ............................................102

3-29 The selected variables and regression coefficients selected (p<0.05) from a
stepwise regression using the Landscape, Focal Wetland, and Sampling
variables listed in Table 3-27 to represent the avian based
N M S ordination axes. ........................................ .......................................... 102

3-30 The significant (p<0.10) avian indicator values of land use.............................104

3-31 The significant (p<0.10) avian indicator values of LDI quartiles........................105

3-32 Mean proportion of sampled bird species predominate foraging strategy by
land use and L D I quartile............................................. ............................ 108

3-33 Mean proportion of sampled bird species predominate nesting strategy by
land use and L D I quartile............................................. ............................ 109















LIST OF FIGURES


Figure pge

1-1 Systems diagram of landscape biotic composition storage and driving energies,
including hum an influence.......................................................................... .. .. ...2

2-1 The distribution of the survey sites by land use and Florida region....................48

2-2 The mean proportion of land cover type within 200 meters of each focal
w etland by land use categories.. ......................... .............................................50

3-1 Mean and standard deviation of amphibian species richness within the
wetland and within the wetland and the 200m buffer by land use.......................77

3-2 Mean amphibian species richness within the wetland and within the wetland
and the 200m buffer by LDI quartile.. ........................................ ............... 77

3-3 The mean tadpole count, mean number of tadpole species, mean number of
anuran species calling and mean maximum intensity of anuran species calling
b y la n d u se ....................................................... ................ 8 0

3-4 The mean tadpole count, mean number of tadpole species, mean number of
anuran species calling on cassette tapes, and mean maximum intensity of
anuran species calling by LDI quartile. ..................................... ............... 80

3-5 Unrotated joint plot of the amphibian based NMS ordination site scores for
axis one and axis three by land use with the strength and relationship of three
environmental variables (FWCI, LDI, and maximum depth).............................83

3-6 Unrotated joint plot of the amphibian based NMS ordination species scores for
axes one and three with the strength and relationship of three environmental
variables (FW CI, LDI, and maximum depth)........................................................88

3-7 The proportion of all sites and only sites that had standing water in 2003
where fish were encountered by land use. .................................. .................92

3-8 The proportion of the amphibian community that attaches eggs to substrates,
has free floating eggs, attaches or has floating eggs, and has terrestrial egg
develop ent by land u se .......................................................................... ....... 93









3-9 The proportion of the amphibian community that attaches eggs to substrates,
has free floating eggs, attaches or has floating eggs, and has terrestrial egg
develop ent by LD I quartile. ........................................... .......................... 94

3-10 Unrotated joint plot of the avian based NMS ordination site scores for axes
one and three with the strength and relationship of the landscape variables
LD I and FW C I. .......................................................................99

3-11 Unrotated joint plot of the avian based NMS ordination site scores for axes
one and two with the strength and relationship of three landscape variables
LDI, FWCI, and proportion of the landscape within 200 meters of the
wetland that was used for agriculture. ...................................... ............... 100

3-12 The mean proportion of avian species within each feeding guild by
L D I qu artile ................................................................................. ............... 10 7

3-13 The mean and standard deviation of the proportion of avian species that
maintain feeding territories at each site by LDI Quartile and Land Use............108

3-14 The mean proportion and standard error of the exotic and Neotropical
migrant birds recorded by LDI Quartile and Land Use..................................... 110

3-15 The mean proportion and standard error of bird species that have significant
(p<0.1) population trends in the Breeding Bird Survey data for Florida since
1 9 6 6 .......................................................................... 1 1 1

3-16 The mean weight (g) and standard error of avian species detected at the
study sites by LD I quartile and Land use ..........................................................112

3-17 Loglo of avian mass versus logo of avian territory or home range size............12















Abstract of Dissertation Presented to the Graduate School
of the University of Florida in Partial Fulfillment of the
Requirements for the Degree of Doctor of Philosophy

AMPHIBIAN AND AVIAN SPECIES COMPOSITION OF FORESTED
DEPRESSIONAL WETLANDS AND CIRCUMJACENT HABITAT:
THE INFLUENCE OF LAND USE TYPE AND INTENSITY

By

James A. Surdick Jr.

August 2005

Chair: Mark T. Brown
Major Department: Environmental Engineering Sciences

Wetlands provide wildlife habitat; however, circumjacent anthropogenic land uses

influence abiotic and biotic aspects of wildlife habitat and in turn possibly which species

will persist in a landscape. The main objective of this study was to record the amphibian

and avian species composition of depressional forested wetlands embedded in landscapes

of varying human land use intensity and relate compositional differences to wetland and

landscape characteristics. Presence/absence surveys for amphibian and avian species

were conducted at 111 small (<2.5 ha) forested depressional wetlands throughout Florida

within four common land use types, natural or reference, silviculture, agriculture, and

urban/residential.

Despite a modest relationship between land use and species richness, the results of

nonmetric multidimensional scaling ordinations and multi-response permutation

procedures indicated a strong association with amphibian and avian species composition

and land use. Indicator species, species sensitive and tolerant of human land uses, were









identified. The relevance of these results may broaden with the finding that not only

were there species-specific responses, but also predominate life history traits of the avian

and amphibian community varied with land use context. For example, amphibians that

were obligatory ephemeral pond breeders decreased with increasing land use intensity.

Within the avian community, insectivores, bark gleaners, canopy gleaners, territorial

species, ground nesters, and cavity nesters decreased, while omnivores, herbivores,

ground gleaners, canopy nesters, and exotic species increased, with increasing land use

intensity.

The Landscape Development Intensity index was a remotely measured predictor of

amphibian and avian species assemblages. Amphibian species composition was also

significantly associated with the Florida Wetland Condition Index (FWCI), distance to

the nearest wetland, maximum wetland water depth, wetland water pH, wetland water

specific conductivity, and geographical location of the study sites. Avian species

composition was also significantly associated with the proportion of agricultural land

within 200 meters, FWCI, proportion of retention ponds/canals within 200 meters, and

geographical location of the study sites.

The results of this study suggest that merely preserving wetland habitat in

developed landscapes will not be enough to support a wildlife community indicative of

natural settings. However, wetlands embedded in even the most intensive land uses

provide valuable wildlife habitat.














CHAPTER 1
INTRODUCTION

Statement of the Problem

The increasing spatial extent and intensity of human uses of land worldwide are

affecting both biotic and abiotic aspects of ecological systems and altering ecological

processes (e.g., Vitousek et al. 1997). The result of these alterations is a landscape that is

becoming dominated by ecological systems that Odum (1962) termed, emerging

ecosystems associated with man, and later as interface ecosystems, whose landscape

conditions and driving energies are different from their natural counterparts found in

more remote landscapes (Odum 1971). It is suggested that biological integrity of a

system is maintained when the species composition and functional organization are

comparable to the natural habitats of the region (Karr and Dudley 1981). The biological

integrity of ecological communities within human dominated landscapes is a concern of

many, especially those who are interested in conserving native biota. Despite this

concern, there is still no clear understanding of the changes in wildlife composition along

a gradient of human land use intensity.

In just 100 years Florida's population has increased from 530 thousand to nearly 17

million people (U.S. Census Bureau 2003). By the year 2025, almost 21 million people

are projected to live in Florida (Cambell 1997). With the past and projected population

increases in Florida come major habitat transformations and changes to energy flows

within and between different land systems. Even more changes in the future are likely as

the intensity and spatial extent of Florida's population increases continue. Human










induced changes to ecological communities are complex, and include both direct and

indirect effects.


S Runin Bio- Humar
gRaun i ivers In pact


iestiodiversity



Pred

Vegetation Insects-
Sunnsects Insectivores





Detritus Structure



Landscape


Figure 1-1. Systems diagram of landscape biotic composition, storage, and driving
energies, including human influence.

Figure 1-1 is a systems diagram of an aggregated food web that includes main

driving energies and human influence. Shown are aggregated compartments of the food

web, each containing many species that combined make up landscape biodiversity.

Every landscape has a suite of species that are adapted to its driving energies and habitat

structure, with specific species responding to subtle differences in its physical structure

and the timing and intensities of driving energies. The diagram shows the effect of

humans on each compartment and ultimately on species composition as both direct

(harvesting of species or stress as a result of pollutants or toxins) and indirect (changing

runoff characteristics or increasing nutrient inflows). In addition, the direct effects can

have indirect affects as individual species within compartments adjust to changes in









others. For instance, Polis et al. (1997) suggest bottom up and top down indirect effects

that may result from direct effects. Terborgh et al. (1999) give as an example, hunting

pressure on top carnivores that may affect the entire ecosystem indirectly by influencing

species abundances and persistence in lower levels of the food web. The interplay of

direct, indirect, and cumulative effects makes the identification of specific mechanisms of

shifts in species composition difficult.

With the changes in land use that Florida has experienced in the past 100 years (in

both spatial extent and intensity of use) presumably there have been changes in wildlife

use and persistence. However little research has been conducted that would lend insight

into these changes especially related to the impacts of differing intensity of uses. The

main research question that this dissertation addresses is directed at this lack of

information: How do amphibian and avian communities respond to the anthropogenic

influences associated with different land use intensities?

While this question begs to be answered for the entire landscape and for all wildlife

species, it was obvious from the beginning that the study of human influences on wildlife

communities should be narrowed to a specific ecosystem type and a subset of wildlife in

order to reduce the size of the study to one that was manageable within the time and

resource constraints available. Isolated wetlands offer an interesting point of reference in

that they are embedded in a landscape matrix and thus can be easily treated as individual

cases within varying landscape intensities. Further, isolated wetlands offer the potential

to study avian and amphibian populations as they are relatively productive habitats for

these species. Finally, the macroinvertebrate and vegetative communities of isolated

wetlands have been shown to respond to land use intensity of the circumj acent landscape









(Lane 2003, Reiss 2004); thus, changes in amphibian and avian communities might be

expected.

Plan of Study

In this dissertation, depressional forested wetlands (Cowardin et al. 1979) and their

immediately surrounding landscape (within 200m) that are dominated by human land

uses to varying degrees were studied to better understand wildlife responses to human

land use intensity. The study took place throughout Florida and involved 111 wetlands in

landscapes that were apriori characterized as either reference (no human dominated land

uses), silvicultural, agricultural, or urban. The main objective of this study was to

measure avian and amphibian species composition in these wetlands and their

immediately adjacent uplands and relate it to human land use intensity.

In Florida small (<2.5ha) isolated forested wetlands or dome swamps persist within

primarily four land use categories, reference or natural, silviculture, agriculture, and

urban/residential (Monk and Brown 1965, Kautz 1993). These wetlands serve many

functions including the often cited, support of wildlife (e.g., Moler and Franz 1987,

Semlitsch and Bodie 1998, Russell et al. 2002). Yet, the self-organization of the wildlife

community of dome swamps within the four main land use categories, or along a gradient

of human land use intensity, is little understood. There have been many studies

comparing the pair wise wildlife composition amongst common land uses of a region

(e.g., Smith and Schaeffer 1992, Delis et al. 1996, Daily et al. 2001, Guerry and Hunter

2002, Renken et al. 2004); however, no published accounts could be found that

simultaneously explore the wildlife composition of agriculture, natural, silviculture, and

urban/residential. Similarly, studies have explored the biotic response to an implied or

qualitative gradient of anthropogenic land use intensity (e.g., Blair 1999, Chambers et al.









1999, Miller et al. 2003, Norris et al. 2003); however, rarely is the land use intensity

gradient in these studies a quantitatively based measure. After a review of the available

information Hart and Newman (1996) concluded that the impacts of human land use on

wildlife, especially amphibians, of isolated wetlands in Florida, should be a priority for

future research in Florida.

This dissertation explores several questions related to amphibian and avian species

composition in Florida. How similar are species assemblages within four common land

use types (agriculture, reference, silviculture, and urban)? Do indices of land use

intensity or wetland condition correspond with differences in species composition? What

environmental variables are significantly associated with amphibian and avian species

composition? Which species are indicative of specific land uses or levels of land use

intensity? Finally, are certain guilds (groups of species that share life history

characteristics) sensitive or tolerant to different land use types or levels of land use

intensity?

Study Systems

No part of Florida's 1,400,000 ha lies above 105m (mean sea level). A relatively

flat landscape and temperate to subtropical climate that averages approximately 1.35 to

1.54m of rainfall a year lead to the formation of numerous wetlands and water bodies. In

1780 approximately 54% of Florida was wetlands, but land development reduced the

proportion to 29% by 1980 (Dahl 1990). Of the remaining wetlands in Florida nearly one

third are forested freshwater wetlands with a large proportion typed as stillwater swamp

forests which consist primarily of cypress basin swamps and gum ponds (Ewel 1990).

The focal habitat type in this study was dome swamps which are a subset of

stillwater swamp forests. Dome swamps include cypress domes, cypress ponds, cypress









heads, cypress galls, gum ponds, bayheads, pine barren ponds, and Citronelle ponds

(Florida Natural Areas Inventory 1990). Dome swamps tend to be circular with the

tallest trees occupying the center and tree heights decrease towards the outer edges

(Clewell 1986). The surface water of dome swamps is hydrologically isolated from other

waterbodies and wetlands. The hydroperiod of dome swamps varies but most have

standing water periodically throughout the year (Clewell 1986, Heimburg 1986, Ewel

1990, Folkerts 1997). A general estimate of the hydroperiod for dome swamps is

between 200 and 300 days (Florida Natural Areas Inventory 1990). Variability in dome

swamp hydroperiod can result from differences in geomorphology, vegetation type, soil

type, ground water connectivity, local precipitation patterns, and human disturbance.

Typically, water levels are deeper in the center of dome swamps where peat

accumulations are greatest and decrease towards the periphery where mineral soils

dominate (Spangler et al. 1976, Ewel 1990). Maximum water depths are usually no more

than a meter (Monk and Brown 1965, Clewell 1986). The center of a dome may develop

a treeless pond surrounded by marsh vegetation giving the dome a doughnut like

appearance (Vernon 1947). Likewise, freshwater marsh vegetation or shrubs fringe many

domes where water level fluctuation and fire frequency are greatest.

Like hydroperiod, the fire frequency within dome swamp wetlands is partially

dependent on wetland type and the surrounding landscape composition. Cypress dome

swamps may be dependent on periodic fire where the outer edges may burn as frequently

as every 3-5 years and the interior every 100 to 150 years. Gum pond and bay wetlands

burn less frequently (Florida Natural Areas Inventory 1990).









Isolated basin wetlands have low productivity and growth rates compared to other

wetland habitats (Mitsch and Ewel 1979, Brown 1981). Productivity in dome swamps

may be influenced by both the available nutrients and hydroperiod (Brown 1981, Ewel

1990). Depression wetlands are primarily dependent on rainwater or groundwater

infiltration for recharge and have relatively small catchment basins; thus, wetland surface

waters tend to be oligotrophic (Brown 1981). Peaty soils aid in lowering the surface

water pH which further decreases nutrient availability to dome swamp vegetation (Monk

and Brown 1965, Ewel 1990). It is not uncommon to have pH levels in natural dome

swamp wetlands range between 3.5-4.5 (Brown 1981, Florida Natural Areas Inventory

1990, Warner and Dunson 1998).

Still water swamps are primarily classified by the dominant overstory vegetation.

For instance, cypress domes are dominated by pond cypress (Taxodium ascendens), gum

ponds by black gum (Nyssa biflora), and bayheads by three species of bay tree (Magnolia

virginiana, Perseapalustris, and Gordonia lasianthus). However, all three often share

vegetation components of the other habitat types (Monk and Brown 1965, Ewel 1990).

Other typical overstory species in dome swamps include Pinus elliottii, P. serotina, P.

taeda, and Acer rubrum. Typical midstory and shrub species include Ilex myrtifolia, I.

cassine, Lyonia lucida, Itea virginiana, Annona glabra, Myrica cerifera, Cephalanthus

occidentalis, Cyrilla racemiflora, and Vaccinium corymbosum (Monk and Brown 1965,

Florida Natural Areas Inventory 1990).

Flatwoods, clayhills, sandhills, and wet and dry prairies covered more then 60%

of Florida (Kautz et al. 1993) and are the typical unadulterated habitats surrounding dome

swamps (Clewell 1986, Abrahamson and Hartnett 1990). These systems have evolved to









periodic fire every one to ten years and frequently have Pinus spp. as the dominant

overstory vegetation (Florida Natural Areas Inventory 1990). In areas that receive

periodic fire, flatwoods, sandhills, and clayhills achieve an open savanna like habitat

structure and in the absence of fire succeed into hardwood forest (Engstrom et al. 1984,

Clewell 1986, Wolfe et al. 1988). Nutrient cycling and availability varies by upland

habitat type, but primarily these systems are oligotrophic with relatively low productivity

(Art and Marks 1971).

For at least 5,000 years much of Florida was pine savanna with interspersed

wetlands (Watts 1971, Clewell 1986, Watts et al. 1992). It is only within recent times

that many of the habitats and driving energies that much of Florida's biodiversity has

adapted to are experiencing rapid human induced alteration.

Wildlife Response to Land Use Type

"Reference" Habitats

Throughout the rest of the text "reference" refers to undeveloped habitats that

have received relatively low anthropogenically derived inputs and alterations, and are the

least impacted sites in this study. The reference sites are reasonable examples of

Florida's native communities (e.g., pine flatwoods, sandhills, dry prairies, cypress

domes). There is a set of amphibian and avian species that are commonly found in

Florida reference habitats.

Florida has 53 native amphibians (Enge 1997) and three established exotic anuran

species. Depression wetlands are important habitats for many of Florida's amphibians

and a subset requires isolated depressional wetlands for larval development. There are 14

amphibian species that require, and 21 species that will utilize, small isolated

depressional wetlands as breeding sites in Florida (Moler and Franz 1987).









There have been 196 confirmed breeding bird species in Florida (FFWC 2003).

More specifically, Rowse (1980) recorded 52 summer resident bird species in north

Florida flatwoods, and Workman (1996) recorded 14 summer bird species in 9 north

Florida cypress domes. No avian species in Florida is dependent on small isolated

wetlands; however, facultative use is common among several species (Hart and Newman

1995).

At any one location in Florida there is a regional pool of species available to

respond to changes in the landscape. Regional species richness increases with the

increasing number of habitat types, and this has been suggested as a reason why there is a

general pattern of decreasing species richness for amphibians and birds from north to

south throughout the Florida peninsula (Means and Simberloff 1987, Engstrom 1993).

For example 19 out of 26 of Florida's salamander species are found only in the

Panhandle and/or north Florida (Enge 1997). This suggests there may be more species

sensitive to development in north Florida simply due to a larger species pool. At the

same time South Florida's warmer annual temperatures and lower frequency of days

reaching 0 C may allow for the establishment of more temperature sensitive exotic

species capable of colonizing both disturbed and natural habitats (Simberloff et al. 1997).

Although there is an attempt to utilize "natural" habitats (e.g., parks and

preserves) as benchmarks of reference condition in studies of ecosystem change even the

benchmarks themselves are experiencing, and often have experienced, varying degrees of

human induced alterations that have the ability to influence biotic composition (e.g.,

Rooney et al. 2004).









The extinction and extirpation of species is the most obvious alteration to a

region's biotic composition. Some wildlife species now missing from Florida's systems

due to direct or indirect effects of humans include the passenger pigeon (Ectopistes

migratorius), Carolina parakeet (Conuropsis carolinensis), ivory-billed woodpecker

(Campephilusprincipalis), and the Florida red wolf (Canis rufusfloridanus) (Kautz

1993). Similarly, the geographical ranges of numerous species have been reduced,

extirpating them from habitats where they once interacted (e.g., Florida panther, Felis

concolor coryi, from most of Florida). Without replacement these species' role in

systems is lost from both developed and reference landscapes. Several extinct or

extirpated species are top carnivores whose absence from their hierarchical positions has

allowed a shift in food chain dynamics where lower level carnivores and omnivore

populations increase with concomitant potential to affect lower levels of the food chain

(e.g., Crooks and Soule 1999).

The vegetation structure of today's reference habitats may be very different than

in the past. Virtually all of Florida's old growth timber was harvested between 1870 and

1925 (Clewell 1986). Most of today's preserves, national forests, national wildlife

refuges, and military bases were clear-cut immediately before their designation (Clewell

1986). Thus the average age of the oldest overstory trees in nearly all of Florida's natural

areas are no more than 80 years and the trees in the majority of Florida's forested systems

are probably much younger (Wolfe et al. 1988). The effect on forest wildlife of the

decrease in the density of older large trees and the loss of their distinct structural

characteristics is largely unknown (Harris and Vickers 1984). For instance, cavities

occur more often in older and larger trees, and populations of cavity dependent species









may be very different from what they were before widespread human alteration of

Florida's forested habitats.

Timing and natural procession of fire is another forcing function in Florida's

landscapes that has also been significantly altered by human actions. Starting in the

1920s large scale programs to stop natural forest fires and prescribed burns were enacted

by federal officials (Clewell 1986). Many of Florida's natural systems are dependent on

fire for their maintenance. Obstructions such as roads, canals, and open fields inhibit

fires from covering the extent that traditional fires would have. Even the timing of

landscape fires has largely changed in recent times where winter burns are more frequent.

The historical fire season in Florida was most likely between late April and June when

habitats were driest and the advent of summer convective rainstorms brought frequent

lightning strikes (Robbins and Myers 1992). Several pineland plants require summer

fires to stimulate flowering (Wolfe et al. 1988, Robbins and Myers 1992).

Many of Florida's native species have adapted to habitats that experience frequent

ground fires that help to maintain an open savanna like habitat with numerous grasses and

forbs (Engstrom et al. 1984). For example, the size, density, and productivity of red-

cockaded woodpecker (an endangered southeastern pine savanna specialist) social units

were highly correlated with ground cover composition and natural pine regeneration, both

of which were surrogate measures of local fire history (James et al. 1997). Furthermore,

red-cockaded woodpeckers laid larger clutches the year after their territories had been

burned (James et al. 1997). Engstrom et al. (1984) documented the gains and losses of

bird species over a 15 year period of fire suppression in a north Florida pinewoods.

Notably, loggerhead shrike, Bachman's sparrow and blue grosbeak disappeared within a









few years of fire suppression and species that favor brushy areas dominated, common

yellowthroat, indigo bunting, rufous-sided towhee, and white-eyed vireo.

There are relatively few recent baseline studies of amphibian and avian

assemblages in reference Florida landscapes; however, mounting evidence suggests some

species decline or even disappear as human inputs to habitats increase (e.g., Blair 1996)

which implies a subset of species exists that is indicative of the lowest intensity land uses

in Florida.

Agriculture

Approximately 1000 years ago Apalachee and Timucuan Indians maintained

croplands in Florida, and in parts of the fertile clayhills of the northern panhandle their

fields stretched "as far as the eye could reach" (in Clewell 1986). Spanish colonizers

brought cattle, hogs, and other grazing animals and for almost 300 years much of Florida

was open rangeland where the only fences were around croplands (Clewell 1986,

Daugherty 1989, Ewel 1990). The amount of land area influenced by grazing was

reduced with the advent of the 1940s fence law requiring pastures to have fences, and

rangeland started becoming less extensive and more intensive (Clewell 1986). During

this same time period cropland intensification increased with the advent of widespread

manufactured fertilizer use and essentially the traditional fallow field rotation system

ceased (Wolfe et al. 1988).

High rates of land conversion to agriculture occurred from 1936 to 1987 when

agricultural land increased from approximately 17.5% to 29.7% of the state (Kautz 1993).

Today, the amount of agricultural land appears to have leveled off and may even be

decreasing (Economic Research Service 2002). Based on data from Vesterby and Krupa

(2001) approximately 36.5% or 5.1 million ha of Florida is cropland or grazed by









domestic animals. Cropland is a higher intensity land use and it comprises 29% of all

agricultural lands or 1.5 million ha. In Florida the most common types of agriculture are

rangeland, improved pasture, dairy, citrus, row crops, sugar cane, and greenhouse/nursery

(FASS 2005). Given the large percent of land area utilized for agriculture there are

surprisingly few studies addressing amphibian or avian use of Florida farmland

landscapes.

Amphibians and agricultural landscapes

Meshaka (1997) described the herpetofauna of a ranch in central Florida and

documented 19 amphibian species in various habitats. Many of the species utilized the

abundant ephemeral ponds on the property. Eastern spadefoot (Scaphiopus holbrookii)

was expected, but reported missing from the ranch. Hydrological manipulations of the

improved pasture which maintain water levels high during the dry season and lower than

normal during the wet season were stated as the best explanation for the absence of this

highly fossorial species. A comparison with a lower or higher intensity land use was not

made.

Folkerts (1997) commented on the relatively low occurrence of amphibians in

Citronelle ponds (abrupt topographical depressions within the Citronelle formation). The

lower than expected amphibian species richness was attributed to extensive disturbance

to the surrounding terrestrial habitat. Unlike many ephemeral isolated wetlands

Citronelle ponds are surrounded by relatively fertile sandy loam soils used extensively for

row cropping (Wolfe et al. 1988, Folkerts 1997). Amphibian species frequently

encountered in Citronelle ponds included Acris gryllus, Hyla cinerea, Rana catesbeiana,

Rana grylio, Pseudacris nigrita, Pseudacris crucifer, and Gastrophryne carolinensis

(Folkerts 1997).









In locations other than Florida, agricultural landscapes have been implicated in

altering amphibian species composition, richness, and abundance. For example, in

Australia, Jansen and Healey (2003) attributed observed decreases in frog communities,

species richness and some individual species populations to increased grazing intensity.

Knutson et al. (2004) did not compare natural to agricultural wetlands but did find frog

species richness and reproductive success was similar in Minnesota ponds and wetlands

surrounded by row crops and non-grazed pasture. Ponds used by livestock for water had

higher turbidity, higher phosphorus and lower amphibian reproductive success. Species

richness was highest in ponds with lower total nitrogen levels (Knutson et al. 2004). A

significantly greater number of two spadefoot toad species was found in isolated playa

wetlands surrounded by crops than in those surrounded by lightly grazed pasture.

Species richness and abundance for the remaining species did not differ by agricultural

land use type (Gray et al. 2004).

Amphibians exhibit species specific responses to agricultural landscapes that

depend on life history characteristics and landscape attributes. Kolozsvary and Swihart

(1999) studied amphibian use of forest patches and found American toads (Bufo

americanus) and gray treefrogs (Hyla chrysoscelis) were tolerant to most agricultural

landscapes, ranid frogs responded to the nearness of adjacent wetlands, redback

salamanders (Plethodon cinereus) were correlated with the amount of forest area, and

several species responded to wetland hydroperiod. Ray et al. (2002) found presence of

common toads (Bufo bufo) declined with cultivated field area and alpine newt (Triturus

alpestris) presence declined with vineyard area surrounding ponds.









American toads (Bufo americanus) and northern leopard frogs (Ranapipiens) were

negatively associated with the amount of forested area within 1 km of small (<0.5ha)

isolated wetlands in a Maine agricultural landscape. Five other amphibian species were

positively associated with forested area and two species were most likely to occupy

ponds with adjacent forest (Guerry and Hunter 2002). The presence of four Ohio

amphibians had a positive association with the amount of forest cover within 200m of

breeding ponds (Porej et al. 2004). The remaining species had negative responses to

cumulative lengths of paved roads within 1km of the breeding ponds and the distance to

the nearest five wetlands (e.g., Notophthalmus viridescens)(Porej et al. 2004).

Herbicide, pesticide, and fertilizer use is prevalent in agricultural landscapes.

Nutrients from fertilizers and animal waste have the potential to alter wetland

biogeochemical processes and species composition. Reproductive success of the

northwestern salamander (Ambystoma gracile) and northern red-legged frog (Rana

aurora) were found to be significantly lower in wetlands receiving agricultural runoff (de

Solla et al. 2002). High ammonia concentrations and high biochemical oxygen demand

are suspected in causing the observed difference in hatching success. Algal mats

frequently form in eutrophic waters and Palis (1997) observed that Flatwoods

salamanders (Ambystoma cingulatum) are not found in wetlands with excessive amounts

of algae. Houlahan and Findlay (2003) found the presence/absence of 12 of 13

amphibian species was negatively correlated to total nitrogen levels in 74 Ontario

wetlands. Nitrates and nitrogen fertilizers at ecologically relevant levels are known to

have toxic and lethal effects on amphibian species (Hecnar 1995, Adolfo et al. 1999).









In a study in the Prairie Pothole wetlands of Canada, 288,000 wetlands

(approximately 17% of the wetlands in the area) were estimated to exceed Canadian

pesticide guideline levels for the protection of aquatic life (Donald et al. 1999). Berrill et

al. (1998) found lethal and sublethal affects on three larval anuran species exposed to

relatively low doses of a common agricultural pesticide, endosulfan. American toad

(Bufo americanus) tadpoles were more tolerant to endosulfan exposure than wood and

green frogs (Rana sylvatica andR. clamitans). Bridges and Semlitsch (2000) also found

species specific lethal and sublethal responses as well as population specific responses

within Rana sphenocephala to Carbaryl, a common crop and garden insecticide. Finally,

larval Xenopus laevis exposed to ecologically relevant levels (0.01 to 40 ppb) of atrazine,

the most widely used herbicide in the U.S., exhibited hermaphroditism and demasculation

(Hayes et al. 2002).

Not only do the active ingredients in pesticides have an apparent effect on

amphibians but the surfactants used in common pesticides were found to have a narcotic

effect. For instance, Mann and Bidwell (2001) found that after exposure to surfactants

tadpole feeding activity was reduced and the effects were compounded under low

dissolved oxygen conditions.

Birds and agricultural landscapes

A comparison of avian richness on four agricultural land use types in west-central

Florida revealed a relationship between bird species richness and land use type (Cutright

1981). Species richness was greatest in native grazed forest habitats (included dome

wetlands), followed by unimproved pasture, improved pasture, and cropland supported

the lowest number (Cutright 1981). The most abundant species during the summer

season in the pasture habitats were eastern meadowlark, cattle egret, and northern









bobwhite quail, and in the cropland mourning dove and red-winged blackbird were the

most common. Shrub dependent species (e.g., Carolina wren and white-eyed vireo) and

vultures dominated the grazed forest habitats. Hirth and Marion (1979) found the eastern

meadowlark and northern bobwhite quail were the most common summer residents in

native rangeland (grazed flatwoods) of south Florida, and summer granivore densities

were twice that of insectivores.

A single species study by Morrison and Humphrey (2001) found higher densities

and higher nesting success of the federally threatened crested caracara (Caracara

cheriway) within Florida cattle ranches than on lands managed as natural areas. The

loggerhead shrike is often positively associated with lower intensity agricultural lands

and other open areas, and population declines in Florida and elsewhere have been

attributed to decreases in low intensity agricultural lands since the 1940s (Cade and

Woods 1997).

Several studies conducted in tropical areas indicate avian species richness is higher

in or near remnant forest patches than in the surrounding agricultural matrix (Daily et al.

2001, Luck and Daily 2003, Naidoo 2004). Avian species composition is also distinct in

tropical regions between lower and higher intensity agricultural land uses (Daily et al.

2001, Naidoo 2004).

DesGranges and Boutin (1996) analyzed avian population trends along Breeding

Bird Survey (BBS) routes in Quebec and found species experiencing population declines

were associated with lower intensity agriculture (e.g., pastureland) while increasing

species were associated with the recent increases of more intensive agriculture (e.g., cash

crops). In England population declines of farmland birds were associated with a trend in









increasing agricultural intensity (Chamberlain et al. 2000). In Iowa bird species richness

declined on a continuum of landscapes types from forest to row-crop monocultures (Best

et al. 1995). In North Dakota the best indicator of duck abundance in shallow wetlands

was the proportion of upland area comprised of agricultural land (Austin et al. 2001).

Three levels of grazing intensity were associated with changes in avian species

composition and richness in Utah. Metrics of land use intensity were created using

species dominance and richness. However, these metrics were only significant for the

high impact sites (Bradford 1998). Finally, Beecher et al. (2002) reported higher bird

species richness on organic farmland than on nonorganic farmland, 54 and 39 species

respectively. They attributed the higher bird species richness to greater foraging

opportunities provided by the increased biomass of non-crop vegetation at sites that were

not treated with herbicide.

Ford et al. (2001) reviewed why many species of birds have declined in the

eucalyptus woodlands throughout the agricultural zone of Australia. Although species

from several feeding guilds have declined, insectivores dependent on native vegetation

experienced the biggest declines. Hypotheses of mechanisms responsible for declines in

some species included poor dispersal capabilities between habitat patches, competition

with increasing populations of larger generalist species, habitat specificity and the

inability to adapt to vegetation changes in fragments, increased nest predation by

mesopredators in habitat fragments, and degradation of food resources (Ford et al. 2001).

Declining and increasing avian species populations to changes in land use and

varying avian species richness by land use types suggest there are avian species

indicative of land use intensity within agricultural habitats. For instance, brown-headed









cowbird, eastern meadowlark and Brewer's sparrow were associated with lower intensity

agriculture and ring billed gull, rock dove and homed lark with higher intensity

agriculture (DesGranges and Boutin 1996, Bradford et al. 1998). No studies were found

that utilized birds as statistically significant indicators of agricultural land use or

agricultural land use intensities.

Silviculture

Approximately 43% of Florida is managed for commercial forestry and more then

27% of that area is planted in dense stands of Pinus elliottii, P. taeda or P. clausa (Kautz

1993). A mature stand of planted pine superficially appears to be the least altered of the

anthropogenic land-use types discussed in this paper, and perhaps silvicultural areas are

less intensively managed than both agricultural and urban areas. However, mechanical

site preparation, fertilizer augmentation, herbicide application, relatively short rotation

cycles, and densely stocked stands cause a significant shift in the composition and

structure of forest vegetation (Clewell 1986). Likewise, the conversion from native

landscapes to pine plantations appears to be associated with a shift in vertebrate species

composition.

Amphibians and silvicultural landscapes

The flatwoods salamander (Ambystoma cingulatum) may be sensitive to silviculture

practices in Florida. Means et al. (1996) observed a drastic decline in a large A.

cingulatum population over a 22-year period. Although unable to demonstrate a direct

link, they attributed the decline to the silvicultural practices occurring in the surrounding

landscape. The direct effect of mechanical site preparation (e.g., roller chopping, disking,

and bedding) on fossorial species is poorly understood. Wolfe et al. (1988) stated the

amphibians found in pine plantations, such as oak toad (Bufo quercicus) and pinewoods









treefrog (Hylafemoralis), entered from adjacent communities following large scale

silvicultural related disturbances. Similarly, a recovery of flatwoods herpetofauna

populations three years after a north Florida clear-cut was attributed to interspersed

cypress dome habitats that provided refugia during the early stages of forest succession

(Enge and Marion 1986).

Ephemeral wetlands in silvicultural areas may be important refugia and habitat for

several amphibian species. Russell et al. (2002) recorded 20 amphibian species utilizing

edge habitat of five ephemeral wetlands embedded in a South Carolina silvicultural

landscape, while in the adjacent uplands only 8 species were recorded. Amphibian use of

the wetlands was negatively correlated to the amount of hardwoods in the uplands and the

distance to the nearest wetland, and positively correlated to the conifer basal area

surrounding each wetland (Russell et al. 2002).

A common practice in flatwoods managed for silviculture is to ditch the isolated

wetlands to increase connectivity and flow. Furthermore, frequently wetlands are

encircled with a ditch to retard fire. As a result, wetland hydroperiods are altered. On

lands in north Florida used for timber production, Vickers et al. (1985) found nearly 4

times as many terrestrial reptiles and toads in ditched cypress ponds and 1.4 times more

frogs and salamanders in unditched cypress ponds. Furthermore, ditching can connect

once formerly isolated wetlands to sources of predatory fish (Babbitt and Tanner 2000).

The species specific response of amphibians to fish predation influences the amphibian

species composition of wetlands (Kats et al. 1988).

Herbicides and fertilizers are also potential anthropogenically induced stressors

within silvicultural landscapes. Intensively managed Florida pine plantations often have









herbicide and fertilizer applied two times during each stand rotation. Hatch et al. (2001)

observed avoidance behavior in several western amphibians exposed to paper towels

dosed with urea (a common forest fertilizer). However, the amphibians did not avoid

soils dosed with urea, and exposed individuals experienced higher mortality and lower

feeding rates than controls.

An insecticide, fenitrothion, used to control spruce budworm in boreal forests

caused paralysis in several tadpole species. Bullfrog (R. catesbeiana), green frog (R.

clamitans), and spotted salamander (Ambystoma maculatum) were the most sensitive

followed by the wood frog (R. sylvatica) and leopard frog (R. pipiens). The American

toad tadpoles (B. americanus) were the most resilient (Berrill et al. 1995). Similarly

Berrill et al. (1994) found three tadpole species died or were paralyzed upon exposure to

a herbicide, triclopyr, and an insecticide, fenitrothion, and the order of increasing

sensitivity was R. catesbeiana, R. clamitans, and R. pipiens.

Florida pine plantations are generally clear-cut on a 25-30 year rotation cycle

(Clewell 1986). Several researchers have documented differences in the amphibian

community in clear-cut areas versus forested areas. Enge (1984) found that amphibian

abundance was reduced 10-fold in North Florida clear-cuts while amphibian species

richness was not affected. Major reductions in Scaphiopus holbrookii, Gastrophryne

carolinensis and Rana sphenocephala were observed.

North Carolina salamander abundance was five times greater and species richness

two times greater in mature forests (50-70 years) than in clear-cuts (Petranka et al. 1993).

Amphibian response to clear-cuts appears to be species specific where some species are

sensitive and others tolerant of tree harvesting (e.g., deMaynadier and Hunter 1998).









Intuitively, arboreal species are impacted by the removal of trees, but clear-cutting also

reduces relative humidity and increases insolation and evaporative water loss from the

soil (Raymond and Hardy 1991). The "drying out" of the habitat may partly explain the

observed declines in amphibian abundance and richness within clear-cut areas.

In most industrial and commercial forests in Florida clear-cut areas are immediately

replanted with an even aged, densely stocked monoculture. Even aged stands of trees

have been reported to have lower local abundances of amphibians than uneven aged

mixed forest; however, a reduction in species richness or composition was not detected

(Renken et al. 2003). Mature even aged stands have higher densities of redback

salamanders (Plethodon cinereus) than regenerating and sapling even aged stands

(DeGraaf and Yamasaki 2002).

Amphibian species richness and composition among different isolated wetlands

within managed forests of a region appear to be consistent (Russell et al. 2002) that

suggests amphibian composition may be used as an indicator of this land use.

Birds and silvicultural landscapes

Avian abundance and species richness is higher in Florida silvicultural landscapes

that contain forested wetlands (Marion and O'Meara 1982, Harris and Vickers 1984).

Furthermore, the ecotone of small cypress wetlands and clear-cuts supports higher

breeding bird abundances and species richness than the ecotone with planted pine stands

(Harris and Vickers 1984). The vegetation height heterogeneity of cypress domes is the

suspected reason why bird density and diversity are greater than in the surrounding pine

plantations or clear cuts (Marion and O'Meara 1982). Cypress domes also have more

cavities than pine stands and approximately 30% of the flatwoods/cypress dome avian

community are cavity nesting species (Rowse 1980, McComb et al. 1986). Finally,









forested wetlands within a pine plantation matrix may experience pulses of increased

avian richness by providing refugia for forest dependent species after adjacent pines are

harvested (Marion and O'Meara 1982).

Comparisons of north Florida pine plantations divided into three levels of site

preparation intensities indicate that avian species richness and abundance are highest in

the low and intermediate intensity stands (Harris et al. 1975). Furthermore, all plantation

sites (n=9) had lower avian richness and abundance than mature stands of longleaf (Pinus

palustris) or slash (Pinus elliottii) pine. Finally, mature longleaf pine stands had more

species than mature slash pine stands (Harris et al. 1975).

In Oregon, Chambers et al. (1999) reported the avian response to three intensities

of silvicultural treatments and uncut control stands. The avian community of the low

intensity treatments was most similar to the control stands. Avian community

composition in the moderate and highest intensity treatments was significantly altered.

The number of species that declined was approximately equal to the number that

increased in the moderate and highest intensity treatments. Similarly, Tittler et al. (2001)

found the abundance of 10 species increased, 7 decreased, and 10 remained the same in

silvicultural treatments relative to controls in Alberta, Canada. Four species were only

found within the silviculture treatments. Keller et al. (2003) found species richness

increased 2-6 years after clear-cutting then decreased between 7-25 years after cutting

and then increased again after 25 years. They concluded avian guilds respond to

successional stages of forest structure and productivity (Keller et al. 2003).

Estades and Temple (1999) included the landscape mosaic in exploring the effect

of habitat fragmentation on Chilean avian communities in managed forest systems.









Fragmentation effects were species specific. The avian community was significantly

correlated to the vegetation adjacent to the study forest fragments. For example, the

abundance of cavity nesting species was positively correlated to the proximity of native

forest patches.

Snags are often removed or at very low densities in southeastern pine plantation

forests. A habitat manipulation experimental study in loblolly pine forests (P. taeda)

found avian abundance, richness, and diversity, density of woodpecker territories, density

secondary cavity nesting species, and the density of Neotropical migrants to be greater in

plots that had snags than those where snags were removed (Lohr et al. 2002).

Several studies have investigated the response of avian guilds to silviculture. For

instance, ground nesters, cavity nesters, flycatchers, and long distance migrants were

negatively associated with monotypic spruce plantations in Norway (Hausner et al.

2003). Insectivorous birds consistently exhibit decreases in selectively logged forests

relative to primary or unlogged forests (e.g., Thiollay 1992, Taylor and Haseler 1995,

Mason 1996, Stratford and Stouffer 1999). More specifically, sallying and leaf gleaning

insectivores have significantly lower species richness in selectively logged forests while

bark gleaning insectivores, frugivores and omnivores increase (Owiunji and Plumptre

1998). Finally, ground nesting species in northern Minnesota forests were most

susceptible to nest-predation within 100 meters of clear-cut areas, so much so, that for

one species the edge habitat was a population sink (Manolis et al. 2003).

Urban/Residential/Roadways

The fastest growing land use in Florida is urban and residential development. From

1949 to 1987 urban land area increased almost 500% (Kautz 1993). Currently 13.4 % of









Florida's land area is composed of roads or urban/residential developments (Vesterby and

Krupa 2001).

Amphibians and urban landscapes

Very few studies explore amphibian species richness, composition, and or

abundance in urban landscapes. However a particularly relevant study by Delis et al.

(1996) compared anuran species richness and abundance within wetlands surrounded by

residential housing with wetlands surrounded by undeveloped land near Tampa, Florida.

Species richness and abundances of anurans were greater in the wetlands surrounded by

undeveloped land. Furthermore, anuran species richness and composition in the wetlands

surrounded by undeveloped land were more similar to a 1974 amphibian study conducted

in the same area prior to the residential development. Delis et al. (1996) claimed that

four species were sensitive to development because they were only found in the wetlands

surrounded by undeveloped land, Bufo quercicus, Scaphiopus holbrookii, Hylafemoralis,

and H. gratiosa. Three species, Rana sphenocephala, R. grylio, and R. catesbeiana, had

higher abundances in the wetlands surrounded by development. Changes to both the

upland and wetland habitats in the urban environment were thought to be responsible for

the loss of anuran species at the developed sites (Delis et al. 1996).

Life history traits of the extirpated (Sensitive) and the relatively more common

(Tolerant) species may help to explain potential causal mechanisms for the observed

differences. The four species Delis et al. (1996) found extirpated from wetlands

surrounded by development were fossorial (Bufo quercicus, Scaphiopus holbrooki, and

Hyla gratiosa) and/or arboreal (Hylafemoralis and H. gratiosa). Changes in the

surrounding terrestrial habitat may have reduced the available burrowing substrate for

fossorial anurans. In a burrowing experiment, adult Scaphiopus holbrookii were unable









to burrow into grass sod, and were much slower burrowing into landscaping gravel than

sand. Juvenile Scaphiopus holbrookii did not burrow into common residential substrates

such as sod, gravel, and saturated soils (Jansen et al. 2001). Reduction in the amount of

area affording refugia has the potential to increase this species exposure to predators and

the elements.

A study in Italy of 84 wetlands within developed landscapes found that the

wetlands with the highest amphibian species richness were free of fish, received high

levels of insolation, and were near other wetlands that had high amphibian species

richness (Ficetola and De Bernardi 2004). The most common species were pollution

tolerant, did not require large patches of terrestrial habitat, and were capable of moving

through human dominated landscapes by using hedgerows and canals. The most

sensitive species appeared to be influenced by a combination of wetland isolation and

habitat alteration. In general the sensitive species were dependent on terrestrial habitats

for part of their life cycle, highly susceptible to road traffic, and required water bodies

free from fish predation (Ficetola and De Bernardi 2004).

Along a rural to urban gradient, a 10 km long 2 km wide transect in Connecticut,

two amphibian species, Rana sylvatica and Ambystoma maculatum, were found only in

landscapes above 30% forest cover, Notophthalmus viridescens, persisted only in

landscapes above 50%, forest cover while two species, Plethodon cinereus and

Pseudacris crucifer, were found along the entire urban gradient (Gibbs 1998). The

species dispersal abilities were inversely correlated to their persistence in fragmented

habitats. That is, the most sedentary species persisted in habitat fragments and were

found in the highest densities (Gibbs 1998).









There are several potential mechanisms influencing amphibian species abundance

and composition within the urban environment. Wetlands that remain in urban areas are

often incorporated into the cities' stormwater systems or utilized for tertiary wastewater

treatment. Furthermore, very little of the native terrestrial habitat surrounding wetlands is

allowed to persist, and road density, herbicide and fertilizer use all increase.

Urbanized watersheds as well as wetlands utilized for stormwater control

experience highly fluctuating water depths, which have been correlated to the amount of

impervious surface in catchment basins (Richter and Azous 1995). Azous and Horner

(1997) captured fewer amphibians in wetlands where water level fluctuation exceeded 20

cm. Egg masses attached to vegetation were suspected to be incapable of withstanding

extreme water depth fluctuations.

Heavy metals bioaccumulate in organisms using stormwater wetlands. Three

species of fish (Lepomis microlophus, L. macrochirus, and Micropterus salmoides)

collected from stormwater ponds in the Orlando, Florida area had significantly higher

concentrations of heavy metals (Ag, Cd, Ni, Cu, Pb and Zn) than fish caught in natural

lakes (Campbell 1995). Furthermore, some amphibians experience increased embryonic

and larval mortality rates, slower developmental rates, reduced feeding rates, and lower

breeding success when exposed to different heavy metals. These effects may be

compounded under low pH conditions (Home and Dunson 1994, 1995a,b,c, Rowe et al.

1996).

Large amounts of water are used each day in urban areas. The resulting

wastewater is often partially treated and then released into the environment. Ephemeral

ponds within effluent-irrigated fields in Pennsylvania are used to treat wastewater high in









chlorine, boron and nitrate. High boron concentrations were found to reduce hatching

success of B. americanus and cause deformities in R. sylvatica, A. jeffersonianum, and A.

maculatum (Laposata and Dunson 1998). When compared to natural temporary ponds

the wastewater-irrigated pond amphibians had significantly fewer egg masses, lower

hatching success and lower larval survival (Laposata and Dunson 2001). In Florida,

Jetter and Harris (1976) noted initially high anuran use of cypress depression wetlands

receiving wastewater. However, anuran use nearly stopped as dissolved oxygen levels

dropped with increasing inputs of waste.

Birds and urban landscapes

There have been numerous studies of avian composition, abundance, and species

richness in urban environments throughout the world. However, studies of avian

diversity in urban environments within Florida or the Southeastern U.S. are limited.

Woolfenden and Rohwer (1969) described the breeding birds within three study

plots located in the suburbs of St. Petersburg and Gulfport, Florida. Older residential

suburbs had three times as many breeding pairs than newer residential suburbs, 600 and

200 pairs per 100 acres respectively. The proportion of vegetative cover, a potential

mechanism explaining the observed trend, in the new suburb was estimated to be 1.4%

while that in the more mature suburbs were 28.3 and 10.3%. The breeding bird density in

the mature Pinellas county suburbs were also twice the densities found in grassland,

marsh, brush and scrub, oak hickory forests, beech maple forest, and mixed coniferous-

deciduous forests (Udvardy 1957). A total of eleven species were recorded breeding in

the suburbs and the house sparrow, mourning dove, blue jay and northern mockingbird

comprised 90% of all sightings (Woolfenden and Rohwer 1969). Based on earlier studies

of native habitats that had been replaced by suburban development Woolfenden and









Rohwer (1969) listed 10 species that respond negatively to development; red-tailed hawk,

red-shouldered hawk, bald eagle, yellow-billed cuckoo, chuck-will's-widow, red-

cockaded woodpecker, scrub jay, brown-headed nuthatch, pine warbler and Bachman's

sparrow.

In a comparison of avian use of two forested riparian corridors, in a north Florida

urban area and a state preserve, community and species specific differences were

detected (Smith and Schaefer 1992). Densities of five insectivores were lower in the

riparian forest surrounded by urban development while the density of four omnivores and

an insectivore increased (Smith and Schaefer 1992).

In general, bird abundance and biomass increases and species richness and

community evenness decreases in an urban matrix relative to less fragmented natural

habitat areas (Woolfenden and Rohwer 1969, Emlen 1974, Mills et al. 1989, Blair 1996,

Clergeau et al. 1998, Sorace 2002, Traut and Hostetler 2004). There are many potential

mechanisms responsible for this trend. Human supplied feeding stations were cited as a

likely reason why Emlen (1974) recorded a 65-fold increase in herbivorous birds'

biomass in Tucson, Arizona relative to the surrounding native habitat. Species that do

not maintain feeding territories and those species that utilize human structures for nesting

also tend to be more successful in urban environments (Emlen 1974).

Studies focusing on the abundance of birds in urban environments have found

that the population response is often species specific and varies with the habitat qualities

within the urban environment (Emlen 1974, Williamson and DeGraaf 1981, Mancke and

Gavin 2000, Hostetler and Knowles-Yanez 2003). The population densities of 21 of 36

bird species were affected by the proximity of buildings to study woodlots in









Pennsylvania. The density of 10 species increased while 11 species decreased in

woodlots adjacent to buildings (Mancke and Gavin 2000). Hostetler and Knowles-Yanez

(2003) surveyed birds in older residential, newer residential, parkland, and golf courses in

Phoenix, AZ, and found abundance of 4 of the 26 species encountered were significantly

correlated with the amount of specific urban land use type. For instance, abundance of

killdeer varied with golf course area, house sparrow with single lot residential and

mourning dove with medium density residential. The remaining species that did not

show a correlation with any specific land use type were thought to be associated with the

unique vegetation characteristics found within the different study areas (Hostetler and

Knowles-Yanez 2003).

Avian species that share life history characteristics (i.e., guilds) have

demonstrated population responses to urban land uses (e.g., Flather and Sauer 1996).

The density of houses within 100 m of forest remnants has been shown to have a strong

affect on Neotropical migrant birds while short-distance migrants and permanent

residents showed no observable trend. Furthermore, urban 25 ha woodlots had a smaller

Neotropical migrant community than 4 ha woodlots surrounded by lower intensity land

uses but without nearby houses (Friesen et al. 1995).

In New England, Kluza et al. (2000) also found Neotropical migrants and forest

interior species responded negatively to the density of housing in a forest matrix. More

specifically, abundances of ground nesting species were greatest in forest areas with the

lowest housing density while abundances of blue jays, a reported nest predator, were

greatest in areas of moderate housing densities. Vegetation characteristics were not

significantly different between areas (Kluza et al. 2000). Nest predators were suspected









to be responsible for the observed trends, and nest predation rate may increase as the land

use intensity contrast between remnant forest patches and adjacent land uses increases

(e.g., Hanski et al. 1996, Suarez et al. 1997).

Mills et al. (1989) found the densities and richness of native territorial birds were

strongly correlated to the volume of native vegetation in urban areas and to a lesser extent

a negative correlation with housing density. Native territorial birds were not correlated

with exotic vegetation volume. However, exotic and nonterritorial birds were positively

correlated with exotic vegetation volume and the amount of lawn area (Mills et al. 1989).

There is an assumed gradient in urban land use intensity from rural areas to the

center of cities. Several studies have utilized the rural to urban gradient to examine the

role humans have on avian communities (McDonnell and Pickett 1990). The urban

gradient approach also enables researchers to identify, often species-specific, thresholds

of human development on bird communities (Miller et al. 2001). Studies conducted in

different regions of the world exploring the rural to urban gradient have indicated there

are avian species that are urban avoiders, urban adapters, and urban exploiters

(McKinney 2002).

Blair (1996, 2001) ranked urban land areas from the least altered to the most

altered by humans; biological preserve, recreational area, golf course, residential

neighborhood, office park and business district. One or more bird species reached their

maximum density in each of the urban land uses sampled (Blair 1996). Native species

occurred predominately in the lower intensity land uses while generalists and exotic

species were most common in the business districts. The avian community similarity was

much higher between the most heavily developed sites in Ohio and California then the









similarity between native habitats in the two ecoregions (Blair 2001). Species richness,

diversity, and biomass were greatest at sites of intermediate development (Blair 1996,

2001).

The community similarity between wintering birds in five town centers (30%) of

Finland was lower than the similarity between apartment complexes (54%) and single

family residential areas (54%)(Jokimaki and Kalsanlahti-Jokimaki 2003). A threshold

was detected in the impact on the avian community in towns between 35,000 and 105,000

residents, suggesting the regional land use context has an affect on avian species

composition within urban landscapes. Finally, along with city size and urban land use

type, the local habitat structure is thought to be especially important in shaping avian

community structure (Jokimaki and Kalsanlahti-Jokimaki 2003).

Crooks et al. (2004) compared avian abundance and richness in native habitat,

fragments, and urban transects in southern California. Fragments had the highest species

richness and abundance. The significant vegetation differences between native and urban

landscapes were suspected to be responsible for the trends in avian populations.

Furthermore, they detected species indicative of either urban or core habitat, and many of

the urban species were also encountered by Blair (2001) in Ohio and northern California.

Different bird species reached peak abundance within forest fragments in Seattle

surrounded by three different levels of urbanization (urban, suburban, and exurban)

(Donnelly and Marzluff 2004). Most species associated with native forest habitat

remained in fragments greater than 42 ha while urban dependent species were associated

with fragments surrounded by greater than 40% urban land cover. Urban dependent









species were positively correlated and native bird species were negatively correlated with

the amount exotic forb and shrub cover (Donnelly and Marzluff 2004).

Avian species composition changed with the number of native trees and shrubs,

ground cover, and tree density within Colorado riparian lowland areas along a rural to

urban gradient (Miller et al. 2003). Settlement intensity was negatively associated with

the migrant and low nesting species, while resident and cavity nesting species were

positively associated urbanization (Miller et al. 2003). The species richness and density

of birds in riparian forests of California along an urban gradient decreased with

increasing number of bridges within 500m, decreasing distance to nearest building, and

decreasing volume of native vegetation (Rottenborn 1999). The avian species

composition was influenced by the same variables with the addition of the proportion of

building and pavement surface cover within 500m. Foliage and bark gleaning

insectivores, cavity nesters, and ground nesting species all showed significant declines

with increasing urbanization. Finally, a set of tolerant and sensitive species to

urbanization was identified (Rottenborn 1999).

Significantly different bird communities were also found in four urban habitats

along a development continuum in Australia. Notably, insectivorous and nectivorous

bird species declined along the gradient from parks with native vegetation, to residential

areas with native vegetation, to residential areas with exotic vegetation to newly

developed areas denuded of most vegetation (White et al. 2004).

Clergeau et al. (1998) observed a decrease in insectivorous species and an

increase in omnivorous species on two rural to urban gradient studies on different

continents. Two types of bird groups capable of utilizing urban environments were









identified. Omnivorous species adapted to food resources supplied in urban

environments and species that exploit resources found in the urban environment and their

usual habitat. Several of the urban adapted species, European starling, rock dove, and

house sparrow were found in cities in Canada and France (Clergeau et al. 1998).

A literature review by McKinney (2002) identified the traits of avian species

groups indicative of areas along the rural to urban gradient. Species predominately found

in urban centers, urban exploiters or synanthropes, are often dependent on human

resources, can feed on wind dispersed seeds or are omnivorous, and nest on rocky cliff

like surfaces, or are cavity nesters capable of inhabiting human structures. Urban

adapters are often considered edge species for their tendency to be located at the ecotone

of different habitats. Urban adapters are predominately granivores, omnivores or aerial

sweepers, and tree, shrub, or cavity nesters. Urban avoiders, often called forest interior

species, tend to be tree foraging insectivores, neotropical migrants and or ground nesting

species.

Roads

Roads and the vehicles that use them traverse through many different landscapes

and habitat types. The 3.8 million km of roads in the contiguous United States cover 1%

of the land area, but it has been estimated that the ecological effect spans over 22%

(Forman 2000). Roads have both direct and indirect affects on wildlife populations. The

direct affects include vehicular induced mortality and suitable wildlife habitat is lost with

the construction of roads. Indirect affects include modification of animal behavior,

habitat fragmentation, alteration of the surrounding chemical and physical environment,

corridors for the spread of exotics, and increased access by humans (see Trombulak and

Frissell 2000).









Road density within two km of wetlands in Canada was negatively correlated with

the wetland's plant, amphibian, reptile, and bird species richness (Findlay and Houlahan

1997). Furthermore, the correlation was stronger when older road densities were used

suggesting the full effect of road density on the species richness of amphibians and birds

in wetlands can take decades (Findlay and Bourdages 2000).

The probability of six different amphibian species being killed while attempting to

cross a Denmark road with a traffic load of 134 vehicles/hour was estimated to be 0.34 to

0.61 and on roads with 625 vehicles/hour the probability rose to 0.89 to 0.98 (Hels and

Buchwald 2001). For two of the species it was estimated that 10% of the population

within 250 m of the road was killed annually (Hels and Buchwald 2001). Fahrig et al.

(1995) found reduced amphibian abundance and density, and increased mortality with

increasing traffic intensity. Finally, the probability that a moor frog (Rana arvalis)

population is present in suitable habitat drops from 55% to 30% when the study area is

adjacent to a highway (Vos and Chardon 1998).

Road density is often correlated with other aspects or human development making

it difficult to determine the independent effect of roads on wildlife populations. When

the effects of isolated forest patch characteristics were corrected for, avian species

richness was lower in forest fragments within 2 km of a major highway (Brotons and

Herrando 2001). At the same time, forest species richness in forest fragments was

negatively correlated to the distance to roads with higher traffic loads while some

ubiquitous species showed a positive relationship with highways and residential

roadways (Brotons and Herrando 2001). Species richness and relative abundance of

birds in Big Bend National Park, Texas was negatively associated with distance to nearest









road, distance to nearest development, and the interaction between these two variables

(Gutzwiller and Barrow 2003). Finally, a significant effect of traffic intensity was found

on the avian population (reduced 12-54%) within 100 m of roads with 5,000 cars a day

and within 500 m for roads with 50,000 cars per day in agricultural landscapes without

any other apparent land use influences (Reijnen et al. 1996).

Studies on the Response of Amphibians and Birds to Multiple Land Use Types

Studies of the wildlife response to landscapes of varying and multiple land use

intensities indicate biodiversity is primarily affected by the proportion of land use types

and distances between land use types of different intensities. There appears to be a

wildlife compositional gradient that roughly corresponds to a land use intensity gradient.

Wildlife response to land use type is species specific, and often an individual species

response is consistent between studies of similar land uses. Where species response, as

well as measures of biodiversity, seems to differ between studies of similar land uses

(e.g., agriculture) the intensity of land use within the land use type is often the

explanatory difference. Finally, there is also evidence suggesting species that share

certain life history characteristics respond in a similar manner to habitat fragmentation,

land use types and intensities (Whitcomb et al. 1981, O'Connell et al. 2000, Rodewald

and Yahner 2001, Norris et al. 2003).

Anuran species richness and abundance within Midwestern landscapes was

negatively associated with the amount of urban land and positively associated with

upland forest, wetland forest, and emergent wetlands. There was not a consistent trend

with the amount of agricultural land. Anurans had a positive relationship with agriculture

in Wisconsin and a negative trend in Iowa. The predominance of more intensive

agriculture in Iowa (e.g., row crops) and, in general, less intensive agriculture in









Wisconsin (e.g., pasture and hayfields) was offered as an explanation for the observed

trends (Knutson et al. 1999).

Amphibian species richness of ephemeral to semipermanent wetlands located in an

agricultural (n=9) and urban (n=12) matrix in Minnesota was negatively associated with

road density, distance to nearest wetland, and a very strong effect with the proportion of

urban land adjacent to focal wetlands (Lehtinen et al. 1999). The proportion of

agricultural land was not selected as a predictor of species richness. Species thought to

be particularly sensitive to urban influences included Notophthalmus viridescens and

Hyla chrysoscelis. The American toad (B. americanus) in urban landscapes was

positively correlated with the amount of forest cover (Lehtinen et al. 1999) while Guerry

and Hunter (2002) found a negative association with forest cover for this species in an

agricultural landscape.

Anurans of the New Jersey Pine Barrens appear to respond in a predictable manner

to agricultural and urban development surrounding breeding wetlands. Two species were

considered pineland specialists (Hyla andersonii and Rana virgatipes), four species were

wide ranging throughout New Jersey (e.g., Rana sphenocephala, Rana clamitans, and

Bufofowleri), and four species were considered border entrant or those only capable of

entering the pineland community in habitats disturbed by humans (e.g., Hyla versicolor

and Acris crepitans). The anuran community gradient corresponded to an environmental

gradient. An increase in pH and a decrease in specific conductance and floating

vegetation were associated with increasing human disturbance and amphibian species

composition (Bunnell and Zampella 1999).









The occurrence of 10 of 13 amphibians in 74 Ontario wetlands was positively

correlated with the proportion of forest cover and wetlands on adjacent lands and

negatively correlated with road density and nutrient levels (Houlahan and Findlay 2003).

The effects of adjacent land uses on amphibian composition were strongest within 200m

of the focal wetland. Bufo americanus was the only species positively correlated with

land use intensity and nutrients levels. All other species were negatively correlated with

water nutrient levels. Two species, Rana clamitans and R. catesbeiana, did not show an

association with land use intensity. Although the proportion of surrounding land use in

either agriculture or urban was compared with amphibian species richness, abundance,

and species frequency of occurrence the proportion of forest cover was a better predictor

in all models (Houlahan and Findlay 2003). The proportion of forest cover may be a

surrogate measure of land use intensity because the replacement land use was

conglomeration of often high intensity land uses (e.g., cropland, highways, urban areas).

A multi taxa study in Minnesota observed bird richness and diversity generally

decreased with increasing agriculture surrounding 15 riparian wetlands in Minnesota.

Amphibian abundance increased with the proportion of wetlands and decreased with

increasing rangeland and urban land uses. Amphibians were also influenced by specific

wetland characteristics, such as ditching or eutrophication. Land uses within 500 and

1000 m of the wetland, the shortest distances studied, had the strongest relationship with

amphibian and bird diversity (Mensing et al. 1998).

Landscape scale studies of avian composition indicate bird communities respond to

the patch area size of forest or wetland cover, proportion of forest cover, and edge density









(Miller et al. 1997, Brennan and Schnell 2005). Yet the effect of land use intensity with

in the nonforested land uses on adjacent forested habitats is little understood.

Determining the biological effect of land uses of different land use intensities and

energy flows could have important implications in land use planning. For instance, in the

American west there has been some debate as to whether ranch lands or residential

development has a lower impact on native biodiversity (see Knight et al. 1995 and

Wuerthner 1994). Maestas et al. (2003) compared avian, mesopredator, and plant

communities across a gradient from reserves to ranches to exurban development. Seven

bird species had higher densities in residential areas while two species had higher

densities on ranches, one species on reserves, and three species on both ranches and

reserves. The plant and mesopredator communities were more similar on reserves and

ranches than residential developments.

Bird community response appears to be correlated to assumed anthropogenic land

use intensity gradients. The species abundances and composition of a subandean bird

community in continuous native forest was more similar in forest fragments surrounded

by lower intensity exotic tree plantations than pastures (Renjifo 2001). Avian community

response was species-specific and independent of trophic group and foraging strata

(Renjifo 2001). The more heterogeneous vegetation structure of the plantations versus

pastures was hypothesized as the mechanism for the gradient in bird community

composition. Gradual changes in avian species composition along a land use intensity

gradient (primary forest, secondary forest, agroforestry, annual crops) have also been

observed in Indonesia (Waltert et al. 2004). In Pennsylvania long distance migrants and

forest dependant species were more common and edge species less common in forest









remnants surrounded by silviculture versus those surrounded by agriculture (Rodewald

and Yahner 2001). Finally, a summary of 34 studies in tropical forests indicated species

richness of birds, ants, and lepidoptera decreased in areas converted to agriculture but not

in areas that were selectively logged (Dunn 2004).

Rodewald and Yahner (2001) compared forested landscapes within either an

agriculture or silviculture matrix and concluded that the type of disturbance strongly

influenced the breeding bird community. In general, more forest species declined and

edge species increased in forested landscapes disturbed by agriculture relative to

silviculture disturbances. Cavity nesters, resident species, downy woodpecker, American

crow, indigo bunting, and brown-headed cowbird were significantly more abundant in

forest with adjacent agricultural land uses. Long-distance migrants, forest-canopy

nesters, forest-understory nesters, yellow-billed cuckoo, and hooded warbler were more

abundant in forests adjacent to silvicultural treatments (Rodewald and Yahner 2001).

Norris et al. (2003) explored the avian community composition of forests along a

land use intensity gradient from mature, slightly disturbed, successional, and highly

disturbed (i.e., selectively logged and grazed) forests in Iowa. A vegetation disturbance

gradient was used to rank sites and was essentially an index of land use intensity. Bird

species richness and abundance were not good predictors of the forest disturbance

gradient. However, species groups responded to the forest disturbance gradient.

Neotropical migrants, species of management concern, and area-sensitive species were

most common in forests of low land use intensity while permanent residents, short

distance migrants, and ground nesters were more abundant in forests that had higher land

use intensity. Although only forests were studied, habitat heterogeneity between forest









types was implicated as the driving mechanism for avian community response (Norris et

al. 2003).

The results of these studies indicate that there is a response of amphibians and bird

communities to general land uses and in most landscapes there appears to be a gradient

from natural to silviculture to agriculture to urban.

Habitat Fragmentation

In many landscapes, isolated patches of natural habitat remain in a mosaic of

human land uses. Yet, our understanding of the effects on wildlife populations in

fragments is incomplete. There has been limited success applying the Equilibrium

Theory of Island Biogeography to natural habitats surrounded by human dominated

landscapes (Wiens 1995). Similarly, studies of fragmentation have had difficulty

establishing trends between habitat pattern and biodiversity (e.g., McGargial and

McComb 1995, Lehtinen et al. 1999, Hostetler and Holling 2001). These approaches

appear to be inadequate in their ability to accurately describe observed trends in species

composition, abundance, and persistence in fragmented habitat patches. There are two

main factors influencing wildlife populations in fragmented habitat patches not accounted

for by the island biogeography theory or traditional fragmentation studies. The

surrounding landscape type influences both species-specific dispersal capabilities and the

ecological quality of the habitat patch. The focus cannot only be on habitat pattern, but

needs to include the dynamics of the landscape mosaic (Wiens 1995, Rodewald 2003).

The species that respond (i.e., increase or decrease) to habitat fragmentation may

be evidence that environmental change and species dispersal capabilities are important

determinants of wildlife occupation of habitat patches within human influenced

landscapes. In general, species with poor dispersal capabilities, sedentary species (Wiens









1995, Walter et al. 1999), and habitat specialists (Canaday 1996, Ford et al. 2001, Craig

2002) tend to be the most sensitive to fragmentation. Similarly, wide ranging species and

habitat generalists tend to thrive in fragmented landscapes (Miller et al. 1997, Ford et al.

2001, Craig 2002).

From a species perspective habitat quality and dispersal capabilities are important

factors influencing species composition in isolated habitat patches or fragmented

landscapes. In turn, landscape composition and configuration contribute to the degree at

which a fragment is isolated and suitable.

The composition or intensity of human land use surrounding habitat fragments

can be highly variable. Landscapes that receive periodic or low intensity anthropogenic

inputs (e.g., selective logging, native rangeland) appear to be the least restrictive to

wildlife movement (McGargial and McComb 1995, Wigley and Roberts 1997). Not only

are some species more likely to disperse through low intensity land uses there is often a

subset of the original species persisting in the "degraded" habitat. These residual and

peripheral species have the ability to influence species recruitment, predator-prey

dynamics, and resource competition within natural habitat patches (Ford et al. 2001).

Even high intensity anthropogenic landscapes are not barriers to dispersal for some

species (e.g., Mazerolle 2001, Bulger et al. 2003). Again, it is a combination of a species

behavior and the composition of the surrounding landscape that determines dispersal

permeability.

Inhibiting dispersal is not the only influence adjacent landscapes can have on the

biota of habitat patches. Juxtaposed habitats interchange abiotic and biotic components.

Habitat patches are often influenced by the neighboring landscape reducing habitat









quality for some species (e.g., Debinski and Holt 2000, Davidson et al. 2002). There may

be a gradient of environmental change with increasing human use of the surrounding

landscape (e.g., Blair 1996, Findlay and Houlahan 1997). However, quantifying human

land use and environmental degradation is difficult. Common trends in habitat patches

adjacent to human landscapes are increased nutrient inputs, reduced fire periodicity, and

"edge effects". The intensity of these effects is highly variable and dependent on the

landscape mosaic.

Wildlife populations are influenced by the synergistic effects of habitat loss,

habitat fragmentation, and habitat degradation (Ford et al. 2001). Fragmentation studies

that do not account for the intensity of human disturbance may lead to an

oversimplification or overstating of observed trends. Establishing disturbance trends

with human land use types and land use intensities may increase the accuracy of

predicting the effects of habitat fragmentation on wildlife populations.

Amphibian and Avian Assemblages as Indicators of Land Use Intensity

Amphibian and bird assemblages have been used as an index of human habitat

disturbance. Species richness and diversity vary inconsistently with habitat disturbance

(e.g., Brooks et al. 1998), however, species composition and guild assemblages appear to

offer more potential for quantifying the biotic effect of land use intensity and habitat

degradation on wildlife communities (Brooks et al. 1998, O'Connell 2000).

An evaluation of the amphibian, bird, fish, invertebrate, and plant community

response to the proportion agriculture, urban, grassland, forest, and water circumjacent

116 Minnesota wetlands revealed 79 metrics significantly correlated to land use

(Galatowitsch et al. 1999). Metrics potentially applicable to other wetland systems

included: proportion of wetland birds, wetland bird richness, proportion of insectivorous









birds, importance of Carex, and importance of invasive perennials. Only 2 amphibian

and 23 bird metrics were significantly correlated to land use and possibly an artifact of

their overall diversity. Of all the taxa studied birds were considered the most useful and

amphibians the least useful for monitoring changes in wetlands related to land use

(Galatowitsch et al. 1999). Another north temperate study also found amphibians could

not be used as the only indicator of biotic integrity for ephemeral ponds and wetlands.

Amphibians, crayfish and fish assemblages were needed to accurately detect

anthropogenic disturbance gradient in Indiana (Simon et al. 2000).

O'Connell et al. (1998, 2000) developed an index of biological integrity for the

uplands of the Mid-Atlantic States using songbird response guilds. From literature

searches songbirds were typed either specialist or generalist for eight guild categories;

trophic status, foraging substrate, nest placement, primary habitat, patch size, number of

broods, migratory status, and nest predators (O'Connell et al. 1998). The proportion of

the songbird community that was omnivore, insectivore, single-brood species, and forest

area-sensitive species was significantly different for six levels of habitat integrity

(O'Connell et al. 1998). Sites that were typed as being of poor quality by the songbird

IBI were dominated by agriculture and urban land uses (O'Connell et al. 2000). Finally,

the proportion of forest, landscape-level diversity, and canopy height were selected as the

best predictors of the songbird IBI (O'Connell et al. 2000).

Similarly, Croonquist and Brooks (1991) ranked species into avian response guilds

and compared a protected watershed with a highly developed watershed. Response guild

scores were significantly different between the two watersheds. As the proportion of area

developed increased from the headwaters to the mouth in the impacted watershed the









response guild scores decreased. For example, the proportion of Neotropical migrants

decreased and the proportion of edge and exotic species increased with increasing land

use intensity. At the same time, mammalian guilds did not respond to habitat disturbance

in a predictable manner (Croonquist and Brooks 1991).

It is increasingly apparent that multiple taxa assemblages are better at quantifying

human habitat disturbance. Several studies have indicated that there is often not cross-

taxon congruence with response to habitat disturbance, especially when species richness

is compared (e.g., Lawton et al. 1998, Lund and Rahbek 2002, Provencher et al. 2003,

Su 2004).

Landscape Development Intensity (LDI) Index

Although a comparison of biodiversity between discrete land use types within a

region may reduce variability and increase inference on the effect of land use type on

biodiversity, the typical landscape is often a patchwork of individual landowners, land

use types, and land use intensities. Thus the wildlife assemblage of a region is often

influenced by a variety of land use intensities, and studies that reduce the inherent

variability of landscapes may be reducing realism and applicability of study results. The

question then becomes how to study the response of wildlife composition in the "real"

world but yet have a quantifiable land use intensity to make comparisons with other

studies. The Landscape Development Intensity (LDI) index offers a potential mechanism

for quantifying the human influence to landscapes (Brown and Vivas 2005).

The LDI index is a measure of development intensity based on the energy use per

unit area (Brown and Vivas 2005). Accounting for all of the anthropogenically derived

materials and energies within land uses may not seem pertinent to studies of wildlife

composition. However, with increasing human energy use there appears to be a gradient









of direct, indirect, and cumulative impacts, which have the ability to influence wildlife

abundance, composition, behavior, distribution, and survival. The LDI index quantifies

the human land use gradient which is correlated with increasing noise pollution, light

pollution, human presence, traffic noise, and human structures all of which are examples

of factors that have the ability to affect wildlife (e.g., Bautista et al. 2004, Bird et al.

2004, Crawford and Engstrom 2001, Klem et al. 2004, Longcore and Rich 2004, Riffel et

al. 1996, Slabbekoorn and Peet 2003). Traditional studies of wildlife response to

forest/nonforest situations that do not account for the intensity of human land use within

the matrix may be missing an aspect important in shaping wildlife assemblages.














CHAPTER 2
METHODS

Plan of Study

The methods employed in this study of wildlife response to land use intensity

included three years of field data collection of a number of biotic and abiotic parameters

in 111 isolated forested wetlands and data analysis using several statistical techniques.

The following methods section is organized into three subsections: site selection, field

data collection, and statistical analysis methods.

Site Selection

This study concentrated on small isolated, depressional, forested wetlands in

Florida. Wetlands varied in size from 0.1 to 2. ha (Table 2-1). Site selection was

conducted to insure that approximately one quarter of the total 111 study sites were

sampled within each of Florida's four wetland ecoregions (Lane 2000)(Figure 2-1). In

addition, wetlands were selected throughout Florida to represent four apriori landscape

settings; reference, silviculture, agriculture, and urban/residential. Because of the

difficulty of obtaining access to wetlands and the fact that depressional forested wetlands

are not randomly distributed throughout Florida, selection of sample wetlands was not

conducted randomly.









Table 2-1. Mean wetland size (ha) and the minimum and maximum wetlands sampled
per land-use category and region. Means were not significantly different
between land use types or Florida region (one way ANOVA, Tukey multiple
comparisons, land use types (p=0.266) and between regions (p=0.066)).
mean (s.d.) min max
Reference 0.66 (0.41) 0.20 1.75
Silviculture 0.49 (0.24) 0.22 0.96
Agriculture 0.81(0.46) 0.12 2.11
Urban 0.74 (0.48) 0.10 2.03

Panhandle 0.60 (0.38) 0.10 1.75
North 0.73 (0.38) 0.26 1.85
Central 0.88 (0.52) 0.12 2.03
South 0.61 (0.44) 0.18 2.11


ASO


2003 surveys
SAgriculture
Reference
x Silviculture
A Urban
Panhandle
North
SCentral
South


*A
'* (

r


Figure 2-1. The distribution of the survey sites by land use and Florida region.









Table 2-2 lists the number of wetlands sampled in each of the regions by apriori

landscape matrix. Twenty-seven wetlands were sampled in the panhandle, thirty in the

northern peninsula, twenty-nine in the central region, and twenty-five in southern Florida.

Twenty-seven study wetlands were located within predominately agricultural landscapes,

thirty-three were in reference landscapes, eleven wetlands were in or adjacent to pine

plantations, and forty wetlands were located within an urban/residential matrix.

Table 2-2. Count of 2003 sites by land-use and Florida region.
Reference Silviculture Agriculture Urban Total
Panhandle 7 4 6 10 27
North 7 7 6 10 30
Central 11 0 8 10 29
South 8 0 7 10 25
Total 33 11 27 40 111


Access to wetlands was gained by contacting landowners and land managers

directly and with the aid of agricultural extension agents. Sites were grouped into blocks

of 4 to 6 sites incorporating at least one low intensity site (e.g., reference or silviculture)

with higher intensity sites (e.g., agriculture and urban). Each wetland within a site block

(four to six sites) was sampled twice within a four-day period (generally two to four sites

were sampled per day) for a total of 111 wetlands sampled between 30 May and 27

August 2003. Wetlands within the site blocks were sampled to minimize driving time

regardless of land use type. Site order was reversed for the second site visits when

possible. An attempt was made to sample site blocks alternately among the four Florida

eco-regions throughout the 2003 sampling period.

Figure 2-2 shows the mean proportion of land cover type within 200 meters of each

focal wetland for the four apriori land use categories. The mean proportion of

undeveloped land comprised greater than 20% of the land area within 200 meters of the









agriculture, silviculture and urban land use categories, and greater than 97% of the area

for reference wetlands. Agriculture, silviculture, and urban sites all had means greater

than 70% of the land cover within 200 meters that defines each category.


iA-t]


* Undeveloped
] Pine Plantation
*Agriculture
] Urban


R S A U
Sites by land use

Figure 2-2. The mean proportion of land cover type within 200 meters of each focal
wetland by land use categories. R=reference sites, S=silviculture sites
A=agriculture sites, and U=Urban/residential sites.

The dominant land use surrounding 13 of the 27 agriculture sites was cattle pasture.

Five sites had row crops or spray fields and three sites had citrus plantations as the

dominant surrounding agricultural land use. Six apriori agriculture sites consisted of a

combination of pasture and row crops. Sixteen of the 19 sites that had cattle pasture had

improved pasture (e.g., evidence of planted non-native grasses, drainage ditches and

canals, clear-cut forests etc.) and the remaining three were within native rangeland.

Among the reference sites two wetlands were in city parks, four were in county

parks and seven were in state parks, preserves or reserves. Two sites were on military

bases, two were in national forests and three were in national wildlife refuges or reserves.

Finally, six sites were within state forest, three were on private conservation tracts, and

four were on water management district properties. Approximately 20 of the 33


S1.0

. 0.8
E
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8 0.6
4-
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o 0.4
o
g 0.2
0n
n n









reference sites allowed game hunting and selective timber harvest operations within the

landscape surrounding the study sites.

Five of the 11 sites categorized as silviculture sites were adjacent to pine

plantations within state forests and one was in a national forest. The remaining five

silviculture sites were on private timber company land.

The dominant urban land use surrounding twelve of the 40 a priori urban sites was

categorized as industrial, commercial, or institutional and four of these also had a large

proportion of the surrounding landscape as arboretum or silviculture. Twenty-two urban

sites were adjacent to primarily residential land uses of varying human densities, and 6

sites were a combination of residential and commercial or institutional. Four of the 22

residential sites also had golf courses as a dominant surrounding land use, and finally,

two of the residential sites had a large proportion of city park or maintained open space

within 100 meters.

Field Sampling

Sampling of amphibian and avian species was conducted within the limits of each

focal wetland and the circumjacent landscape within 200 m. Habitats adjacent to the

focal wetland were included in the survey because many amphibians are dependent on

uplands during part of their life cycle and only use wetlands seasonally. Furthermore, the

relatively small size of the focal wetlands in this study may only be a component of some

avian species home range (Schoener 1968), and restricting species detections to just the

wetland during a two-visit synoptic sample could underestimate the wildlife use of the

habitat. Two hundred meters was chosen because it encompasses the area of most avian

territory sizes (Schoener 1968), and is the maximum radius of audible detectability for

most eastern forest bird species (Whitcomb et al. 1981). Two hundred meters from the









wetland edge also includes the mean minimum "core" terrestrial habitat area for most

amphibians (Semlitsch and Bodie 2003).

All site surveys were conducted by the same observer. Survey start and end times

were recorded for each survey and surveys were started anytime during the daytime

hours. The air temperature, cloud cover (weather bureau sky condition codes), and

estimated wind speed (Beaufort land based scale) were recorded at the beginning of every

survey. Surveys were not started during heavy rains.

Visual and Auditory Encounter Surveys

Each survey began by walking a 15-minute transect through the wetland that

started five meters from the wetland edge at the closest physical access point to the

wetland. The surveyor slowly walked an approximate straight line through the center of

each wetland stopping periodically to listen and visually scan the surroundings. One

minute was spent surveying at opposite sides of the wetland while standing in the

wetland/upland ecotone. Species flying above the sites that were suspected to be

foraging (e.g., swallows, vultures, etc.) or fleeing from the site were included. After the

completion of the 15-minute transect, species observed opportunistically while

conducting other onsite measurements and sampling were also recorded. The time

elapsed from the survey inception was recorded for the first observation of each species

detected in the wetland and within the landscape, as well as how the species was first

detected (song, call, or visual, and for birds if it was flying). Logs, trash, and other cover

objects within the wetland boundary were turned when possible to detect sheltering

species.









Automated Recorder Surveys

An automated recording system (Bedford Technical model ARS-o6X001) was

deployed at 96 wetlands during the first site visit. To protect the microphones from

inclement weather they were placed inside an inverted plastic water bottle about 2.5cm

back from the cut off bottom end (Dodd 2003). The microphone and plastic cover were

tied to a branch with fishing line, at a slight downward angle, suspended 1.5 to 2m above

ground level, and directed at the deepest area in the center of each wetland. Leaves and

small branches within one meter of the microphone were removed. The timer on each

automated recorder was set to record for one minute every hour for a total time of

approximately 45 hours.

The same observer listened to all of the cassette tape samples in the lab with a

headset. All species heard singing or calling and the time of day were recorded for each

one-minute acoustic sample and the maximum number of anurans heard calling was

coded by species; 1= one individual heard, 2 = more then one heard but individuals

distinct, 3 = multiple calls but individuals cannot be distinguished. Finally, notes were

made of all identifiable noises (e.g., airplanes, automobile traffic, fireworks, rain, wind,

etc.) and whether the noises obscured the detection of calling vertebrates during each

one-minute acoustic sample.

Combined Sampling Effort

Two 15-minute transect surveys were conducted at all 111 sites in 2003 for a total

of 3,330 standardized minutes. An additional 7,753 minutes were spent at the sites

documenting species observations beyond the 15-minute time period along with 3,762

minutes of audible cassette tape recordings, for all sites combined. The mean amount of

total minutes spent recording species composition, including audio recorder minutes, at









each agriculture, reference, silviculture, and urban site was 142.3, 145.6, 122.2, and 121.3

minutes respectively. The amount of time spent sampling reference sites was

significantly greater than at urban sites. There was not a significant difference in the

amount of time spent surveying at all other land use types. Audio loggers were used at

24 agriculture sites, 25 reference sites, 9 silviculture sites and 34 urban sites, 89%, 76%,

82%, and 85% of the sites respectively. The sites at which cassette recordings were

taken, and the amount of audible minutes for each site, are given in Appendix A.

Dip-net Sampling

At sites that had standing water twenty dip-net sweeps were used to sample for

larval anurans and salamanders during one of the two visits to each site. The dip-net was

a U.S. standard 30 mesh D-frame net and each sweep was approximately 1.0m long. The

twenty dip-net sweeps were stratified across each wetland, perpendicular to the 15-

minute survey transect, targeting representative microhabitats. Caudates and larval

anurans (tadpoles) were identified, tallied, and released after each sweep (Conant and

Collins 1991, Altig et al. 1999). Some specimens could not be reliably identified to

species in the field (e.g., Rana sphenocephala, R. capitol, Rana grylio, Hyla squirella,

Bufo terrestris, and some very small individuals) and were given a unique unknown

grouping (e.g., unknown a, unknown b etc.), tallied and released. All fish captured

during the sweeps were identified to genus, and species when possible, tallied, and

released. The presence of known macroinvertebrate tadpole predators were also recorded

but not tallied (e.g., Belostomatidae, Decapods, Dolomedes sp., Dytiscidae, Gyrinidae,

Nepidae, Notonectidae, Odonata).









Terrestrial Insect Sampling

A 38 cm diameter sailcloth sweep net was used to sample terrestrial

macroinvertebrates during one of the site visits at 82 of the sites by walking across each

wetland and sweeping the net in an approximatel80 degree arc across shrubs and forbs

100 times. The collected insects and vegetation were placed in a site-specific plastic bag

then put in an ethyl acetate killjar with the bag left open. Each site bag was removed

from the killjar after 20 minutes, placed on ice and/or refrigerated until they could be

processed. Processing was conducted in the lab and involved separating all

macroinvertebrates visible without the aid of microscopes or hand lenses from pieces of

vegetation. Macroinvertebrates were identified to order, grouped by species, and

counted. Each macroinvertebrate order and collected vegetation was weighed separately

to the nearest milligram.

Other Biotic Sampling

Prior to the sampling for amphibian and avian fauna in 2003 the sites were sampled

in either 2001 or 2002 for macrophytes, aquatic macroinvertebrates, diatoms, tree

composition, percent canopy cover and basal area (Reiss 2004). As part of the Reiss

(2004) study several metrics using macrophytes were calculated for the wetlands

including macrophytes sensitive or tolerant to development, the proportion of

macrophytes native or exotic to Florida, the proportion of macrophyte species that were

wetland dependent, and an average coefficient of conservatism score based on

macrophyte presence were calculated for each site (Reiss 2004). Reiss (2004)

summarized these metrics along with macroinvertebrate and diatom metrics into a Florida

Wetland Condition Index (FWCI) score for each site. In that same study each site was









given an average Wetland Rapid Assessment Procedure (WRAP) score using methods

outlined by Miller and Boyd (1999).

Environmental Variables

Water Sampling

A grab water sample was taken from an undisturbed area of each wetland that had

standing water at least 10cm deep following the standard operating procedure of the

Florida Department of Environmental Protection (FDEP) for water quality measurements.

A hand held YSI-55 Dissolved Oxygen meter was used to measure dissolved oxygen and

water temperature in the same general location immediately prior to each water sample.

Water samples were sent to the FDEP Central Chemistry Laboratory, and analyzed for

ammonia-nitrogen (EPA 350.1), color (EPA 110.2), nitrate/nitrite-nitrogen (EPA 353.2),

pH (150.1), specific conductance (EPA 120.1), total Kjeldahl nitrogen (EPA 351.2), total

phosphorus (EPA 365.4), and turbidity (EPA 180.1).

The maximum water depth at each site was measured by walking across each

wetland, and recording the water depth in the deepest area encountered to the nearest

centimeter. For the vast majority of the sites the deepest area was near the center of the

wetland. Maximum depth measurements were not taken in areas that were obviously

animal burrows, tree tip-ups, ditches, canals, or excavated pools.

Fire History

During the site visits evidence of any fire in the surrounding landscape as well as

within the sample wetland was recorded. Fire evidence included charred, stumps, snags,

and logs, or burn marks on the boles of living trees. An estimate was made about

whether the fire had occurred more or less than 6 years ago. This was a gestalt estimate

based on vegetation height and composition as well as characteristics of the charring. In









some instances landowners or land managers were able to date the last fire in the area. A

sample wetland and surrounding landscape were scored separately as follows, zero for no

evidence, one for old evidence (>6 years ago), and two for recent fire evidence (<6 years

ago).

Landscape Variables

Several landscape indices were calculated from Land Use/Land Classification

(LU/LC) coverages, aerial photos, and site visits as follows: 1) Landscape Development

Intensity (LDI) index using 100m buffer around each sample wetland and current LU/LC,

2) a historical 200m LDI using 1974 LU/LC data, 3) the percent of the 200m buffer

consisting of the following land types; undeveloped land, pine plantation, agriculture,

urban/residential, roadway, wetland, contiguous forest, and canal/retention pond, 4)

distance to nearest paved road and two closest wetlands.

Landscape Development Intensity Index

A 100m buffer LDI score was calculated with ArcView software (ESRI 1995) for

all of the sample wetlands (Brown and Vivas 2005). The LDI index is intended to

represent the amount of human impact on systems. The assumption is the amount of

human energy use in the landscape is associated with the amount of change to adjacent

systems through direct, secondary, and cumulative affects. The LDI score for each site

was calculated as follows.

LDITotal=Z %LUi LDIi

LDITotal = LDI ranking for 100 meter buffer around each forested wetland site

%LUi = Proportion of the 100 meter buffer in land cover/use i

LDIi = Landscape development intensity coefficient for land use i









The LDI coefficients for each land cover/use (FDOT 1999, FDEP 2002) are

derived from previous studies of energy use within common land use systems and

represent the amount of nonrenewable energy (e.g., electricity, fuels, fertilizers, and

pesticides) used per unit area (Brown and Vivas 2005). The range of coefficients has

been normalized on a scale from 1 to 10. A land use coefficient of 1 equals no

nonrenewable energy use (i.e., natural system) and land use coefficient of 10 is the most

intensive land use (i.e., central business district that has a average height of 4 stories).

For comparative purposes other typically encountered land cover/uses encountered in this

study and their LDI coefficients are pine plantation (1.58), native rangeland with

livestock (2.02), improved pasture with livestock (3.74), row crops (4.54), low density

single family residential (6.90), high density single family residential (7.55), two lane

highway (7.81), four lane highway (8.28), and commercial (9.18)(a more complete list

can be found in Brown and Vivas 2005).

Digital Orthophoto Quads (DOQ) from 1999 (LABINS 2002) and site visits in

2001 and 2002 were used to classify land cover/use within 200 meters of each wetland.

An LDI score based on the 1974 land use coverage (FDEP 2002) within 200m of each

site was also calculated.

Landscape indices

In addition to the LDI calculations several other indices of landscape context were

tabulated within the 200m buffer of each wetlands. These included the proportion of land

area comprised of the buffer and the sample wetland that was agriculture, undeveloped,

silviculture, urban, roadway, canal/retention pond, and wetland. The proportion of each

land use was calculated using ArcView software (ESRI 1995, FDEP 2002), 1999 DOQs,

and site visits. A 200 m buffer was used because this was the area of biotic sampling,









and it was shown by Houlahan and Findlay (2003) that land uses within 200m have the

strongest effect on amphibian assemblages.

Land areas with very little to no annual human derived energy inputs were

classified as undeveloped (e.g., pine flatwoods, dry prairies, hardwood hammocks,

wooded vacant lots). Land areas categorized as agriculture were primarily used for

growing crops and or raising domesticated animals. It was often difficult to determine

the livestock grazing intensity in forests and wetlands, especially from aerial photos, and

these areas were designated undeveloped, as were, woodlots in urban and agricultural

areas. Silviculture consisted of clear-cuts and pine plantations, however selectively cut

forests were included with the proportion undeveloped. Golf courses, recreational fields,

cleared lots, buildings, and roads along with their associated "green space" (i.e., yards

and right-of-ways) were classified as urban.

Many of the classifications were not mutually exclusive. For example, the

proportion of the 200 m buffer considered roadway was also included in the proportion

classified as urban. However, single track "jeep trails" or forest roads that were on the

original soil substrate and were not graded were included with the land use on which they

occurred (e.g., agriculture or reference). The proportion of the total area classified as

wetland was calculated, and included in the proportion undeveloped. Canal/retention

pond area was calculated alone but also included within the land use in which they

occurred (e.g., silviculture, agriculture, urban).

The variable "% contiguous forest (200m)" was the area of the sample forested

wetland plus the forest adjacent to the focal wetland as a proportion of the combined

wetland and 200 m buffer area. For example only the forested portion of the focal









wetland would be contiguous forest at sites surrounded by wet or dry prairie. Citrus

groves, forested fence lines, and pine plantations greater than 5 years from planting were

included as contiguous forest area if adjacent or continuous with the focal wetland.

Forests were considered areas with greater then 25% canopy cover and breaks in the

forest canopy greater then 25m were considered discontinuities of the forest canopy.

The distances from the focal wetland to the nearest wetland, the second nearest

wetland, paved road, major river system (FDEP 2002), and forest cover were measured

using ArcView software (ESRI 1995). Distance to the two closest wetlands was

measured from the focal wetland perimeter to the closest edge of any type of wetland

delineated in FLUCCS code. Distance to paved road was measured from the focal

wetland edge to the closest paved road using 1999 DOQs. There is some indication that

major floodplain forests may be wildlife corridors, thus the distance to the nearest river

system from the focal wetland edge was measured using an FDEP (2002) GIS coverage

of Florida rivers. Finally, the distance to the nearest forest patch from the focal wetland

edge was measured for sights without contiguous upland forest.

Data Analysis

First, comparisons were made between land use types using the measured

environmental and landscape variables. Second, associations between LDI scores and

landscape and environmental variables are presented. Then relationships between the

species assemblages and landscape content and environmental variables were explored

looking for possible explanations for observed trends in amphibian and avian species

composition. Finally, characteristics of the species life history traits were compared with

environmental variables, land use type, and land use intensity.









Water Chemistry and Landscape Variables

Environmental variables were measured and then compared for significant

differences between land use types and LDI quartiles (sites arranged into approximately

four equal sized groups and ranked based on LDI scores; LDI quartile one consisted of

the sites with the lowest LDI scores and LDI quartile four consisted of the sites with the

highest LDI scores). All variables were tested for normality, Kolmogorov-Smirnov

Normality Test, and transformed when necessary, (power, logarithmic, and arcsine

squareroot transformations). If normality was still not met, non-parametric tests were

conducted, and parametric tests on normally distributed data. Univariate correlations

were used to compare the site LDI scores with all other environmental and site sampling

variables.

Species Richness

Presence or absence of amphibian and avian species were recorded for each site.

The mean species richness, total species richness, and the minimum and maximum

number of species observed at the sites for each taxon were compared by land use type,

LDI quartile, and Florida region. Visual estimations of species accumulation curves

(species richness by time) suggested that the majority of the avian and amphibian species

pool was sampled at each site.

Amphibian frequency of site occurrence was calculated by land use type and for all

sites. Amphibian species frequency of occurrence is equal to the number of sites species

X was found by land use divided by the number of sites in Florida counties within species

X's range (ARMI 2004) by land use type.









Species Composition

To test for differences in the species composition of the apriori land use types and

LDI quartiles multi-response permutation procedure (MRPP) was used in PC-ORD

(McCune and Mefford 1999). MRPP is a nonparametric test of differences between

groups (Mielke 1984). The Sorensen (Bray-Curtis) distance measure was selected to

calculate the distance matrix used in the MRPP analyses (McCune and Grace 2002).

Amphibian and avian community structure were ordinated using nonmetric

multidimensional scaling (NMS)( Kruskal 1964, Mather 1976) with PC-ORD software

(McCune and Medfford 1999). Ordination is used to reveal dominant patterns in data by

arranging items (sites or species) along an axis or multiple axes (McCune and Grace

2002). NMS is an ordination technique that does not assume normality and uses a

repeated search to maximize the rank correspondence between the original

multidimensional space and distances in the reduced dimension of the ordination space

(Peterson and McCune 2001).

For the amphibian and avian NMS ordinations 40 runs with a random starting

configuration were performed with the data using the Sorensen (Bray Curtis) distance

measure. A run consisted of a series of solutions that started with 6 axes and stepped

down in dimensionality to 1 axis. Then the data was randomized (species presence

shuffled by site) and 50 runs were completed. Stress (an inverse measure of fit) is

calculated for the runs with the real data and the randomized data. The best ordination

solution is then selected based on the real run with the lowest stress, and dimensionality

was determined when adding another dimension did not reduce the stress by 5 or less.

Finally, the selected dimensionality had to have a final stress lower than 95% of the

randomized runs.









Rarely encountered species can create noise in the NMS ordination results, and it is

recommended that species that occur at fewer than 5% of the sites should be deleted

(McCune and Grace 2002). Thus, for the amphibian NMS only species that occurred at 2

or more sites were included and for the avian NMS only species that occurred at 3 or

more sites were included. Including amphibians that occurred at two or more sites and

birds that occurred at three or more sites in the NMS ordinations resulted in models with

the lowest stress.

Correlations were conducted between NMS ordination axis scores and the

measured site environmental variables. The correlation between axes scores and the LDI

and FWCI were of interest and if not already the strongest correlation of the

environmental variables was displayed along with the strongest correlation coefficients

with the remaining ordination axis in joint plots of the data. With joint plots the angle

and length of the vector indicates the strength of the relationship between the

environmental variable and ordination axes scores (McCune and Grace 2002).

Stepwise multiple regression with forward and backward selection (alpha=0.05)

was used to explore the relationship between the NMS axes scores with landscape

variables, wetland variables, and sampling attributes. The variables listed in Appendix B

were screened for covariation and a choice was made between highly correlated (>0.7)

variables. The resulting independent variables were used in the stepwise multiple

regressions can be found in Table 2-3. Exploring the influence of LDI on the NMS axes

scores was of interest; however, LDI was also strongly correlated (>0.7) with the % urban

and the % undeveloped. The percent urban land and percent undeveloped land are

considered important in influencing species composition, so all (LDI, % urban, %









undeveloped) were included for selection in the stepwise regression analyses. Similarly,

latitude and longitude of study sites were strongly correlated (=-0.71), however both

variables were included for selection in the stepwise multiple regression models. Total

phosphorus was replaced separately by all of the nitrogen based water chemistry

variables and separate regression analyses were conducted to explore the relationship

between the amphibian NMS axes scores and nutrients. However, nitrogen has been

associated with causing stress in amphibians, and all of the nitrogen based water

chemistry measures were individually included in separate runs of the multiple regression

analyses.

Estimating habitat condition can be done at three levels of assessment: remotely,

rapidly with site visits, or by more thorough means, this is a goal of the EPA and others

wishing to evaluate landscapes (Brooks et al. 1999). Thus amphibian and avian based

NMS axes scores were first only compared with 14 landscape variables that can be

calculated remotely (without a site visit), then again with the landscape variables along

with variables measured within the focal wetland and attributes of sampling (e.g., survey

length, starting time, Julian date of first site sample). Finally, stepwise multiple

regression models were created for amphibian based NMS axes scores using landscape,

focal wetland, sampling attributes, and water chemistry parameters (Table 2-3). Water

chemistry parameters were not included in avian stepwise multiple regression models.













Table 2-3. Independent variables included in stepwise regressions predicting the
amphibian based NMS axes scores.
Focal Wetland and
Landscape Sampling Attributes Water Chemistry
LDI % Broad Leaved Trees Dissolved Oxygen
1974 LDI % Evergreen Trees Total Phosphorus
% Agriculture (200m) % Exotic Trees pH
% Canal (200m) Basal Area Conductivity
% Roadway (200m) Fire Evidence in Wetland
% Silviculture (200m) Maximum Depth (cm)
% Undeveloped (200m) FWCI
% Urban (200m) Amphibian Species Richness
% Wetland (200m) Fish Species Richness
Distance to Nearest Wetland (m) Plant Species Richness
Distance to Paved Road (m) Julian Date
Focal Wetland Area Starting Time (survey 1)
Latitude Starting Time (survey 2)
Longitude Survey Length (min)


Indicator Species

The amphibian NMS ordination was used to calculate species scores with weighted

averaging in PC-ORD (McCune and Mefford 1999). Site ordination scores were used to

calculate the average position for each species along the ordination axis. The large

sample size of sampled birds made it difficult to construct an observable graph thus avian

species scores were not displayed.

Indicator Species Analysis (Dufrene and Legendre 1997) was used to select

amphibian and avian indicators of land use and LDI quartile in PC-ORD (McCune and

Mefford 1999). Indicator Species Analysis supplies indicator values to species based on

their faithfulness of occurrence to apriori groups from a range between 100, perfect

indication, and zero, no indication. Indicator values are calculated using a species









relative abundance and relative frequency (McCune and Grace 2002). To test for

statistical significance 1000 Monte Carlo randomized runs were conducted for each

Indicator Species Analysis. Avian Indicator values were conducted for land use and LDI

quartile. The indicator values for amphibians were strongest when LDI quartiles 3 and 4

were combined and when reference and silviculture, and agriculture and urban sites were

combined.

Community Characteristics

Amphibian and avian species have life history traits that may influence their

response to anthropogenic land uses or land use intensity. Within each class species

share some of these life history traits (i.e., guilds) and may react in a similar manner to

land use change. Identification of life history traits that respond to habitat characteristics

of land uses or land use intensity may make it possible to apply results to other regions

and to the exploration of driving environmental mechanisms of species composition

change with habitat alterations.

Amphibian community

The hydrology of wetlands is often altered with increasing land use intensity. To

explore trends in amphibian community change with increasing land use intensity,

amphibians were grouped into either obligate or facultative users of ephemeral wetlands

for reproduction (Moler and Franz 1987), and one amphibian species sampled was not

dependent on wetlands for reproduction (Appendix E). Comparisons were made between

the proportion of the amphibian communities' ephemeral wetland dependency between

land uses and LDI quartiles. Amphibian eggs attached to vegetation in wetlands that

have highly variable water levels may be prone to sedimentation or experience periods of

anoxia (Azous and Homer 1997). To explore the relationship between egg placement









strategy and land use intensity amphibian species were first grouped into one of four

strategies (Appendix E)(Bishop 1943, Conant and Collins 1991, Bartlett and Bartlett

1999). Comparisons were made by strategy between the proportion of the amphibian

community that attaches their eggs, has free floating eggs, is capable of attaching or has

free floating eggs, and has terrestrial development, by land use category and LDI quartile.

Avian community

Avian species were grouped by, trophic guild (carnivore, herbivore, insectivore,

and omnivore) (adapted from Martin et al. 1951, De graaf et al. 1985, Poole and Gill

200X), foraging substrate/technique (aerial screener, bark gleaner, canopy gleaner,

ground gleaner, hawker, sallier, and water based)(adapted from De graaf et al. 1985),

wetland dependency (Poole and Gill 200X, De Graaf et al. 1985), native/exotic,

Neotropical migrant, establishes feeding territories (Poole and Gill 200X), nest location

(canopy, cavity, ground, human structure, parasitic, and shrub)(Ehrlich et al. 1988), and

Florida population trend (significant increasing and decreasing species (p<0.1) between

1966-2002)(Saur et al. 2003). Species weights derived from Sibley (2000) (Appendix F).














CHAPTER 3
RESULTS

The organization of the results section begins with comparisons of attributes of

the sample wetlands and circumjacent landscape by land use. Then the site LDI index

scores are compared with physical, biological, landscape, and water chemistry attributes

of the study sites. Finally, amphibian and avian richness and composition results are

presented and possible explanatory mechanisms for the observed trends are explored.


Wetland Characteristics

Wetland Vegetation

Generally, there were few differences in the macrophyte species richness,

composition of the wetland overstory, canopy cover, and mean basal area of the wetland

vegetation between wetlands grouped by apriori land uses (Table 3-1). The dominant

overstory woody vegetation of the sampled wetlands was Taxodium ascendens followed

by Nyssa biflora (Table 3-2). Taxodium ascendens with N. biflora comprised more than

40% of the basal area at 108 of the 111 sites. Broad-leaved trees averaged less than 24%

of the basal area at all sites combined and were generally more common at urban and

silviculture wetlands. More of the basal area at agriculture and urban wetlands consisted

of exotic woody species and no exotic woody species were detected within the basal area

measurements at the reference and silviculture wetlands.









Table 3-1. The mean and standard deviation of the wetland macrophyte species richness,
percent canopy cover, and tree basal area (>10.2 dbh) by land use.
Macrophyte Basal area
species richness % canopy cover* m2/ha
Reference 34.7 (10.5)a 81.1 (7.6)a 29.2 (8.8)a
Silviculture 29.2 (14.8)a 81.7 (7.7)ab 30.5 (8.7)a
Agriculture 37.9 (13.5)a 82.4 (8.6)ab 35.4 (13.1)a
Urban 36.6 (10.4)a 85.5 (6.0)b 32.9 (13.1)a
Land uses that share letter superscripts do not have significantly different means (*= One-
way ANOVA, Tukeys multiple comparisons, **= Mann-Whitney U-test, p<0.05).


Table 3-2. The mean proportion and standard deviation of the basal area of the wetland
overstory vegetation by land use.
shrub Taxodium* Nyssa*
evergreen broad-leaved species exotic speciesascendens biflora
R 0.12(0.15)a 0.11(0.16)a 0.05(0.12)a 0.00(0.00)a 0.83(0.18)a 0.05(0.10)a
S 0.12 (0.05)a 0.37 (0.29)b 0.06 (0.06)a 0.00 (0.00)a 0.57 (0.28)b 0.21 (0.19)b
A 0.08 (0.11)a 0.14 (0.17)a 0.02(0.03)a 0.01(0.03)b 0.80(0.20)ac 0.07(0.11)ac
U 0.08(0.12)a 0.26 (0.26)b 0.02(0.03)a 0.04 (0.13) 0.70(0.25)bc 0.11(0.16)bc
= Shrub species were woody species normally classified as shrubs (e.g., Cyrilla
racemiflora, Myrica cerifera), and were sampled in the wetland basal area measurements
(stems >10.2cm dbh). = The most dominant and *= the second most dominant tree
species. Land uses that share letter superscripts do not have significantly different means
(Mann-Whitney U-test, p<0.05).

Wetland Water Depth

The mean maximum water depth for all sites that had standing water at the time of

the first visit in 2003 was 44.7 cm (Table 3-3). There was not a significant difference

between the mean maximum depth of agriculture, reference, silviculture and

urban/residential land use categories (one-way ANOVA, Tukey multiple comparisons,

(p=0.401)). However, there were large differences in the percentage of sites that did not

have standing water; 35% of the urban sites were dry at the time of the first visit followed

by silviculture (27%), agriculture (15%) and reference sites (9%)(Table 3-3).









Table 3-3. The mean and standard deviation of the maximum wetland water depth at
sites with standing water by land use, and the count of sites without standing
water at the time of sampling by land use.
water depth dry sites
Reference 47.8 (21.4)a 3
Silviculture 53.9 (22.0)a 3
Agriculture 41.8 (25.8)a 4
Urban 41.0 (24.7)a 14
Shared letter superscripts indicate no significant difference (one-way ANOVA, p<0.05).

Wetland Water Chemistry

Water samples were analyzed for chemical constituents at 87 of the 111 sites. Sites

where water samples were not taken were dry or had water depths that were less than

10cm during the first site visit. There was not a significant difference in water chemistry

parameters between land-uses for temperature, color, and nitrite-nitrate-nitrogen (Table

3-4). Reference sites had significantly higher dissolved oxygen levels, and lower pH,

turbidity, conductivity, ammonia, TKN and total phosphorus than both agriculture and

urban sites. Reference and silviculture sites had significantly lower nutrient levels

(ammonia and total phosphorus) than both agriculture and urban sites. Agriculture and

urban sites did not have statistically significant differences for the measured water

chemistry parameters.

Wetland Terrestrial Insects

Mean insect species richness, mean total individuals captured, and mean biomass

were significantly greater at agriculture (42, 146, 1.3g) sites than at reference (27, 59,

0.5g) and urban (33, 84, 0.5g) sites (One-way ANOVA, Tukeys multiple comparisons,

p<0.05). Silviculture sites had the lowest sample size (n=9) and mean insect species

richness, mean total individuals captured, and mean insect biomass (32, 87, 0.4g

respectively) were not significantly different from the other land use categories.









Table 3-4. Mean and standard deviation of water chemistry parameters by land use.
Land uses that share letter superscripts do not have significantly different
means (Mann-Whitney U-test, p<0.05).
Reference Silviculture Agriculture Urban
Sites sampled (n) 30 8 23 26
Dissolved oxygen (mg/L) 3.4 (1.9)b 2.0 (1.3)ab 1.7 (1.5)a 1.5 (1.5)a
Temperature (C) 26.2 (1.6)a 25.9 (1.5)a 25.9 (1.5)a 26.0 (2.0)a
Color (PCU) 243 (146)a 351 (234)a 390 (425)a 228 (191)a
pH 5.2 (1.2)b 4.4 (0.4)b 6.5 (0.7)a 6.2 (1.3)a
Turbidity (NTU) 2.2 (3.0)b 1.9(1.7)a 17.4 (48.7)a 10.3 (31.5)a
Conductivity (umhos/cm) 93 (171)b 46 (18)b 174 (203)a 173 (238)a
Ammonia (mg N/L) 0.02 (0.02)b 0.02 (0.01)b 0.74 (1.92)a 0.11 (0.13)a
N02-N03-N (mg N/L) 0.01 (0.02)a 0.01 (0.01)a 0.01 (0.03)a 0.01 (0.02)a
TKN (mgN/L) 1.13 (0.37)b 1.28 (0.38)ab 4.13 (6.23)a 1.58 (0.54)a
Total phosphorus (mg P/L) 0.04 (0.03)b 0.04 (0.02)b 1.59 (3.53)a 0.18 (0.17)a


Circumjacent Land Use Characteristics

Table 3-5 displays the mean distance to the nearest and second nearest wetland,

distance to the closest paved road, and the mean proportion of wetland, road, and

canal/retention pond within 200 meters of each site. The most significant differences in

these measures were between reference and urban sites. For example, the mean distances

to the two closest wetlands at urban sites, 213 and 423 m, were more than double that of

reference sites. The mean distance to a paved road was two orders of magnitude lower at

urban sites than the distance for agriculture, silviculture, and reference sites. Reference

sites have the highest mean proportion of wetland within 200 meters (p<0.01), while

agriculture and urban sites had significantly more canals, ditches and retention ponds

within 200 meters than reference and silviculture sites.









Table 3-5. The mean and standard deviation (in parentheses) of attributes of the
landscape surrounding the sample wetlands by land use type. Shared letter
superscripts indicate means that are not significantly different (p>0.05).
distance to (m) Reference Silviculture Agriculture Urban
nearest wetland 96a (79) 158a (136) 136a (103) 213b (196)
2nd nearest wetland 170a (115) 299ab (212) 252a (145) 423b (259)
paved road 1600a (1410) 1948a (2201) 1248a(1172) 57(152)
proportion of 200m buffer
wetland 0.25 (0.01) 0.09a (0.07) 0.13a (0.09) 0.07b (0.07)
road 0.02a (0.02) 0.02b (0.02) 0.01b (0.01) 0.08 (0.09)
canal/retention pond 0.00a (0.00) 0.00a (0.00) 0.02b (0.02) 0.03b (0.06)


Landscape Development Intensity

The mean and median 100 meter buffer LDI score for all sites sampled in 2003

combined was 3.49 and 3.64 respectively. Fourteen sites received the lowest possible

LDI score (LDI=1.0), and the highest site LDI score was 7.8 (appendix d). Table 3-6

shows the 111 sites grouped by 100 meter buffer LDI score quartiles. Quartile 1 is

comprised entirely of, the apriori, reference category sites and has the least amount of

variation in LDI scores. Quartile 2 contains all of the silviculture sites along with sites

from the other three major land use categories. Quartile 3 was a mix of agriculture and

urban sites. Quartile 4 consisted almost entirely of urban sites with only one high

intensity agriculture site.

Table 3-6. The range, mean, and standard deviation of LDI scores by LDI quartile, and
the number of sites by land use type for each LDI quartile.
LDI Sites
quartile range mean stdev n ag(n) ref(n) siv(n) urb(n)
1 1.00-1.12 1.03 0.04 29 0 29 0 0
2 1.20-3.64 2.25 0.86 27 9 4 11 3
3 3.66-5.21 4.40 0.51 28 17 0 0 11
4 5.24-7.78 6.44 0.82 27 1 0 0 26

The mean LDI score for the 27 agriculture sites was 3.9 with a 0.7 standard

deviation. The lowest LDI score for agriculture sites was 2.4 and the highest score was









5.3. The mean LDI score for the 33 reference sites was 1.1 and the scores ranged from

1.0 to 1.8. The eleven silviculture sites had the smallest range of LDI scores, 1.5-1.8,

with a mean score of 1.6. Finally, the 40 urban sites had the largest range of LDI scores,

2.2-7.8, with a mean score of 5.7.

The 100m LDI scores were significantly correlated to many features of the

landscape that have the potential to influence amphibian and avian composition (Table 3-

7). Examples of the strongest correlations include the proportion of forest area

contiguous with, and including, the sample wetland within 200m (r=-0.71), evidence of

fire in the landscape (r=-0.75), and the amount of undeveloped land (excludes pine

plantation) within 200m of the focal wetland (r=-0.75). Significant but weaker

correlations included the proportion of roadway within 200m (r=0.48), distance to paved

road (-0.45), the distance to the nearest and second closest wetland (r=0.39, r=048),

wetland area within 200m (r=-0.40), fire evidence in the dome (r=-0.53), dissolved

oxygen (r=-0.41) and pH (r=0.45).

Amphibians

Amphibian Species Richness

A total of 23 amphibian species were detected during the 2003 field season giving

a total of 588 species by site detections. Four species, Ambystoma talpoideum, Bufo

woodhousiifowleri, B. marinus, Notophthalmus perstriatus, were encountered at only

one site, and two species, Rana capito aesopus and Scaphiopus holbrookii, were each

detected at only two sites.










Table 3-7. The Pearson correlation coefficients between the
and selected environmental variables.
Environmental Variable Pearsons p-value
% Undeveloped Land (200m) -0.71 0.00
% Roadway (200m) 0.48 0.00
% Canal/Retention Pond (200m) 0.34 0.00
% Wetland (200m) -0.40 0.00
% Contiguous Forest (200m) -0.75 0.00
Nearest Wetland (m) 0.39 0.00
2nd Nearest Wetland (m) 0.48 0.00
Distance to Paved Road (m) -0.45 0.00
Focal Wetland (ha) 0.19 0.05
Fire Evidence in Dome (1/0) -0.53 0.00
Fire Evidence in Landscape (1/0) -0.75 0.00
Latitude 0.02 0.83
Longitude -0.04 0.65
Julian Date 0.07 0.48
Water Dissolved Oxygen -0.41 0.00
Water Temperature -0.11 0.30
Water Ammonia 0.10 0.37
Water NO2NO3-N 0.13 0.22
Water TKN 0.11 0.30
Water Total Phosphorus 0.09 0.40
Water Turbidity 0.11 0.32
Water Color -0.04 0.74
Water pH 0.45 0.00
Water Conductivity 0.20 0.06
Maximum Water Depth -0.20 0.04
Fish Species Richness -0.13 0.21
Amphibian Species Richness -0.32 0.00
Insect Species Richness 0.09 0.40
Insect Count 0.11 0.33
Insect Biomass 0.03 0.76
Plant Richness 0.09 0.37
Basal Area 0.15 0.12
% Canopy Cover 0.23 0.07
% Evergreen Tree Species -0.07 0.48
% Broad Leaved Trees 0.23 0.02
% Exotic Trees 0.27 0.00


site's 100m buffer LDI score

N
111
111
111
111
111
111
111
111
111
111
111
111
111
111
87
87
87
87
87
87
87
87
87
87
111
90
111
82
82
82
109
111
63
108
108
108









Table 3-8 shows the amphibian species richness by land use type, LDI quartile,

and Florida region. Species detected using dip nets, cassette recordings and site visits

were included. Among four land use types, the mean species richness at reference sites

was significantly different than the mean species richness at urban sites. The minimum

and maximum number of amphibian species sampled at reference or silviculture sites

were three and ten, and at urban and agricultural sites the minimum was one and the

maximum was nine. Mean species richness for LDI quartile 1 was significantly greater

than the other three quartiles. However, LDI quartile 2 had the highest total species

richness. There was not a significant difference in the mean species richness by region.

The panhandle sites had the highest total species richness of the four sample regions.

Table 3-8. Amphibian species richness by land use, 100 meter LDI quartile, and Florida
region. Species richness was determined by using a composite richness of all
sampling methods (e.g., duration of the entire site visit, dip net sweeps, and
cassette recordings).
sites (n) mean (s.d.) total range
Reference 33 6.5 (2.0)a 20 (3-10)
Silviculture 11 5.3 (2.4)ab 16 (3-10)
Agriculture 27 5.3 (2.2)ab 18 (1-9)
Urban 40 4.2 (2.2)b 19 (1-9)

LDI Quartile 1 29 6.8 (2.0) 19 (3-10)
LDI Quartile 2 27 5.1 (2.2)a 20 (2-10)
LDI Quartile 3 28 4.7 (2.1)a 17 (1-8)
LDI Quartile 4 27 4.5 (2.4)a 18 (1-9)

Panhandle 27 6.0 (2.4)a 19 (3-10)
North 30 5.3 (2.3)a 18 (1-9)
Central 29 5.3 (2.4)a 15 (1-10)
South 25 4.5 (2.1)a 15 (1-8)

all sites 111 5.3 (2.3) 23 (1-10)
Note: Land uses with similar letter superscripts do not have significantly different means
(ANOVA, Tukey multiple comparisons, p<0.05).









Figures 3-1 and 3-2 show the mean amphibian richness by land use and LDI

quartile as detected by only using the two 15 minute surveys for each site. The pattern of

only the reference sites being significantly different than the urban sites is repeated as in

Table 3-8. The amphibian species richness decreases with increasing LDI quartile;

however, the relationship is only significant for LDI quartiles 1 and 4.

Amphibian Species Frequency of Occurrence

Table 3-9 shows the frequency of occurrence for all species by land use and all

sites combined corrected for species-specific ranges (i.e., known county occurrences,

(ARMI 2004)). Rana sphenocephala and Hyla cinerea were encountered most frequently

(64% of the sites) followed closely by H. squirella (60% of the sites). Silviculture sites

were only sampled in the panhandle and north Florida regions thus Bufo marinus and

Osteopilus septentrionalis, south Florida species, were not included in the frequency of

occurrence for this land use. The species' with the highest frequency of occurrence at

silviculture sites were H. femoralis, R. sphenocephala, and Acris gryllus (91%, 82%, and

73% of the sites respectively). Hylafemoralis and A. gryllus also had the highest

frequency of occurrence at reference sites (85% of the sites). Hyla cinerea, H. squirella,

and R. sphenocephala had the highest frequency of occurrence at both agriculture and

urban sites.












U)
a,
.c
o


ab
6 a T
T ab


I- J
U)
4-
0)
i 3



E
M 0-


a
TS


aI
Till


a
T-


a


r s


I inside the wetland U inside the wetland and the landscape

Figure 3-1. Mean and standard deviation of amphibian species richness within the
wetland and within the wetland and the 200m buffer combined by land use
type. Species richness was determined during two 15-minute surveys for each
site. Means that share letters were not significantly different between the land
use types (One-way ANOVA, Tukey multiple comparisons, p<0.05).



6

C5
a a a a



C,
--


E
0
Q1 2 Q3 Q4

I inside the wetland I inside the wetland and the landscape

Figure 3-2. Mean and standard deviation of amphibian species richness within the
wetland, and within the wetland and the 200m buffer combined by LDI
quartile. Species richness was determined during two 15-minute surveys for
each site. Means that share letters were not significantly different between the
LDI quartiles (p<0.05, One-way ANOVA, Tukey multiple comparisons).









Amphibian Breeding Effort

The mean number of tadpoles sampled at all sites with standing water during 20 dip

net sweeps was 9.6. However, the standard deviation was 15.1 even after excluding an

outlier that had approximately 1266 Scaphiopus holbrookii tadpoles collected during 20

sweeps (also excluded in Figures 3-3 and 3-4). The mean number of tadpoles and the

mean number of tadpole species was significantly different for only reference sites versus

urban sites (Figure 3-3). The mean number of species calling and their mean maximum

calling intensity were not significantly different between land uses. At all sites combined

the mean number of tadpole species, mean number of species calling, and their mean

maximum calling intensity were 1.7, 3.9, and 1.8 respectively.

LDI quartile one had the highest mean number of tadpoles, tadpole species, species

calling and mean maximum calling intensity (Figure 3-4). However, only the mean

number of tadpoles and the mean number of tadpole species found at quartile 1 versus

quartile 4 sites were significantly different (p<0.05).

Amphibian Species Composition

Table 3-10 shows the results of the MRPP test for significant differences in

amphibian species composition by dominant land use and LDI quartile. The amphibian

species composition was not significantly different between reference and silviculture

sites, and between agriculture and urban sites. All other between land use type

comparisons were significantly different at the p=0.01 level. The effect size or chance-

corrected within-group agreement (A) statistic shows the biggest difference in amphibian

composition between land use types was the reference and urban sites (A=0.122)

followed by reference versus agriculture sites (A=0.101). Results of the MRPP by LDI

quartile show that only the third and fourth quartiles did not have a significant difference












Table 3-9. Amphibian frequency of occurrence by land use type and all sites combined.a
Reference Silviculture Agriculture Urban All Sites
Acris gryllus 0.85 0.73 0.41 0.15 0.48
Ambystoma talpoideum 0.00 0.00 0.07 0.00 0.01
Bufo marinus 0.00 NA 0.00 0.11 0.06
Bufo quercicus 0.52 0.36 0.19 0.05 0.25
Bufo terrestris 0.06 0.09 0.22 0.28 0.18
Bufo woodhousiifowleri 0.14 0.00 0.00 0.00 0.04
Eleutherodactylus planirostris 0.06 0.09 0.30 0.35 0.23
Gastrophryne carolinensis 0.27 0.18 0.52 0.35 0.35
Hyla chrysoscelis 0.07 0.09 0.08 0.16 0.11
Hyla cinerea 0.58 0.27 0.81 0.68 0.64
Hylafemoralis 0.85 0.91 0.11 0.10 0.41
Hylagratiosa 0.39 0.18 0.11 0.00 0.16
Hyla squirella 0.52 0.36 0.74 0.65 0.60
Notophthalmus perstriatus 0.08 0.00 0.00 0.00 0.02
Notophthalmus viridescens 0.09 0.00 0.07 0.05 0.06
Osteopilus septentrionalis 0.13 NA 0.33 0.24 0.23
Pseudacris ocularis 0.69 0.50 0.16 0.18 0.35
Rana capitol aesopus 0.03 0.00 0.00 0.03 0.02
Rana catesbeiana 0.22 0.09 0.41 0.26 0.27
Rana clamitans 0.39 0.36 0.47 0.28 0.36
Rana grylio 0.52 0.18 0.30 0.10 0.29
Rana sphenocephala 0.58 0.82 0.67 0.63 0.64
Scaphiopus holbrookii 0.00 0.09 0.00 0.03 0.02
=species frequency of occurrence was calculated using only sites that were located in
counties with species-specific records of historical occurrence (ARMI 2004).

























tadpole count tadpole species species calling max intensity


Figure 3-3. The mean tadpole count, mean number of tadpole species, mean number of
anuran species calling and mean maximum intensity of anuran species calling
by land use. Land uses that share letters do not have significantly different
means (Mann-Whitney U-test, p<0.05). (Only sites with standing water and
where 20 dip-net sweeps were conducted were included)(Species calling and
mean maximum intensity based on sites with audio logger data only)


* Q1
* Q2
I Q3
I Q4


tadpole count tadpole species species calling max intensity


Figure 3-4. The mean tadpole count, mean number of tadpole species, mean number of
anuran species calling on cassette tapes, and mean maximum intensity of
anuran species calling by LDI quartile. Quartiles that share letters do not have
significantly different means (Mann-Whitney test p<0.05). (Only sites with
standing water and where 20 dip-net sweeps were conducted were
included)(Species calling and mean maximum intensity based on sites with
audio logger data only)









in amphibian species composition, and the negative A value indicates there was less

agreement within these groups than expected by chance. The largest significant

difference in species composition occurred between LDI quartile 1 and 4 followed by

LDI quartile 1 and 3, (A=0.150 and 0.135).

Figure 3-5 shows axis one and three of a nonmetric multidimensional scaling

(NMS) ordination using presence/absence data for all amphibian species found at two or

more sites in 2003. The final stress on the ordination was 17.5 and was stable after 183

iterations. The sites in Figure 3-5 are coded by their apriori land use classification.

Approximately 78% of the variance was represented with three NMS axes, and the

greatest proportion of variation was explained by axis three (35%)(Table 3-11).

The angle and length of the arrows in Figure 3-5 represent the relationship with the

environmental variables and the ordination scores. The environmental variables LDI and

FWCI were significantly correlated to the y-axis (axis 3) and the maximum site depth

was significantly correlated with the x-axis (axis 1)(Table 3-12). The correlation

between axis three and the independent variables FWCI (Reiss 2004) and the percentage

of sensitive plants were the strongest correlations, 0.77 and 0.78 respectively. Axis three

ordination scores had the strongest and the most significant (p<0.01) Pearson's

correlations with the site-specific independent variables (Appendix b). Axis one was

significantly correlated (Pearson, p<0.01) with maximum water depth, percent canal/ditch

within 200 meters, Julian date, survey length and latitude. The correlation between axis

one and water depth had the only coefficient greater than 0.5. Axis 2 (not shown) was

significantly correlated (Pearson, p<0.01) with amphibian richness and maximum water

depth, -0.37 and -0.25 respectively.












Table 3-10. MRPP test for significant differences in amphibian species composition
between land use types and LDI quartiles.
Variable A* T-statistic p-value
Land use
All land uses 0.109 -17.66 p<0.000
AvsR 0.101 -14.99 p<0.000
A vs S 0.094 -8.63 p<0.000
AvsU 0.001 -0.12 p=0.393
Rvs S 0.008 -0.81 p=0.189
RvsU 0.122 -22.85 p<0.000
S vs U 0.076 -9.76 p<0.000

LDI Quartiles
All quartiles 0.096 -15.74 p<0.000
1 vs 2 0.024 -3.34 p<0.006
1 vs 3 0.135 -18.72 p<0.000
1 vs 4 0.150 -21.02 p<0.000
2 vs 3 0.038 -5.39 p<0.000
2 vs 4 0.047 -6.71 p<0.000
3 vs 4 -0.004 -0.53 p=0.658
Land use abbreviations: A=agriculture, R=reference, S=silviculture, and U=urban.
* = Agreement statistic, chance corrected within group agreement.


The reference site with the lowest site score on the y-axis (NMS axis 3) had a rural

highway cutting through its southern boundary and was located in a relatively small (336

ha), isolated county park. Four of the five agricultural sites with positive y-axis scores

(NMS axis 3) were wetlands located on public conservation lands that were lightly

grazed. The urban sites with positive y-axis scores were wetlands adjacent to either

undeveloped land or very recent urban developments.











1.5 Axis 3
jFWCI


* x


x x

L x
x


LOI


aximum Depth agriculture
x reference
x silviculture
1.5
x urban
xx


-1.5


Figure 3-5. Unrotated joint plot of the amphibian based NMS ordination site scores for
axis one and axis three by land use with the strength and relationship of three
environmental variables (FWCI, LDI, and maximum depth). Axis 2 is not
shown.

Table 3-11. Coefficients of determination (R2) between amphibian based NMS
ordination distances and distances in the original space. Proportion of the
variance represented by each axis is listed incrementally and cumulatively.


IncrementCumulative
0.247 0.247
0.187 0.434
0.348 0.782


Axis 1

-1.5


x
x
^C~


Axis
1
2
3









Table 3-12. Pearson's correlation coefficients (r) and Kendall's Tau comparisons with
NMS ordination axes based on amphibian species composition and LDI,
FWCI, and maximum water depth.
LDI FWCI Maximum Depth
axis 1 r 0.028 0.005 0.521
tau 0.039 -0.013 0.379
axis 2 r 0.026 0.128 -0.246
tau 0.030 0.072 -0.178
axis 3 r -0.683 0.776 0.280
tau -0.478 0.570 0.223


Table 3-13. Independent variables included in stepwise regressions predicting the
amphibian based NMS axes scores.
Focal Wetland and
Landscape Sampling Attributes Water Chemistry
LDI % Broad Leaved Trees Dissolved Oxygen
1974 LDI % Evergreen Trees Total Phosphorus
% Agriculture (200m) % Exotic Trees pH
% Canal (200m) Basal Area Conductivity
% Roadway (200m) Fire Evidence in Wetland
% Silviculture (200m) Maximum Depth (cm)
% Undeveloped (200m) FWCI
% Urban (200m) Amphibian Species Richness
% Wetland (200m) Fish Species Richness
Distance to Nearest Wetland (m) Plant Species Richness
Distance to Paved Road (m) Julian Date
Focal Wetland Area Starting Time (survey 1)
Latitude Starting Time (survey 2)
Longitude Survey Length (min)

Table 3-14 shows the results of a stepwise regression using the landscape variables

(Table 3-13) to represent the NMS ordination scores based on amphibian composition at

the 111 sites. Axis 3 had the highest adjusted R2 (59.5) among the three axes. LDI,

percent urban, percent agriculture, distance to the nearest wetland, latitude and longitude

were all selected as Landscape predictors of axis 3. Axis 1 had a relatively low adjusted

R2 (16.3) and two variables, the area of the focal wetland and latitude, were selected.

There were no landscape variables that were correlated with axis 2 at the 0.05 level.









Table 3-14. The variables and regression coefficients selected (p<0.05) from a stepwise
regression using the Landscape variables listed in Table 3-13 to represent the
amphibian based NMS ordination axes.
Selected Variables axis 1 axis 2 axis 3
LDI -0.10
Focal Wetland Area 0.30
% Urban (200m) -0.79
% Agriculture (200m) -0.70
Distance to Nearest Wetland (m) 7.40E-04
Longitude 0.09
Latitude 0.14 0.15

Adjusted R-square 16.32 59.54


Table 3-15 shows the results of a stepwise regression using Landscape, Focal

Wetland and Sampling variables (Table 3-13) to represent the NMS ordination scores

based on amphibian composition at the 111 sites. Axis 3 had the highest adjusted R2

(70.9) followed by axis 1 (37.1) than axis 2 (27.3). When comparing Table 3-14 with

Table 3-15 the adjusted R2 of axis three improved by more then 10% with the addition of

focal wetland specific measures. None of the Sampling Effort variables were selected as

predictors of axis 3. The selected variables for axis 3 included; LDI, FWCI, proportion

agriculture (200m), distance to the nearest wetland, latitude and longitude. Maximum

depth, Julian date and latitude were selected as predictors of axis 1. Finally, predictors of

axis two were primarily wetland or sampling effort variables.

Table 3-16 shows the results of a stepwise regression using a total of 32 landscape,

focal wetland, sampling, and water chemistry variables (Table 3-13) to represent the

NMS ordination scores based on amphibian composition at 87 sites. Only the adjusted

R2 for axis 3 (73.6) improved from Tables 3-14 and 3-15 with the addition of the water

chemistry variables. Water conductivity and pH along with LDI and FWCI were selected









as predictors of axis 3. Axis 1 and 2 did not have any water chemistry parameters

selected as predictors.

Table 3-15. The variables and regression coefficients selected (p<0.05) from a stepwise
regression using the Landscape, Focal Wetland, and Sampling variables listed
in Table 3-13 to represent the amphibian based NMS ordination axes.
Selected Variables axis 1 axis 2 axis 3
LDI -0.09
FWCI 5.7 E-03 0.03
% Agriculture (200m) 0.34
Distance to Nearest Wetland (m) 7.0E-04
Maximum Depth (cm) 8.1E-03 -4.1 E-03
% Exotic Trees (focal wetland) -1.42
Amphibian Species Richness -0.09
Julian Date 6.6E-03
Longitude 0.06
Latitude 0.08 0.09 0.12

Adjusted R-square 37.08 27.34 70.87



Table 3-16. The variables and regression coefficients selected (p<0.05) from a stepwise
regression using the Landscape, Focal Wetland, Sampling, Water Chemistry
variables listed in Table 3-13 to represent the amphibian based NMS
ordination axes.
Selected Variables axis 1 axis 2 axis 3
Amphibian Species Richness -0.09 -0.08
Survey Length (min) 3.1E-03
Maximum Depth (cm) 5.7E-03
Latitude 0.09
% Agriculture (200m) -0.29
LDI -0.09
pH -0.14
Conductivity -5.0E-04
FWCI 0.01


Adjusted R-square


29.96 18.32


73.59









Amphibian Indicator Species

Figure 3-6 shows the same ordination as Figure 3-5, but with the species ordination

scores plotted on axes one and three. The species most associated with sites with high

FWCI scores and low LDI scores were Bufo quercicus, Hylafemoralis, H. gratiosa,

Pseudacris ocularis, and Acris gryllus. The two exotic species Eleutherodactylus

planirostris and Osteopilus septentrionalis along with native B. terrestris were most often

associated with sites that had low FWCI scores and had high LDI scores. Ranid species

along with Hyla chrysoscelis and Notophthalmus viridescens presence was associated

with deeper sites. Eleutherodactylus planirostris, Osteopilus septentrionalis, Hyla

squirella, and B. quercicus were associated with shallow or dry sites. The weighted score

of Gastrophryne carolnensis was the closest to the origin of axis 1 and 3. This suggests

G. carolinensis was not strongly associated with land use intensity or water depths.

Results of indicator species analyses reflect the species pattern observed in Figure

3-6. The majority of the significant indicators of LDI quartiles (Table 3-17) and land use

(Tables 3-18 and 3-19) tend to be indicators of reference sites or low intensity sites (e.g.,

Acris gryllus, Bufo quercicus, H. femoralis, H. gratiosa, Pseudacris ocularis, and Rana

grylio). When LDI quartiles three and four were lumped or agriculture and urban sites

were lumped (Tables 3-17 and 3-19) Eleutherodactylus planirostris, B. terrestris,

Gastrophryne carolnensis, and H. squirella had their highest indicator value at these

higher land use intensity sites. The exotic Osteopilus septentrionalis had increasing

indicator values with increasing land use intensity, however, the pattern was not

significant (p=0.15). An additional 10 amphibian species detected in 2003 were not

significant indicators of land use or development intensity.