Wetland Habitats of North America

CHAP TER 15

Southeastern Depressional Wetlands L . K ATHERINE KIRKMAN, LOR A L . SMITH, and STEPHEN W. GOLL ADAY

Isolated depressional wetlands are common landscape features throughout the Coastal Plain of the southeastern U.S. Broadly defined, they are wetlands that are surrounded by uplands and lack surface water inputs and outputs (Tiner 2003). They occur in topographic depressions ranging from less than 1 ha to several square kilometers, and their hydrology is dominated by precipitation inputs and evapotranspiration losses. Variations among these wetlands are reflected in many regional colloquial names, such as Carolina bays, grady ponds, limesink ponds, cypress domes, oak domes, citronelle ponds, gum ponds, seasonal ponds, and flat-bottom ponds. In the Southeast (and similarly in other regions), depressional wetlands have many of the structural and functional attributes associated with other types of wetlands; however, their ecological role has been greatly undervalued, particularly in regard to the important habitats they provide (National Research Council 1995). Ironically, the absence of a clear surface-water connection contributes to the uniqueness of these wetland habitats; yet this defining feature has also played a role in society’s failure to recognize and protect the ecological services associated with them. Often these wetlands support unique suites of species, including several rare species of conservation concern. Relative to their collective area in the region, isolated depressional wetlands contribute disproportionately to the maintenance of biodiversity (both fauna and flora) (Whigham 1999), particularly when they are embedded within minimally disturbed uplands or buffer zones (Kirkman, Goebel, et al. 2000). Despite this fact, depressional wetlands do not receive the federal protection afforded to other wetland types. Even less awareness exists of the upland-wetland connectivity of ecosystem functions and delivery of ecosystem services. The reason for this may be that these values are less obvious than in other types of wetlands and are not easily measured because the wetlands are integral parts of the landscape (Ewel 1990a; Whigham 1999). For some depressional wetlands in the Southeast, little is known about their hydrologic connectivity to surface flows through subsurface or short-duration surface waters or to groundwater, or even their potential role in regulation of nutrient assimilation processes in agriculturally impacted landscapes (Ewel 1990a; Whigham and Jordan 2003).

In this chapter we characterize the physiognomic variation of depressional wetlands in the region. We emphasize the importance of seasonally fluctuating water levels and frequent fire as drivers of the structural and functional diversity of depressional wetlands, as well as the primary factors controlling the notable similarities in appearance and constituent species that occur across the region. Finally, we identify conservation concerns and information needs that are critical to the development of strategic management and protection plans of these vulnerable southeastern wetlands.

Geology, Hydrology, and Biogeochemistry Distribution and Geomorphology The origins of depressions vary across the region depending on topographic position and underlying geology. In karst regions of southwestern Georgia, southern Alabama, and northern Florida, irregular depressions called limesinks appear to result from solution processes in the underlying limestone bedrock. As mildly acidic rainwater infiltrates soils, it dissolves the calcium carbonate and creates underground voids. The overlying land surface eventually subsides, and an accumulation of sand and clay results in impermeable soil horizons that lead to the ponding of water and a semi-perched water table (Hendricks and Goodwin 1952; Lide, Meentemeyer, et al. 1995). In southeastern North Carolina and throughout the Coastal Plain of South Carolina, depressional wetlands that have a distinctive elliptical shape and northwest-southeast orientation are called Carolina bays and are particularly abundant. Based on their position on sandy, relatively undissected surfaces, the geomorphologic origin of these depressions is attributed to unidirectional wind on water in surface depressions (Soller and Mills 1991). Although numerous wetland depressions exhibit the elliptical shape ascribed to Carolina bays, others within the same range are less well defined geomorphically (Lide 1997). Irregularly ovoid depressions that lack consistent orientation occur throughout the southeastern Atlantic and Gulf Coastal Plain westward to Alabama; these may have originated from dissolution of underlying weathered clay or iron203

oxide cementing materials that resulted in volume loss and soil settling (Folkerts 1997).

Soils Isolated depressional wetlands are situated within Coastal Plain uplands on deep, well-drained sand and loamy sand soils to poorly drained upland soils. Characteristically, these wetlands have relatively impervious clay or sandy clay layers beneath the basin that restricts surface-water and groundwater interactions. The permeability and infiltration rate of these sediments depend on the soil texture and depth. Commonly, in the interior of the depression this aquitard is composed of alternating clay and sand layers several meters deep and is thinner toward the upland margin (Lide, Meentemeyer, et al. 1995). Although the clayey hardpan may remain saturated throughout the period of inundation (Lide, Meentemeyer, et al. 1995), unsaturated layers of clay and sand beneath saturated zones also commonly occur (West, Shaw, et al. 1998). Surficial soils above the clayey layers are extremely variable among wetlands and may range from sandy substrates to organic layers several meters deep (Bliley and Pettry 1979; Stolt and Rabenhorst 1987; Newman and Schalles 1990). A distinction is often made between “peat-based” wetlands and “clay-based” wetlands, reflecting the accumulation of organic materials (Sharitz 2003). In a regional transect from the upper Coastal Plain to the Atlantic Coast, Newman and Schalles (1990) reported a trend of greater peat depth toward the lower Coastal Plain, particularly in more poorly drained upland soils. The degree of organic soil or peat accumulation depends on hydroperiod, nutrient levels, and fire. Peat buildup occurs when primary production exceeds decomposition of the peat substrates (Clymo, Turnen, et al. 1998). Litter decomposition is particularly slow if pH and dissolved oxygen (DO) are low (Ewel 1990b). In wetlands that frequently dry down, organic soils typically do not accumulate, because plant litter rapidly decomposes when substrates are exposed to oxygen (e.g., Battle and Golladay 2007). Also, in wetlands that frequently dry, fire often plays an important role in reducing organic matter accumulation (Schalles and Shure 1989; Ewel 1990b; Sharitz and Gresham 1998; Craft and Casey 2000). Geomorphic controls of wetland processes and community development include the topographic setting and the associated soils of the uplands surrounding the depressions. Evidence that depressional wetlands located in interridge depressions of sandy uplands or in former floodplain terraces tend to have finer-textured soils and longer hydroperiods suggest that topo-edaphic conditions drive variations in hydrologic regime and potential fire frequency (Kirkman, Goebel, et al. 2000; De Steven and Toner 2004). Similarly, cypress swamps or shrubbogs located in the flatwoods of the lower Coastal Plain are surrounded by seasonally wet Spodosols that affect uplandwetland hydrologic interactions (Richardson 2003; Sun, Callahan, et al. 2006).

Hydrology The hydrologic regime of depressional wetlands is strongly driven by precipitation, evapotranspiration, basin morphometry, and topographic position. Thus, timing and duration of inundation vary widely among wetlands, even when in close geographical proximity. Water levels in the wetlands fluctu204 Inland Wetlands

ate seasonally and among years, largely depending on precipitation. Hydroperiod ranges from semipermanent inundation with dry down occurring only during prolonged periods of drought to wetlands that dry nearly every year (Dierberg and Brezonik 1984a; Mitsch 1984; Schalles and Shure 1989; Lide, Meentemeyer, et al. 1995; Kirkman, Goebel, et al. 2000; De Steven and Toner 2004). Seasonal patterns of inundation occur, with filling in winter and spring when precipitation exceeds evaporative water loss. With increasing temperatures and vegetation growth responses, gradual dry down occurs in summer with increased evapotranspiration. Inundation may occur rapidly in response to large rainfall events (Sharitz 2003); however, there may be considerable lag time between change in prevailing weather and change in water level, depending on antecedent wetland water levels (Lide, Meentemeyer, et al. 1995). While hydrologic connectivity of depressional wetlands to groundwater is often assumed to be minimal because of the relatively impervious basin materials, several studies have demonstrated hydrologic coupling during certain conditions that depend on water table depth, subsurface topography, and the permeability of clay layers beneath the wetland. For example, in some depressional wetlands of southwestern Georgia, Hendricks (1954) found that in wet conditions when groundwater was high, lateral seepage into the wetland occurred. Alternatively, when groundwater level was below the ponded-water level, a hydraulic gradient developed and the rate of seepage out of the pond was influenced, provided that the substrate permitted hydraulic connectivity. Hendricks (1954) also demonstrated a water table mounding effect beneath the basin of some wetlands when groundwater levels were lower than the basin of the wetland, but in such conditions, the groundwater level had no control on the rate of seepage. Similar interpretations were drawn from a water budget study in an isolated depressional wetland in South Carolina (Lide, Meentemeyer, et al. 1995). Pyzoha, Callahan, et al. (2008) developed a conceptual model describing hydrology of depressional wetlands, suggesting that during years of above-normal rainfall, surficial groundwater moves laterally into the wetland from the catchment basin associated with the depression. When precipitation and inundation conditions change, such as at the initial onset of drought or at the end of a drought, this gradient is reversed. Shallow groundwater transfers between depressional wetlands and extensive shrub-bog flats are presumed to take place, although quantification of the degree of hydrologic connectivity is lacking (Richardson 2003). In the upper Coastal Plain, ephemeral surface-water connections between depressional wetlands and streams during extreme storm events are also possible (Heimberg 1984). As depressions fill, they act as corridors for surface-water movement. For example, in 1994 and 1998, southwestern Georgia experienced greater than 100-year floods due to tropical storms, and some depressional wetlands became briefly linked to other depressions as well as a nearby fifth-order creek (Michener, Blood, et al. 1998; Battle and Golladay 2002). In such events, the surface flow of water likely serves as a mechanism for nutrient pulses into wetlands, as well as for propagule dispersal between wetlands.

Biogeochemistry and Nutrient Cycling Water chemistry in depressional wetlands is influenced by many factors, including hydroperiod, vegetation, position within the

landscape, surrounding land use, fire, and relative contribution of shallow groundwater versus precipitation. Although they exhibit heterogeneity in water chemistry, most southeastern depressional wetlands derive nutrients primarily from precipitation, and are usually acidic and nutrient poor. Across a regional transect of depressional wetlands in North Carolina and South Carolina, Newman and Schalles (1990) reported a median pH value of 4.6. Similar acidic conditions occur in depressional wetlands in Florida (Dierberg and Brezonik 1984a), whereas somewhat higher average pH has been reported in wetlands of southwestern Georgia (Battle and Golladay 2001). Among wetlands of differing vegetation types, Battle and Golladay (2001) reported similar water chemistry characteristics at the initial period of inundation, presumably reflecting precipitation. However, over the course of the hydroperiod, closed-canopy wetlands differed significantly from cypress savannas or grass-sedge marshes, particularly in maintaining higher levels of PO4 -P, benthic organic matter, and organic C. As water levels receded, NH4 -N and organic C levels increased in the swamps and the water became darkly stained. Differences in water chemistry among wetlands are probably attributable to larger amounts of litterfall and wood inputs into the forested sites relative to those wetlands that are dominated by herbaceous vegetation. Similar seasonal patterns in ammonia have been reported in cypress ponds in Florida when the diluting effect of rainfall was not present (Mitsch 1984). Surrounding land use is a potential factor influencing water quality of depressional wetlands. Because they are topographically low, these wetlands appear to be sinks for sediment and nutrients (Craft and Casey 2000; Whigham and Jordon 2003). Relative to minimally altered sites, elevated N, P, and suspended sediments have been observed in wetland waters located within intensive agricultural areas in Florida (Reiss 2006) and in southwestern Georgia (Battle and Golladay 2007; Atkinson, Golladay 2011). The degree of nutrient retention or transport to surface streams or groundwater is not well documented and appears to depend on the specific hydrologic characteristics of the depressional wetland (Whigham and Jordan 2003), as well as intensive storm events (Battle and Golladay 1999). Some studies of Florida cypress swamps receiving wastewater indicated no significant hydrologic or nutrient connections of the depressional wetland with downstream systems. In this case, nearly all of the introduced organic matter and nutrients was retained through soil processes (Deghi, Ewel, et al. 1980; Dierberg and Brezonik 1984b). However, in a similar study in which intermittent surface flows occurred from a cypress pond to a downstream wetland, export of wastewater nutrients to the downstream wetland were observed (Nessel and Bayley 1984). Regardless of vegetation type, net primary productivity in depressional wetlands is generally thought to be P-limited (Ewel 1990b; Koerselman and Mueleman 1996). In addition to low input of nutrients, acidic conditions in depressional wetlands make nutrients unavailable for plant uptake, partly because of slow decomposition rates of dead organic material, and also because P becomes bound with iron and aluminum. Craft and Chiang (2002) found that plant available N (nitrate), organic N, and total N were greater in wetlands than in longleaf pine–wiregrass uplands and that C:N increased from wetland to upland soils. Even though total P was greater in wetland than in upland soils, most of it was in recalcitrant organic forms. They concluded that periodic inundation results in retention of organic N and P in soil, with greater retention of N than of P. This process results in a shift from N limitation  

of plant growth in uplands toward P limitation or N and P colimitation in wetlands. Some comparisons of productivity in Florida wetlands suggest that relative to swamps receiving flowing water from streams or rivers, depressional cypress swamps that were supplied primarily by rainfall had lower net primary productivity (Brown 1981). However, generalizations regarding the influence of nutrient inputs versus inundation period on decomposition rates in various wetlands are difficult to make, particularly because these roles may change with prolonged drought or prolonged inundation, as well as type of plant material (Brinson, Lugo, et al. 1981; Battle and Golladay 2007). Water level fluctuations may diminish nutrient limitation by encouraging decomposition and nutrient mineralization processes (Brinson, Lugo, et al. 1981). For example, in depressional wetlands with highly fluctuating water levels in Georgia and South Carolina, Watt and Golladay (1999) and Busbee, Conner, et al. (2003) reported litterfall values that were similar to those of alluvial river systems. The degree to which fire influences overall nutrient balances within a wetland is not well understood. In general, fire consumes much aboveground biomass and litter, increasing the rate of nutrient turnover. However, the type of vegetation, soils, timing of rainfall, intensity of fire, and hydrologic regime also influence nutrient turnover and availability. Wilbur and Christensen (1983) found that following burning in shrub-bog wetlands, a considerable enrichment of nutrients occurred, including Mg, K, PO4 -P, NH3-N, and NO3-N. It was unclear as to whether the increase in nitrate in burned peat was a consequence of ash addition, reduced plant uptake, or changes in rates of nitrification and denitrification. Nutrient volatilization associated with fires that consume vegetation in herbaceous wetlands with mineral soils probably results in losses of N and C due to combustion, but little loss of P, similar to that reported in adjacent upland longleaf pine–wiregrass forests (Boring, Hendricks, et al. 2004). Nutrient subsidies associated with fire may also be attributable to the transport of ash into the wetlands, primarily when coupled with rainfall (Battle and Golladay 2003). Changes in soil nutrient concentrations have been examined relative to land uses surrounding depressional wetlands. A consistent pattern of increased total P concentration in soils of wetlands impacted by agricultural runoff relative to reference sites has been reported, although seasonal fluxes have not been thoroughly examined (Paris 2005; Reiss 2006). Craft and Casey (2000) attributed declines in rates of soil P accumulation over the last 100 years to decreased anthropogenic disturbances in a group of depressional wetlands located within landscapes currently managed for conservation of longleaf pine forests. Although they reach tentative conclusions, these studies suggest that P accumulation in wetland soils may be a sensitive metric that could be developed into a useful indicator of depressional wetland condition.  

Vegetation Vegetation of isolated depressional wetlands varies widely because of the spatial and temporal variation in hydrologic regimes, topographic settings, and the influence of fire. The hydrologic regime directly influences vegetation composition by filtering out species based on their tolerance of inundated or dry conditions. Hydroperiod also regulates fire frequency and intensity when prescribed fire management is occurring on Southeaster n Depressional Wetlands 205

A

B

C D

FIG. 15.1. Vegetation types in depressional wetlands. A. Grass-sedge marsh. B. Cypress savanna. C. Cypress-gum swamp. D. Shrub bog. (A–C: Photos from J. W. Jones Ecological Research Center archives; D: Photo from Hugh and Carol Nourse.)  

the landscape. During dry periods, both upland and wetland sites will burn provided that fire fuels are available, but they are not susceptible to fire when inundated. Thus, during dry conditions, fire removes species intolerant of fire. Regional fire suppression and landscape fragmentation have significantly altered the influence of fire in plant community succession in many to most depressional wetlands. Differences in plant community development associated with topographic position and geomorphology are likely linked through the influence of these physical differences on hydrologic regime (Kirkman, Goebel, et al. 2000; De Steven and Toner 2004; Stroh, De Steven, et al. 2008). Plants inhabiting depressional wetlands must cope with anaerobic conditions when soils are saturated or inundated, partial submergence of leaves during inundation, periods of dry soil or even drought conditions, and, potentially, fire. In addition to seed dormancy characteristics, other life-history traits that contribute to success in wetland environments include seed dispersal by animals or water, nutritional value to waterfowl, first-year reproductive maturity of perennials, and flood-induced petiole elongation (Kirkman and Sharitz 1993).

Plant Communities Numerous distinct plant assemblages have been described for Coastal Plain depressional wetlands (Sharitz and Gresham 1998; Sharitz 2003). Structurally, this diverse vegetation can 206 Inland Wetlands

be grouped into three major categories: open-canopied herbaceous-dominated communities, closed-canopy swamps, and shrub-bogs (Fig. 15.1a–d).  

OPEN-CANOPIED HERBACEOUS-DOMINATED COMMUNITIES

Grass-sedge marshes and cypress savannas occur primarily in wetlands having sandy surficial soils over clay that frequently experience dry-down conditions. The vegetation is composed of a species-rich herbaceous flora. Dominant species are often distributed in distinct zonal patterns in response to water depth, fire, or even historical anthropogenic disturbances (Tyndall, McCarthy, et al. 1990; Sharitz and Gresham 1998). Floating species dominate in the deepest part of the depression or across entire wetlands with semipermanent hydroperiods. Emergent grasses and sedges often occur in more intermediate water conditions, and woody shrubs may be present along the wetland edge. Droughts, or periods of greater than average precipitation, may result in shifts in species zones (Stroh, De Steven, et al. 2008). Collins and Battaglia (2001) concluded that zonation is less distinct in wetlands with greater fluctuations in hydroperiod and less gradient in slope. Several interacting factors contribute to directional or cyclical patterns of change in vegetation over time in grass-sedge marshes. Based on chronosequences of aerial photographs, many depressional marshes appear to be temporally stable if

FIG. 15.2. Conceptual model of ecosystem development in depressional wetlands. Drivers (stable

physical features of depression that control hydrologic and fire-disturbance regimes) and filters (climatic and disturbance factors that control the establishment of species) are identified in black boxes. Bold arrows from these boxes indicate the resulting environmental conditions or vegetation from each influencing factor. Abbreviations of vegetation: H (herbs), C (cypress), G (gum), HW (other hardwoods), P (pine). Resulting depressional wetland vegetation is indicated in lowest boxes. (From Kirkman, Goebel, et al. 2000.)

they have escaped direct hydrologic alterations (Kirkman, Lide, et al. 1996; Stroh, De Steven, et al. 2008). Dominance by herbaceous vegetation may be maintained for many decades without encroachment of hardwoods or pond cypress (Taxodium ascendens). Because fire and inundation both inhibit establishment of woody species, the absence of one factor may be offset by the presence of the other, and, consequently, change in the plant community can be very slow. In depressions with prolonged or semipermanent inundation, hydrology alone may exclude woody establishment. Alternatively, extended dry periods in the absence of fire can result in rapid shrub and tree encroachment in which the size of the wetland, the land-use history, or the landform may be factors (Kirkman, Goebel, et al. 2000; Stroh, De Steven, et al. 2008) (Fig. 15.2). In depressional marshes with irregularly fluctuating water levels, a persistent seed bank is usually present and guilds of species adapted to germination under different hydrologic conditions have been identified (Kirkman and Sharitz 1994; Poiani and Dixon 1995; Collins and Battaglia 2001; Mulhouse, Burbage, et al. 2005). Some species are adapted to drought and germinate in exposed soil conditions, others are stimulated to germinate only when inundated, and some are generalists and can germinate in both flooded and dry soil conditions (Hook 1984). The species richness of persistent seed banks in grass-

sedge depressional wetlands is among the highest of freshwater wetlands in North America (Kirkman and Sharitz 1994). Cypress savannas are similar to marshes floristically, but have an open to sparse canopy (25– 50% cover) of pond cypress. For cypress to become established, the wetland must be dry long enough for seeds to germinate (Demaree 1932) and seedlings must attain a height that will not be submerged with inundation, as well as attaining a size that can withstand fire. Once established, mature pond cypress trees are relatively fire tolerant and extremely flood tolerant (Ewel 1995, 1998). Evenaged stands of cypress may reflect past timber harvest or may be the result of episodic recruitment events. With prolonged dry down and absence of fire, cypress savannas can become invaded by hardwoods, particularly oaks (Quercus spp.) and swamp black gum (Nyssa biflora) (Ewel 1998). Vegetation of some shallow depressional wetlands, particularly those at the dry end of the hydrologic gradient, succeed from herbaceous-dominated marshes to mesic hardwood communities with prolonged fire exclusion (DeSteven and Toner 2004). Martin and Kirkman (2009) describe such depressional wetlands in southwestern Georgia as islands of evergreen and semievergreen hardwoods, particularly flood-tolerant live oaks (Q. virginiana), water oaks (Q. nigra), and laurel oaks (Q. laurifolia) within a matrix of fire-maintained longleaf pine  

Southeaster n Depressional Wetlands 207

FIG. 15.3. Generalized model for succession in medium-depth basins in north-central Florida. (From

Casey and Ewel 2006.)

uplands. The largest and oldest trees usually occur in the deepest part of the depression, where these fire-intolerant species can become established. Based on aerial photography and the presence of a persistent seed bank of obligate wetland herbaceous species, the presence of mesophytic oaks suggests that during dry periods they become established and are then protected from fire when the wetlands are ponded or have saturated soils. The accumulation of fire-resistant oak leaves and low light conditions result in little or no shrub layer or herbaceous ground cover. The layer of oak leaves also holds moisture and, consequently, the leaf litter becomes increasingly less prone to carry fire. Thus, over time, as fire is prevented from being carried beneath the oaks, additional oaks become established and the oak-dominated patch encroaches outward toward the upland. As a consequence, these wetlands remain in a persistent alternate state as a hardwood-dominated community because fire exclusion is perpetuated (Martin and Kirkman 2009).

CLOSED-CANOPY SWAMPS

Cypress-gum swamps tend to develop in depressions with longer hydroperiods than those of cypress savannas or grasssedge marshes and may even be semipermanently inundated. These depression swamps occur throughout the Coastal Plain and are generally located in interridge landscape positions or floodplain terraces in the upper Coastal Plain (Kirkman, Goebel, et al. 2000; De Steven and Toner 2004) or are surrounded by pine flatwoods in the lower Coastal Plain and peninsular Florida (Ewel 1998; Casey and Ewel 2006). Surface soils are usually highly acidic organic mucks and peats. Because of the longer hydroperiod, fires are typically less frequent in these wetlands than in cypress savannas or grass-sedge marshes. The canopy composition of cypress-gum swamps ranges from monospecific stands of either pond cypress or swamp black gum to a mixture of both species. The dominance of one species over the other may reflect chance historical establishment conditions and fire events, or even timber harvest patterns in which cypress was selectively removed. Mixtures 208 Inland Wetlands

of slash pine (Pinus elliottii) or loblolly pine (P. taeda) and red maple (Acer rubrum) are also frequently present, as well as numerous shrub species. Characteristic assemblages vary across the region (Newman and Schalles 1990; Folkerts 1997; Edwards and Weakly 2001; Sharitz 2003; De Steven and Toner 2004). In some cypress wetlands, the trees form a domelike profile, with taller, older, and more fire-protected trees in the center and shorter, younger trees on the edges. These are often called cypress domes or gum domes (Ewel 1998). A cypress dome with an open center is sometimes referred to as a cypress doughnut. The closed canopy and frequent inundation of these wetlands result in sparse, herbaceous ground cover. Because pond cypress is more fire tolerant than swamp black gum, more frequently burned sites may be maintained as monospecific stands of cypress (Ewel and Mitsch 1978). Protection from fire, in addition to commercial harvesting of cypress, increases the dominance of hardwoods (Ewel 1990b; Casey and Ewel 2006). De Steven and Toner (2004) examined environmental factors influencing vegetation in depressional wetlands in South Carolina and suggested that hardwoods that colonize following prolonged drought or altered hydrologic conditions depend in part on the forest composition of the adjacent uplands and proximity to other wetlands. Additionally, Kirkman, Lide, et al. (1996) observed that vegetation of smaller wetlands in South Carolina with a history of cultivation has a high probability of developing into mixed stands of flood-tolerant hardwoods after several decades of abandonment from agriculture. Casey and Ewel (2006) examined successional relationships among depressional wetlands in northcentral Florida based on depth of basin and accumulation of organic materials. They reported wider fluctuations in hydroperiod in shallow-basin swamps, and noted that under certain combinations of fire and hydroperiod, slash pine becomes established. They reasoned that fire results in mineral soil for seed germination, but successful establishment requires a subsequent prolonged period with the absence of fire and inundation. In medium-basin swamps, they proposed successional pathways linking cypress or cypress-gum swamps to shrubbogs through a combination of logging, severe fire, and drainage (Fig. 15.3).

SHRUB-BOGS

Shrub-bog vegetation develops in depressional wetlands with peat or sandy peat soils, with long hydroperiods of 6–12 months and fire return intervals of 20– 50 years (Christensen 1988; Casey and Ewel 2006). This vegetation type most commonly occurs in the lower Coastal Plain (Sharitz and Gresham 1998; Richardson 2003; Laliberte, Luken, et al. 2007) or in Florida (Richardson and Gibbons 1993). The vegetation is characterized by dense stands of shrubs such as loblolly bay (Gordonia lasianthus), red bay (Persea palustris), sweet bay (Magnolia virginiana), and other shrub species. Pond pine (Pinus serotina) or pond cypress are occasionally scattered as canopy emergents (Christensen, Burchell, et al. 1981; Sharitz and Gresham 1998). Overall species richness in shrub-bogs is much lower than that of other types of depressional wetland vegetation (Laliberte, Luken, et al. 2007). Otte (1981) suggested that the shrub-bog vegetation develops with increased peat accumulation and in response to nutrient limitation. However, fire, hydroperiod, and nutrients are intricately linked (Christensen 1988). One interpretation of the occurrence of a mix of marsh and shrubbog vegetation is a result of severe fires that burn out deep layers of peat in shrub-bogs (Richardson 2003; Casey and Ewel 2006).  

 

Animal Communities Southeastern depressional wetlands support a remarkably diverse fauna, particularly among aquatic invertebrates (Mahoney, Mort, et al. 1990; Golladay, Entrekin, et al. 1999; Battle and Golladay 2001) and amphibians (Moler and Franz 1987; Dodd 1992; Liner, Smith, et al. 2008). Some animals use depressional wetlands opportunistically, whereas others have specialized life-history characteristics that allow them to persist in temporary aquatic systems (Wiggins, Mackay, et al. 1980; Semlitsch and Ryan 1999). The distinctive fluctuating hydroperiods of depressions limit the presence of predatory fish and many invertebrates and amphibians are adapted to the fishless conditions in these wetlands (Semlitsch 2000).

Invertebrates Depressional wetlands support diverse communities of invertebrates. Arthropods, in particular, tend to be very diverse and abundant. Major groups of arthropods include aquatic insects and crustaceans of subclasses Branchiopoda, Copepoda, Ostracoda, and Malacostraca (Taylor, Leeper, et al. 1999). Generally, copepods, cladocerans, and ostracods are very small taxa ( 500 μm at maturity) (Taylor, Leeper, et al. 1999). Other invertebrate groups observed in depressional wetlands include annelids, nematodes, rotifers, mollusks, bryozoans, sponges, hydrozoans, tardigrades, turbellarians, and water mites (Taylor, Leeper, et al. 1999). Invertebrate assemblages of depressional wetlands vary with season, hydroperiod, and wetland vegetation type (Golladay, Taylor, et al. 1997; Battle and Golladay 2001). Invertebrate abundance is greatest early in the hydroperiod, reflecting the rapid response of taxa capable of surviving the dry period in these habitats. Early in the hydroperiod, predator abundance tends to be low (Schneider and Frost 1996; Moorhead, Hall, et

al. 1998), and physical conditions, such as moderate temperature and abundant DO, are optimal for invertebrate growth (Wiggins, Mackay, et al. 1980; Sklar 1985). After prolonged flooding or dry down, many early-hydroperiod taxa exhibit physiological or behavioral adaptations (migration or production of a resistant life-history stage) (Sklar 1985; Dietz-Brantley, Taylor, et al. 2002) to avoid stressful conditions. The most taxonomically diverse groups of invertebrates in depressional wetlands are Diptera, Coleoptera, Hemiptera, and Odonata (Battle and Golladay 2001; Entrekin, Golladay, et al. 2001); these are primarily migrants that colonize wetlands and oviposit in response to seasonal flooding. Some invertebrate taxa, especially the relatively abundant crustaceans, live in seasonally flooded depressional wetlands but are seldom found in nearby permanently flooded aquatic habitats (Wiggins, Mackay, et al. 1980). This is often attributed to the absence or low numbers of predaceous fish. As with plant communities, marshes have greater diversity and density of aquatic invertebrates than other depressional wetland vegetation types (Battle and Golladay 2001). This may be due to a combination of habitat complexity and trophic diversity. Emergent plants provide both food and habitat for invertebrates, particularly taxa capable of clinging to plants. While few aquatic invertebrates eat living aquatic vascular plants (Newman 1991), decomposing macrophytes, periphyton, and planktonic microorganisms provide high-quality food resources for these organisms (Golladay unpublished data). Closed-canopy depressional wetlands have abundant detrital food resources from the forest canopy, and species assemblages are dominated by detritivorous crustaceans early in the hydroperiod. However, invertebrate abundance and diversity declines later in the hydroperiod, possibly due to low DO concentrations (Golladay, Entrekin, et al. 1999; Battle and Golladay 2001).

Amphibians Most amphibians that use depressional wetlands spend much of their life in adjacent uplands and migrate to wetlands to breed. These species are able to persist in upland habitats during the nonbreeding season by adopting a largely fossorial lifestyle. However, to support an aquatic larval stage, they must lay their eggs in water or in a moist environment that will ultimately be inundated. The length of the larval stage ranges from a few weeks in some frog species to 6 – 8 months for a few frogs and most salamanders. A few fully aquatic amphibians, the dwarf salamander (Pseudobranchus striatus), two-toed amphiuma (Amphiuma means), and siren (Siren spp.), persist in depressional wetlands during dry downs by aestivating in moist substrate or burrows of other animals. Some amphibians breed early in the hydroperiod, presumably to minimize competition and predation, whereas others breed later in the hydroperiod and have other survival strategies. Species such as the pinewoods treefrog (Hyla femoralis) breed only in wetlands that have dried in the previous year and are thus unlikely to have populations of predatory fish (Pechmann, Scott, et al. 1989). Amphibians that breed in depressional wetlands have exceptionally high fecundity and are believed to be longlived, traits that enable them to produce large numbers of offspring when conditions are appropriate (Gibbons, Winne, et al. 2006). A single female eastern spadefoot toad (Scaphiopus holbrookii), for example, can lay > 2,000 eggs at one time and can breed multiple times in one year (Wright 1932). This spe 


Terrestrial adult Eft

Paedomorph

Larva

Eggs

Aquatic adult

FIG. 15.4. Life history of the striped newt (Notophthalmus perstriatus). Adult striped

newts migrate to depressional wetlands to breed. Eggs are laid in the wetlands and hatch into aquatic larvae. As wetlands dry in late spring or summer, larvae transform into terrestrial juveniles called efts. Efts mature in the uplands, after which they return to wetlands to breed, repeating the cycle. In very wet years, if wetlands hold water through the following fall and winter (dashed arrow), larvae may remain in the wetland, retain their aquatic morphology (gills and tail fin), and become sexually mature. The sexually mature larval forms are called paedomorphs. (From Johnson 2001.)

cies also has a short larval period (2–3 weeks), which allows the larvae to take advantage of highly ephemeral water bodies that support fewer predators. One of the most dramatic examples of adaptations to the unpredictable breeding environment of depressional wetlands is that of the striped newt (Notophthalmus perstriatus), which, in addition to being long-lived and highly fecund, is also paedomorphic (Dodd 1993). Paedomorphism is a phenomenon in which an organism with a multistage life cycle becomes reproductively mature while retaining some larval characteristics (Fig. 15.4) (Johnson 2002). Eggs hatch into fully aquatic larvae with external gills for respiration and tail fins for aquatic locomotion. As the wetland dries down in late spring or summer, the larvae transform (i.e., external gills are absorbed and lungs develop to breathe air) into terrestrial juveniles called efts that live a fossorial existence in the uplands until they reach sexual maturity, and eventually migrate overland back to wetlands to breed. However, in years when a wetland holds water through the summer, striped newts may remain in the wetland, retain their external gills while their reproductive organs develop, and then breed the following fall and winter. This complex life history allows the species to mature quickly and breed immediately in extremely wet years, thus avoiding the risk of the additional migration. Landscape connectivity among and between depressional wetlands, terrestrial habitats, and permanent aquatic systems (e.g., streams and rivers) is critical to maintaining the diverse amphibian fauna of depressional wetlands. For example, some amphibians that breed in depressional wetlands migrate considerable distances through uplands to nonbreeding habitat (Gibbons 2003; Semlitsch 2003). Therefore, they need navigable corridors between the two habitats. Although some amphibians exhibit fidelity to their natal wetland, many species also need corridors between wetlands to maintain viable metapopulations (Marsh and Trenham 2001). Altered habi 

210 Inland Wetlands

tats such as agricultural fields, roads, or other forms of human development can act as barriers to amphibian movements because of the increased risk of desiccation or predation (Rittenhouse 2002; Rothermel 2004). Many amphibians are particularly susceptible to mortality on roadways because they migrate en masse to breeding wetlands on rainy nights (Smith, Smith, et al. 2005).

Other Herpetofauna Depressional wetlands are also important habitat for semiaquatic turtles and American alligators (Alligator mississippiensis). Chicken turtles (Deirochelys reticularia) and eastern mud turtles (Kinosternon subrubrum) inhabit wetlands in spring, summer, and early fall and overwinter belowground in adjacent uplands (Buhlmann and Gibbons 2001; Steen et al. 2007). Therefore, the ecotone surrounding wetlands is important to these taxa. Navigable corridors between wetlands are also important to aquatic turtles, which frequently move from depressional wetlands to more permanent water bodies during drought (Buhlmann and Gibbons 2001). Adult male alligators typically inhabit permanent water bodies with large prey, including fish, and utilize open water for courtship. However, smaller adult female alligators may inhabit depressional wetlands or migrate from permanent water bodies to depressional wetlands to nest. Dense emergent vegetation in depressional wetlands offers suitable substrate for constructing nest mounds, and the abundant cover, invertebrate prey base, and low numbers of predators are ideal for juvenile alligator survival (Subalusky, Fitzgerald, et al. 2009). Juvenile alligators leave depressional wetlands as they approach maturity to seek more permanent water bodies that support fish and other large prey (Subalusky, Fitzgerald, et al. 2009). As alligators migrate between habitats, they also use depressional wetlands as tempo-

FIG. 15.5. American alligator (Alligator mississippiensis) in a burrow

within a depressional wetland. The photo was taken during a regional drought, and at least six yellow-bellied sliders (Trachemys scripta) are visible within the burrow. (Photo by Beth and Dirk Stevenson.)

rary stopping points; thus a complex of wetlands within a larger landscape appears to be important to alligator populations. The American alligator also plays an important functional role in southeastern wetlands by creating wallows and dens that offer refugia for aquatic fauna such as amphiuma, siren, and turtles to persist in depressional wetlands during droughts (Kushlan 1974; Mazzotti and Brandt 1994) (Fig. 15.5). Alligators further modify the wetland environment when females construct nest mounds from dead and decaying vegetation. In Florida, alligator nest mounds in marshes are used as nest sites by red-bellied (Pseudemys nelsoni), Florida softshell (Apalone ferox), and eastern mud turtles (Deitz and Jackson 1979). Nest mounds also are colonized by plants intolerant of inundation.

Key Ecosystem Services One of the most important ecosystem services ascribed to southeastern depressional wetlands is providing unique habitats for aquatic and semiaquatic wetland flora and fauna, including numerous endemic, threatened, or endangered species (Sutter and Kral 1994; Kirkman, Drew, et al. 1998; Edwards and Weakly 2001; Dodd and Smith 2003; Sharitz 2003; Smith, Steen, et al. 2006). The relative contribution of these wetlands to regional diversity is disproportionate to their total area (Semlitsch and Bodie 1998). High floristic richness is attributable in part to the variable hydrologic conditions both within and among wetlands, but also to variable fire regimes and the fire-maintained upland-wetland ecotonal habitats (e.g., longleaf pine ecosystem) (Kirkman and Mitchell 2006; Kaeser and Kirkman 2009). Furthermore, at a landscape level, the collective habitats of large and small depressions, wet and dry years, and long and short hydroperiods provide an exceptionally diverse and dynamic assemblage of environmental conditions (Sharitz 2003; Whigham and Jordan 2003). In a survey of rare plants in six states, Edwards and Weakly (2001) found nearly 200 species of concern associated with depressional wetlands, 69 of which were in a threatened status. Most of these were perennial species and occurred in grasssedge marsh and cypress savanna habitats. At least two federally endangered species are associated primarily with ecotonal zones between depressional wetlands and pine-dominated

uplands, including American chaffseed (Schwalbea americana) (Norden and Kirkman 2004) and pondberry (Lindera melissifolia) (Aleric and Kirkman 2005). Amphibians dependent on depressional wetlands as their primary breeding habitat are among the most threatened vertebrates in the Southeast. The Mississippi gopher frog (Rana sevosa), for example, is federally listed as endangered (USFWS 2008), with only one known breeding site for the species remaining. The federally listed (as threatened) frosted flatwoods salamander (Ambystoma bishopi), which was recently recognized as unique from the reticulated flatwoods salamander (A. cingulatum), is known from only a handful of sites in southwestern Georgia and the Florida panhandle. No successful breeding has been reported for this species since the late 1990s. The reticulated salamander also is listed as a threatened species, with fewer than 20 known populations in Florida and southeastern Georgia. Other ecosystem services of depressional wetlands include water storage, nutrient assimilation, and carbon sequestration; however, quantification of most of these in economic terms or environmental consequences is generally lacking, particularly in the southeastern U.S. Depressional wetlands function similarly to other wetlands by providing water storage during periods of heavy rainfall. These wetlands have the potential to improve water quality by assimilating nonpoint-source nutrients released within their basins, particularly in agricultural or urban settings (Dierberg and Brezonik 1984b; Leibowitz 2003; Whigham and Jordan 2003). Although not well documented, depressional wetlands may have important roles in carbon sequestration (Craft and Casey 2000) and the maintenance of human health through food webs that limit zoonotic disease vectors (Kirkman, Whitehead, et al. 2010). Furthermore, nutrient transfer associated with large migrations of amphibians likely represents an important linkage between depressional wetlands and adjacent uplands (Gibbons, Winne, et al. 2006). For example, Smith LL (unpublished data) recorded more than 276,000 juvenile spadefoot toads emigrating from a 0.25-ha wetland in a single breeding event. Although individual toads weighed only about 0.5 g, the total biomass of juveniles exiting the wetland was 560 kg/ha. Although the role of depressional wetlands in the maintenance of regional biodiversity is recognized as an ecosystem service, it is the collective landscape in which these wetlands occur that offers the greatest manifestation of this ecosystem service.

Conservation Concerns Our current national wetland policy fails to recognize and adequately protect depressional wetlands. Without greater recognition of the values and services they provide, the future of these unique wetlands is precarious. Throughout much of the Southeast, depressional wetlands have been altered or destroyed by a wide variety of human activities, including ditching, drainage, agriculture, silviculture, and urbanization (Brinson and Malvárez 2002). Fire suppression within and around many depressional wetlands has promoted successional trajectories toward hardwood domination. In many wetlands, hydrologic modifications have either extended wetland hydroperiods, allowing establishment of fish populations, or have drained the wetlands entirely. Urban development, agriculture, and silvicultural activities have fragmented the larger landscape, presenting challenges to the restoration of these wetlands and regional biodiversity. In addition, potential cliSoutheaster n Depressional Wetlands 211

mate change in the southeastern U.S. may affect the amount and timing of annual precipitation, which, in turn, will modify depressional wetland hydroperiods (Brinson and Malvárez 2002; Stroh, De Steven, et al. 2008). Ultimately, these changes affect wetland vegetation structure, functional processes, and habitat suitability for fauna. Owing to a 2001 Supreme Court decision (Solid Waste Agency of Northern Cook County [SWANCC] v. U.S. Army Corps of Engineers), current U.S. policy fails to adequately recognize and protect the values and services of depressional wetlands because they are deemed “isolated” from other surface waters. Knowledge of the condition of depressional wetlands relative to their support of biotic communities across the Southeast is generally lacking, and methods to assess ecological functions of degraded wetlands are currently limited. Refinement of techniques for creating wetlands or enhancing wetlands through best management practices is necessary to promote policies for delivery of some of the biophysical processes and ecological functions of depressional wetlands. Indices of biotic integrity have been developed for depressional wetlands in Florida (Florida Wetland Condition Indices [FWCIs]; Lane, Brown, et al. 2002; Cohen, Carstenn, et al. 2004; Reiss 2006), which could serve as models for other southeastern states. The FWCIs include both abiotic and biotic metrics, such as water quality parameters, soils, diatoms, macroinvertebrates, and macrophytes. Initial evaluation of the FWCIs revealed that agricultural and urban wetlands had lower biotic integrity relative to reference wetlands (Reiss 2006); however, even in highly urbanized environments, wetlands offered water storage and potential ecological services related to nutrient assimilation. An amphibian index of biotic integrity has been developed for wetlands in Ohio (Micacchion 2002); given the high numbers of amphibian specialists in southeastern depressional wetlands, use of amphibians as indicator taxa warrants investigation. Practical approaches to process-based restoration of depressional wetlands, as well as financial incentives for restoration, are needed. Monitoring and adaptive management can help guide restoration of depressional wetlands. In some cases, wetland seed banks may provide a passive means of revegetation (De Steven, Sharitz, et al. 2006). However, despite the apparent resilience of wetland seed banks, more drastic intervention may be necessary to achieve restoration goals if a threshold to change is surpassed (Martin and Kirkman 2009). Stochastic events such as drought can alter the trajectory of wetland restoration and result in the need for adjustments to management and restoration strategies to achieve desired future conditions (De Steven and Sharitz 2007; Martin and Kirkman 2009). At the national level, the USDA Farm Bill, Wetland Reserve Program (WRP), offers financial incentives to private landowners to restore previously altered wetlands on agricultural lands; however, in parts of the Southeast, the program may be under utilized (De Steven and Lowrance in press). Furthermore, detailed guidelines for restoration in this program are lacking. The current challenge in the Southeast is the development of best management practices for depressional wetlands in agricultural and urban landscapes that can promote components of ecosystem functions. Revegetation of these wetlands with native wetland plants can increase nutrient assimilation and their use by native wildlife. Establishing buffers around the wetlands would further enhance their suitability as wildlife habitat. In many cases, restoration of natural drainage patterns through removal of ditches can provide water quality

212 Inland Wetlands

benefits (Bruland, Hanchey, et al. 2003). Understanding fundamental biogeochemical and hydrologic processes is essential for successfully enhancing and restoring ecological integrity. Correspondingly, a process-based perspective that can promote recognition of the value of inherent services provided by depressional wetlands in the Southeast is central to implementation of policies and management practices that will protect them.

References Aleric KM, Kirkman LK. 2005. Growth and photosynthetic responses of the federally endangered shrub, Lindera melissifolia (Lauraceae), to varied light environments. Am. J. Bot. 92:682–89. Battle JM, Golladay SW. 1999. Water quality and aquatic macro invertebrates in 3 types of reference limesink wetlands in southwest Georgia. In Proceedings of the 1999 GA Water Resour. Conf., Hatcher KJ, editor. Athens, GA: Institute of Ecology, pp. 439–42. Battle JM, Golladay SW. 2001. Water quality and macroinvertebrate assemblages in three types of seasonally inundated limesink wetlands in southwest Georgia. J. Freshw. Ecol. 16:189–207. Battle JM, Golladay SW. 2002. Aquatic invertebrates in hardwood depressions of southwest Georgia. Southeast. Nat. 1:149–58. Battle JM, Golladay SW. 2003. Prescribed fire’s impact on water quality of depressional wetlands in southwestern Georgia. Am. Midl. Natur. 150:15–25. Battle JM, Golladay SW. 2007. How hydrology, habitat type, and litter quality affect leaf breakdown in wetlands on the Gulf Coastal Plain of Georgia. Wetlands 27:251–60. Bliley DJ, Pettry DE. 1979. Carolina bays on the eastern shore of Virginia. Soil Sci. Soc. Am. J. 43:558–64. Boring LR, Hendricks JJ, et al. 2004. Season of burn and nutrient losses in a longleaf pine ecosystem. Int. J. Wildland Fires 13:443–53. Brinson MM, Malvárez AI. 2002. Temperate freshwater wetlands: types, status, and threats. Env. Conserv. 29:115–33. Brinson MM, Lugo AE, Brown S. 1981. Primary productivity, decomposition and consumer activity in freshwater wetlands. Ann. Rev. Ecol. Sys. 12:123–61. Brown SL. 1981. Comparison of the structure, primary production and transpiration of cypress ecosystems in Florida. Ecol. Monogr. 51:403–27. Bruland GL, Hanchey MF, Richardson CJ. 2003. Effects of agriculture and wetland restoration on hydrology, soils and water quality of a Carolina bay complex. Wetl. Ecol. Manag. 11:141–56. Buhlman KA, Gibbons, JW. 2001. Terrestrial habitat use by aquatic turtles from a seasonally fluctuating wetland: implications for wetland conservation boundaries. Chelonian Conserv. and Biol 41:115–127. Busbee WS, Conner WH, et al. 2003. Composition and aboveground productivity of three seasonally flooded depressional forested wetlands in coastal South Carolina. Southeast. Nat. 2:335–46. Casey WP, Ewel KC. 2006. Patterns of succession in forested depressional wetlands in North Florida, USA. Wetlands 26:147–60. Christensen NL. 1988. Vegetation of the southeastern Coastal Plain. In North American terrestrial vegetation, Barbour MG, Billings WD, editors. New York: Cambridge Univ. Press, pp. 317–64. Christensen NL, Burchell RB, et al. 1981. The structure and development of pocosin vegetation. In Pocosin wetlands, Richardson CJ, editor. Stroudsburg, PA: Hutchinson Ross, pp. 43–61. Clymo RS, Turnen J, Tolonen K. 1998. Carbon accumulation in peatland. Oikos 81:368–88. Cohen MJ, Carstenn S, Lane CR. 2004. Floristic quality indices for biotic assessment of depressional marsh condition in Florida. Ecol. Appl. 14:784–94.  

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Collins BS, Battaglia LL. 2001. Hydrology effects on propagule bank expression and vegetation in six Carolina bays. Commun. Ecol. 2:23–33. Craft CB, Casey CP. 2000. Sediment and nutrient accumulation in floodplain and depressional freshwater wetlands of Georgia, USA. Wetlands 20:323–32. Craft CB, Chiang C. 2002. Forms and amounts of soil nitrogen and phosphorus across a longleaf pine–depressional wetland landscape. Soil Sci. Soc. Am. J. 66:1713–21. Deghi GS, Ewel KC, Mitsch WJ. 1980. Effects of sewage effluent application on litter fall and decomposition in cypress swamps. Jour. Appl. Ecol. 17:397–408. Deitz DC, Jackson DR. 1979. Use of American alligator nests by nesting turtles. J. Herpetol. 13:510–12. Demaree D. 1932. Submerging experiments with Taxodium. Ecology 13:258–62. De Steven D, Lowrance R. 2011. Agricultural conservation practices and wetland ecosystem services in a wetland-rich landscape: the Piedmont–Coastal Plain region. Ecol. Appl. 21:S3–S17. De Steven D, Sharitz RR. 2007. Transplanting native dominant plants to facilitate community development in restored Coastal Plain wetlands. Wetlands 27:972–78. De Steven D, Toner MM. 2004. Vegetation of upper Coastal Plain depression wetlands: environmental templates and wetland dynamics within a landscape framework. Wetlands 24:34–42. De Steven D, Sharitz RR, Singer JH, Barton CD. 2006. Testing a passive revegetation approach for restoring Coastal Plain depression wetlands. Restor. Ecol. 14:452–60. Dierberg FE, Brezonik PL. 1984a. Water chemistry of a Florida cypress dome. In Cypress swamps, Ewel KC, Odum HT, editors. Gainesville: Univ. Presses of Florida, pp. 34–50. Dierberg FE, Brezonik PL. 1984b. The effect of wastewater on the surface water and groundwater quality of cypress domes. In Cypress swamps, Ewel KC, Odum HT, editors. Gainesville: Univ. Presses of Florida, pp. 164–70. Dietz-Brantley SE, Taylor BE, Batzer DP, DeBiase AE. 2002. Invertebrates that aestivate in dry basins of Carolina bay wetlands. Wetlands 22:767–75. Dodd CK Jr. 1992. Biological diversity of a temporary pond herpetofauna in north Florida sandhills. Biodiv. Conserv. 1:125–42. Dodd CK Jr. 1993. Cost of living in an unpredictable environment: the ecology of striped newts (Notophthalmus perstriatus) during a prolonged drought. Copeia 1993:605–14. Dodd CK Jr, Smith LL. 2003. Habitat destruction and alteration: historical trends and future prospects for amphibians. In Amphibian conservation, Semlitsch RD, editor. Washington, DC: Smithsonian Institution, pp. 94–112. Edwards AE, Weakly AS. 2001. Population biology and management of rare plants in depression wetlands of the southeastern Coastal Plain, USA. Nat. Areas J. 21:12–35. Entrekin SA, Golladay SW, Batzer DP. 2001. The influence of plant community on chironomid secondary production in two wetland types: cypress-gum swamps and grass-sedge marshes. Archiv. Hydrobiol. 152:369–94. Ewel KC. 1990a. Multiple demands on wetlands. BioScience 40:660–66. Ewel KC. 1990b. Swamps. In Ecosystems of Florida, Myers RL, Ewel JJ, editors. Gainesville, FL: Univ. Presses of Florida, pp. 281–323. Ewel KC. 1995. Fire in cypress swamps in the southeastern United States. In Fire in wetlands: a management perspective. Proc. Tall Timbers Fire Ecology Conf. no. 19, Cerulean S, Engstrom TR, editors. Tallahassee, FL: Tall Timbers Res. Sta., pp. 111–16. Ewel KC. 1998. Pondcypress swamps. In Southern forested wetlands: ecology and management, Messina M, Conner W, editors. Boca Raton, FL: Lewis Publ., pp. 405–20. Ewel KC, Mitsch WJ. 1978. The effects of fire on species composition in cypress dome ecosystems. Fl. Sci. 4:25–31. Folkerts GW. 1997. Citronelle ponds: little-known wetlands of the central Gulf Coastal Plain, USA. Nat. Areas J. 17:6–16. Gibbons JW. 2003. Terrestrial habitat: a vital component for herpetofauna of isolated wetlands. Wetlands 23:630–35. Gibbons JW, Winne CT, et al. 2006. Remarkable amphibian biomass and abundance in an isolated wetland: implications for wetland conservation. Conserv. Biol. 20:1457–65.  

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Golladay SW, Taylor BW, Palik BJ. 1997. Invertebrate communities of forested limesink wetlands in southwest Georgia, USA: habitat use and influence of extended inundation. Wetlands 17:383–93. Golladay SW, Entrekin S, Taylor BW. 1999. Forested limesink wetlands of southwest Georgia: invertebrate habitat and hydrological variation. In Invertebrates in freshwater wetlands of North America: ecology and management, Batzer DP, Rader RB, Wissinger SA, editors. New York: John Wiley & Sons, Inc., pp. 197–216. Heimberg K. 1984. Hydrology of north-central Florida cypress domes. In Cypress swamps, Ewel KC, Odum HT, editors. Gainesville: Univ. Presses of Florida, pp. 72–82. Hendricks EL. 1954. Some notes on the relation of ground-water levels to pond levels in limestone sinks of southwestern Georgia. Trans. Amer. Geophys. Union 35:796–804. Hendricks EL, Goodwin MH. 1952. Water-level fluctuations in limestone sinks in southwestern Georgia. U.S. Geol. Survey Water-Supply Paper 1110-E. Hook DD. 1984. Adaptations to flooding with fresh water. In Flooding and plant growth, Kozlowski TT, editor. New York: Academic Press, pp. 265–94. Johnson SA. 2001. Life history, ecology, and conservation genetics of the striped newt (Notophthalmus perstriatus). Unpublished Ph.D. diss., Univ. of Florida, Gainesville. Johnson SA. 2002. Life history of the striped newt at a north-central Florida breeding pond. Southeast. Nat. 1:381–402. Kaeser MJ, Kirkman LK. 2009. Estimating total plant species richness in seasonally inundated depressional wetlands in the longleaf pine ecosystem. Wetlands 20:373–85. Kirkman LK, Mitchell RJ. 2006. Conservation management of Pinus palustris ecosystems from a landscape perspective. Appl. Veg. Sci. 9:67–74. Kirkman LK, Sharitz R. 1993. Growth in controlled water regimes of three grasses common in freshwater wetlands of the southeastern USA. Aquat. Bot. 44:345–59. Kirkman LK, Sharitz RR. 1994. Vegetation disturbance and maintenance of diversity in intermittently flooded Carolina bays in South Carolina. Ecol. Appl. 4:177–88. Kirkman LK, Goebel PC, et al. 2000. Depressional wetland vegetation types: a question of plant community development. Wetlands 20:373–85. Kirkman LK, Lide R, et al. 1996. Vegetation changes and land-use legacies of depression wetlands of the western Coastal Plain of South Carolina: 1951–92. Wetlands 16:564–76. Kirkman LK, Whitehead EA, et al. 2011. A research framework for identifying potential linkages between isolated wetlands and disease ecology. Ecol. Res. 26:875–883. Kirkman LK, Drew MB, West LT, Blood ER. 1998. Ecotone characterization between upland longleaf pine/wiregrass stands and seasonally-ponded isolated wetlands. Wetlands 18:346–64. Koerselman W, Mueleman AFM. 1996. The vegetation N:P ratio: a new tool to detect the nature of nutrient limitation. J. Appl. Ecol. 33:1441–50. Kushlan JA. 1974. Observations on the role of the American alligator (Alligator mississippiensis) in the southern Florida wetlands. Copeia 1974:993–96. Laliberte L, Luken JO, et al. 2007. The ecological boundaries of six Carolina bays: community composition and ecotone distribution. Wetlands 27:873–83. Lane CR, Brown MT, et al. 2002. The development of the wetland condition index for Florida. Final report to the Florida Department of Environmental Protection. Howard T. Odum Center for Wetlands, Univ. of Florida, Gainesville. Leibowitz SG. 2003. Isolated wetlands and their functions: an ecological perspective. Wetlands 23:517–31. Lide R. 1997. When is a depression wetland a Carolina bay? Southeast. Geogr. 38:90–98. Lide RF, Meentemeyer VG, et al. 1995. Hydrology of a Carolina bay located on the upper Coastal Plain of western South Carolina. Wetlands 15:47–57.Liner AE, Smith LL, et al. 2008. Amphibian distributions within three types of isolated wetlands in southwest Georgia. Am. Midl. Natur. 160:69–81. Liner AE, Smith LL, Castleberry SB, Golladay SW, Gibbons JW 2008. Amphibian distributions within three types of isolated wetlands in southwest Georgia. Amer. Midl. Nat. 160:69–81.  

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 


Mahoney DL, Mort MA, Taylor BE. 1990. Species richness of calanoid copepods, cladocerans and other Branchiopods in Carolina Bay temporary ponds. Am. Midl. Natur. 123:244–58. Marsh DM, Trenham PC. 2001. Metapopulation dynamics and amphibian conservation. Conserv. Biol. 15:40–49. Martin KL, Kirkman LK. 2009. Management of ecological thresholds to re-establish disturbance-maintained herbaceous wetlands of the south-eastern USA. Ecol. Appl. 46:906–14. Mazzotti FJ, Brandt LA. 1994. Ecology of the American alligator in a seasonally fluctuating environment. In Everglades: the ecosystem and its restoration, Ogden JC, Davis SM, editors. Delray Beach, FL: St. Lucie Press, pp. 485–505. Micacchion M. 2002. Amphibian index of biotic integrity (AmphIBI) for wetlands. Final report. Columbus, OH: U.S. EPA. Michener WK, Blood ER, et al. 1998. Tropical storm flooding of a Coastal Plain landscape: extensive floodplains ameliorated potential adverse effects on water quality, fishes, and molluskan communities. BioScience 48:696–705. Mitsch WJ. 1984. Seasonal patterns of a cypress dome in Florida. In Cypress swamps, Ewel KC, Odum HT, editors. Gainesville: Univ. Presses of Florida, pp. 25–33. Moler P, Franz R. 1987. Wildlife values of small isolated wetlands in the southeastern Coastal Plain. In Third southeast. nongame endanger. wildl. symp, Odom RR, Riddleberger KA, Ozier JC, editors. Athens: GA Department of Natural Resources, Game and Fish Division, pp. 234–41. Moorhead DL, Hall DL, Willig MR. 1998. Succession of macroinvertebrates in playas of the southern High Plains, USA. J. N. Am. Benthol. Soc. 17:430–43. Mulhouse JM, Burbage LE, Sharitz RR. 2005. Seed bank-vegetation relationships in herbaceous Carolina bays: responses to climatic variability. Wetlands 25:738–47. National Research Council. 1995. Wetlands: characteristics and boundaries. Washington, DC: National Academy Press. Nessel JK, Bayley SE. 1984. Distribution and dynamics of organic matter and phosphorus in a sewage-enriched cypress swamp. In Cypress swamps, Ewel KC, Odum HT, editors. Gainesville: Univ. Presses of Florida, pp. 262–78. Newman MC, Schalles JR. 1990. The water chemistry of Carolina bays: a regional study. Archiv. Hydrobiol. 118:147–68. Newman RM. 1991. Herbivory and detritivory on freshwater macrophytes by invertebrates: a review. J. N. Am. Benthol. Soc. 10:89–114. Norden AH, Kirkman LK. 2004. Persistence and prolonged winter dormancy of Schwalbea americana L. (Scrophulariaceae) following experimental management techniques. Nat. Areas J. 24:129–34. Otte LJ. 1981. Origin, development and maintenance of pocosin wetlands of North Carolina. Unpublished report to the North Carolina Natural Heritage Program. Raleigh, NC: North Carolina Dept. Natural Resour. Commun. Devel. Paris JM. 2005. Southeastern wetland biogeochemical survey: determination and establishment of numeric nutrient criteria. M.S. thesis, Univ. of Florida, Gainesville. Pechmann JHK, Scott DE, Gibbons JW, Semlitsch RD. 1989. Influence of wetland hydroperiod on diversity and abundance of metamorphosing juvenile amphibians. Wetl. Ecol. Manag. 1:3–11. Poiani K, Dixon P. 1995. Seed banks of Carolina bays—potential contributions from surrounding landscape vegetation. Am. Midl. Nat. 134:140–54. Pyzoha JE, Callahan JJ, et al. 2008. A conceptual hydrologic model for a forested Carolina bay depressional wetland on the Coastal Plain of South Carolina, USA. Hydrol. Process. 22:1689–98. Reiss KC. 2006. Florida wetland condition index for depressional forested wetlands. Ecol. Indic. 6:337–52. Richardson CJ. 2003. Pocosins: hydrologically isolated or integrated wetlands on the landscape? Wetlands 23:563–76. Richardson C, Gibbons JW. 1993. Pocosins, Carolina bays and mountain bogs. In Biodiversity of the southeastern United States, Martin WH, Boyce SG, Echternacht AC, editors. New York: Wiley & Sons, pp. 257–310. Rittenhouse TAG. 2002. Spotted salamander migration at a pond located on a forest-grassland edge. M.A. thesis, Univ. Missouri–Columbia.  

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

214 Inland Wetlands

Rothermel BB. 2004. Migratory success of juveniles: a potential constraint on connectivity for pond-breeding amphibians. Ecol. Appl. 14:1535–46. Schalles JR, Shure DJ. 1989. Hydrology, community structure, and productivity patterns of a dystrophic Carolina bay wetland. Ecol. Monogr. 59:65–85. Schneider DW, Frost TM. 1996. Habitat duration and community structure in temporary ponds. J. N. Am. Benthol. Soc. 15:64 – 86. Semlitsch RD. 2000. Principles for management of aquatic-breeding amphibians. J. Wildl. Manag. 64:615–31. Semlitsch RD. 2003. Conservation of pond breeding amphibians. In Amphibian conservation, Semlitsch RD, editor. Washington, DC: Smithsonian Institution, pp. 8–23. Semlitsch RD, Bodie JR. 1998. Are small, isolated wetlands expendable? Conserv. Biol. 12:1129–33. Semlitsch RD, Ryan TJ. 1999. Migration, amphibians. In Encyclopedia of reproduction, Neill EK, Neill JD, editors. San Diego, CA: Academic Press, Inc., pp. 221–27. Sharitz RR. 2003. Carolina bay wetlands: unique habitats of the southeastern U.S. Wetlands 23:550–62. Sharitz RR, Gresham CA. 1998. Pocosins and carolina bays. In Southern forested wetlands ecology and management, Messina MG, Conner WH, editors. Boca Raton, FL: Lewis Publ., pp. 343 – 77. Sklar FH. 1985. Seasonality and community structure of the backswamp invertebrates in a Louisiana cypress-tupelo wetland. Wetlands 5:69–86. Smith LL, Smith KG, et al. 2005. Roads and Florida’s herpetofauna: a review and mitigation case study. In Amphibians and reptiles: status and conservation in Florida, Meshaka WE Jr, Babbitt KJ, editors. Malabar, FL: Krieger Publ Co., pp. 32–40. Smith LL, Steen DA, et al. 2006. The vertebrate fauna of Ichauway, Baker County, GA. Southeast. Nat. 5:599–620. Soller DR, Mills HH. 1991. Surficial geology and geomorphology. In The geology of the Carolinas, Horton W Jr, Zullo VA, editors. Knoxville: Univ. Tennessee Press, pp. 290–308. Solid Waste Agency of Northern Cook County [SWANCC] v. U.S. Army Corps of Engineers, 531 U.S. 159 (2001) Steen DA, Sterrett SC, et al. 2007. Terrestrial movements and microhabitat selection of overwintering subadult eastern mud turtles (Kinosternon subrubrum) in southwest Georgia. J Herpetol. 41:115–127. Stolt MH, Rabenhorst MC. 1987. Carolina bays on the eastern shore of Maryland. II. Distribution and origin. Soil Sci. Soc. Am. J. 51:399–405. Stroh CL, De Steven D, Guntenspergen GR. 2008. Effect of climate fluctuation on long-term vegetation dynamics in Carolina bay wetlands. Wetlands 28:17–27. Subalusky AL, Fitzgerald LA, Smith LL. 2009. Ontogenetic niche shifts in the American alligator establish functional connectivity between aquatic systems. Biol. Conserv. 142:1507–14. Sun G, Callahan TJ, et al. 2006. Modeling the climatic and subsurface stratigraphy controls on the hydrology of a Carolina bay wetland in South Carolina, USA. Wetlands 26:567–80. Sutter RD, Kral R. 1994. The ecology, status, and conservation of two non-alluvial wetland communities in the south Atlantic and eastern Gulf Coastal Plain, USA. Biol. Conserv. 68:235–43. Taylor BE, Leeper DA, et al. 1999. Carolina bays: ecology of aquatic invertebrates and perspectives on conservation. In Invertebrates in freshwater wetlands of North America: ecology and management, Batzer DP, Rader RB, Wissinger SA, editors. New York: John Wiley & Sons, Inc., pp. 167–96. Tiner RW. 2003. Geographically isolated wetlands of the United States. Wetlands 23:494–516. Tyndall RW, McCarthy KA, et al. 1990. Vegetation of six Carolina bays in Maryland. Castanea 55:1–21. USFWS (US Fish and Wildlife Service). 2008. Endangered and threatened wildlife and plants. 50 CFR 17.11 and 17.12. Washington DC. Watt KM, Golladay SW. 1999. Organic matter dynamics in seasonally inundated, forested wetlands of the Gulf Coastal Plain. Wetlands 19:139–48. West LT, Shaw JN, Blood ER, Kirkman LK. 1998. Correlation of water tables to redoximorphic features in the Dougherty Plain, southwest Georgia. In Quantifying soil hydromorphology, SSSA  

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

 

Special Publication no. 54, Rabenhorst MC, Bell JC, McDaniel PA, editors. Madison, WI: Soil Sci. Soc. Am., pp. 247–58. Whigham DF. 1999. Ecological issues related to wetland preservation, restoration, creation and assessment. Sci. Total Env. 240:31–40. Whigham DF, Jordan TE. 2003. Isolated wetlands and water quality. Wetlands 23:541–49. Wiggins GB, Mackay RJ, Smith IM. 1980. Evolutionary and eco 

 

 

logical strategies of animals in annual temporary pools. Archiv. Hydrobiol. Suppl. 58:97–206. Wilbur RB, Christensen NL. 1983. Effects of fire on nutrient availability in a North Carolina Coastal Plain pocosin. Am. Midl. Natur. 110:54–61. Wright AH. 1932. Life-histories of the frogs of the Okefinokee Swamp, Georgia. New York: Macmillan Co.  

 


This page intentionally left blank