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Chapter 4 – The Ecological Function of South Carolina Stormwater Ponds within the Coastal Landscape

Authors

Dianne I. Greenfield, Ph.D., Belle W. Baruch Institute for Marine and Coastal Sciences, University of South Carolina, Columbia, S.C. Marine Resources Research Institute, S.C. Department of Natural Resources, Charleston, S.C. Currently at the Advanced Science Research Center, City University of New York, New York, NY and the School of Earth and Environmental Sciences, Queens College, CUNY, Queens, NY

Erik M. Smith, Ph.D., Belle W. Baruch Institute for Marine and Coastal Sciences, University of South Carolina, Columbia, S.C. North Inlet-Winyah Bay National Estuarine Research Reserve, Georgetown, S.C.

Andrew W. Tweel, Ph.D. and Kimberly Sitta, Marine Resources Research Institute, S.C. Department of Natural Resources, Charleston, S.C.

Denise M. Sanger, Ph.D., Marine Resources Research Institute, S.C. Department of Natural Resources, Charleston, S.C. ACE Basin National Estuarine Research Reserve

Corresponding Author: Dianne I. Greenfield, Ph.D. (DGreenfield@gc.cuny.edu)

4.1 Background

Development and increased impervious surface cover have numerous consequences for coastal ecosystems. One example, elevated runoff, increases the transport and volume of nonpoint source nutrient inputs, especially nitrogen (N) and phosphorus (P), into receiving watersheds and surface waters (Lewitus et al. 2003; Tufford et al. 2003; Drescher et al. 2007a; Sanger et al. 2015). Ponds control water quality impacts to adjacent receiving waters through the removal/retention of pollutants (nutrients, organic matter, sediments, bacterial pathogens, etc.), as they act as buffers between urban areas, residential and golf course communities, and surrounding water bodies (Lewitus & Holland 2003; Lewitus et al. 2003; Brock 2006; Guinn et al. 2014). In coastal S.C., ponds are typically small (< 1 – 10 acres; see Chapter 1), shallow (< 3 m maximum depth) (Drescher et al. 2007a, b; Smith 2012), and the waters span a wide range of salinities, fresh to estuarine, though freshwater ponds outnumber saline ponds (Lewitus et al. 2003, 2008; Drescher et al. 2007a, b), with salinity regime often related to the degree of direct (pipes) or subsurface (groundwater) connectivity with other ponds and/or nearby tidal creek estuaries (Bunker 2004; Brock 2006; Wisniewski 2014). The salinity and connectivities of these ponds directly impact the ecology of associated biota.

Ponds create unique ecosystems because they generally have reduced flushing capacity associated with high volume residence times, making them susceptible to stagnation, particularly during warmer months, late summer through early fall (Lewitus et al. 2003, 2008; Bunker 2004; Vandiver & Hernandez 2009). Furthermore, they accumulate N and P from fertilizer runoff (Lewitus et al. 2003, 2008; DeLorenzo & Fulton 2009; Drescher et al. 2011b; DeLorenzo et al. 2012). As a result, they are natural “incubators” for the proliferation of algal blooms, many of which produce toxins or cause other adverse consequences, thus termed harmful algal blooms (HABs) (Lewitus et al. 2003, 2008; Siegel et al. 2011; Greenfield et al. 2013, 2014a). Ponds may also serve as valuable permanent and/or transient habitats for a wide range of species spanning multiple trophic tiers (microplankton to vertebrate predators). Given rapid and continued development, ponds have become integral features throughout the S.C. coastal landscape. Consequently, it is both timely and essential to evaluate ponds as functioning ecosystems (Sassard et al. 2014).

4.2 Framework

In the “State of the Knowledge Report: Stormwater Ponds in the Coastal Zone” (Drescher et al. 2007a), information is provided regarding pond ecology, largely relating to nutrient loading and resultant eutrophication responses. Since that contribution, several advances have been made in our understanding of pond ecology that are directly relevant to ecosystem function within the coastal landscape and associated management strategies.

This chapter evaluates the current state of knowledge of ponds and their ecological function within the coastal landscape. To generate this updated review, a wide range of resource formats were evaluated, including primary literature, books, conference presentations (talk and/or poster format), technical reports, student theses, and websites. Citations were deemed appropriate for inclusion within this chapter if they met one or more of the following criteria pertaining to S.C. coastal ponds:

  1. Described nutrient sources, inputs, or utilization,
  2. Considered how abiotic factors (e.g., temperature, salinity) affect nutrient chemistry and/or biota,
  3. Evaluated microorganism populations, dynamics, and/or food-web studies,
  4. Assessed pond vegetation,
  5. Evaluated invertebrate or vertebrate species and/or ponds as their habitats,
  6. Considered how landscape characteristics drive biogeochemical cycling or biology within or among ponds, or
  7. In cases where literature for Sotuh Carolina was limited, we addressed any item (1-6) using comparable regional pond systems (e.g., coastal Georgia, North Carolina, Virginia, and/or Maryland).

We realize that ponds are used as a Best Management Practice (BMP) for inland waters, other coastal states, the Midwest, and internationally. Thus, literature covered here is neither exhaustive nor may certain statements and conclusions herein apply to ponds in those systems. Accordingly, since the hydrology, landscape, and ecology of those systems differ markedly, only relevant studies from S.C. and regional states were included here. Our citations include 105 references as 67 primary literature sources, four books, eight conference presentations, 13 technical reports, three websites, and nine student theses. These are supported by an annotated bibliography (Appendix A4). The following sections discuss how ponds act as ecosystems and their ecological relationships within the surrounding landscape. Highlights of major information gaps and associated recommendations for future directions are provided at the end of this chapter as well as at the end of individual sections.

4.3 Ecological Function within the Coastal Landscape

4.3.1 Physical and Biogeochemical Processes

Ponds increase surface water volume and thus residence time, which has several consequences. For example, warm summer temperatures, shallow depths, and high residence times induce stagnation; this may result in depleted dissolved oxygen levels leading to hypoxia and even anoxia, particularly during the summer and early fall (Lewitus et al. 2008; Serrano & DeLorenzo 2008; Greenfield et al. 2009; Reed et al. 2016). Hypoxia is responsible for the majority (68%) of S.C. fish kills, followed by HABs (27%) then others (Greenfield et al. 2017). HABs and hypoxia may occur together, but blooms are associated with oxygen concentrations ranging from anoxic to hyperoxic (Lewitus et al. 2008; Greenfield et al. 2009, 2014b, 2015), depending on whether the bloom is developing, ongoing, or senescing. When phytoplankton cells actively divide during bloom development, they generate oxygen from photosynthesis (net autotrophy), but during senescence, bacterial remineralization increases and oxygen is consumed (net heterotrophy).

Pond biogeochemistry is highly complex, and recent work has helped transform our understanding of nutrient (notably N and P) cycling and associated microbial processes. For example, in urban ponds, particulate N and P contribute more to the total N (TN) and total P (TP) pools than forested creeks, which have relatively greater contributions of dissolved N and P to the total pools (Tufford et al. 2003). However, the availability of N and P to phytoplankton ultimately affects productivity because algae primarily utilize dissolved nutrient forms and ponds encompass a wide range of nutrient conditions (Drescher et al. 2011b; Smith 2012). For example, a study of 26 S.C. residential freshwater ponds spanning low- to high-density residential development, according to NOAA Coastal Change Analysis Program (C-CAP) land use classifications (Smith 2012, Smith et al. 2015), showed that average TN ranged from 287.9 to 3758.5 μg N L-1 while TP ranged from 3.7 to 394.0 μg P L-1. Though TN and TP concentrations were significantly correlated across these 26 ponds, TP was relatively more variable both across and within ponds. The relative contributions of particulate N and P to the total nutrient pools tended to increase with overall nutrient concentrations, likely due to a greater contribution of nutrients bound in phytoplankton biomass from the particulate pool.

While the relative importance of N and P to pond functioning was originally thought to be primarily driven by salinity, emerging evidence suggests that nutrient cycling processes are more similar in fresh and saline pond systems than previously believed. Specifically, in freshwater ecosystems, the classic paradigm is that lake productivity is usually limited by P availability (e.g., Dillion & Rigler 1974; Prarie et al. 1989), but more recent work suggests this is not always true (e.g., Abell et al. 2010). Consistent with the idea that freshwater productivity is often limited by P availability, phytoplankton biomass (chlorophyll a) in the above-mentioned 26 ponds was strongly correlated to TP concentrations, though experimental evidence suggests that the dependence of freshwater phytoplankton biomass on nutrient availability is more complex than a simple cause and effect relationship. As examples, during nutrient enrichment studies in several of the ponds described above, phytoplankton growth was stimulated by either N alone or, more commonly, by both N and P, rather than just P (Smith, unpubl. data). Similarly, Wisniewski (2014) found significant increases in phytoplankton numbers and biomass in response to N, but not necessarily P, additions using both inorganic (ammonium) and organic (urea) N-sources in fresh to brackish (0 – 5 psu) ponds, and responses were particularly augmented in urea-containing additions. These studies combined suggest that N can act as a limiting or co-limiting nutrient in fresh or nearly-freshwater pond systems. While bottle incubations, like any controlled experimental design, have numerous well-known caveats (e.g., closed systems, limited duration, grazing), results are highly informative for assessing limiting nutrients and other trophic pathways in aquatic ecosystems.

Compared to freshwater ponds, brackish and saltwater ponds are primarily N-limited (e.g., Lewitus et al. 2008; Siegel et al. 2011; Greenfield et al. 2014a; Reed 2014; Reed et al. 2016), but growth and biomass of certain HAB species are enhanced by P (Greenfield et al. 2014a, 2015). Experimental evidence suggests that elevated N, especially urea, an organic N source, stimulates increases in phytoplankton biomass, particularly those of cyanobacteria, dinoflagellates, and other HAB species (Siegel et al. 2011; Reed et al. 2016). For example, Reed et al. (2016) conducted a two-year study of seasonal phytoplankton assemblage responses to three N-forms (ammonium, nitrate, and urea) with and without orthophosphate in a saline pond (mean 19.9 psu, range 11.3 to 32.7 psu). Results showed that N-additions, particularly urea, enhanced phytoplankton growth and HAB species abundances in ponds more than in less-developed landscapes (low to moderately developed tidal creeks), suggesting that ponds may be less resilient to nutrient loadings than comparatively less-developed systems. Furthermore, urea-fueled phytoplankton biomass resulted in greater contribution of phytoplankton-derived dissolved organic carbon (DOC) to the overall DOC pool in ponds and other developed systems than in less-developed tidal creeks (Reed et al. 2015). These findings have direct management relevance because greater than 50% of commercial fertilizers worldwide use urea (Glibert et al. 2006), so fertilization may alter the biogeochemical cycling of carbon in ponds. In sediments, total organic carbon (TOC) levels have been shown to vary widely among ponds (Weinstein et al. 2008), though this variability was not significantly related to land use or sampling location. Furthermore, the differences between fresh and salt water systems need exploration, particularly regarding N-cycling. For example, denitrification rates (reduction of nitrate to other N-forms, ultimately producing dinitrogen (N2) are more rapid in freshwater pond sediments than saline wetlands (Aelion & Warttinger 2010). However, denitrification may be enhanced by pond sediments in both fresh and brackish systems (Drescher 2005) because reaction rates are influenced by hydrology and N-loading (Strosnider et al. 2007), ultimately influencing the N-pool.

Research Gaps

  • Elucidate the parameters (e.g., extent, duration) of hypoxic and anoxic events and their physical, biogeochemical, and climatic regulation.
  • Quantify heterotrophic remineralization and respiration rates.
  • Evaluate nutrient sources and inputs and nutrient biogeochemistry processing within and among ponds.
  • Understand the mechanisms by which management practices (including but not limited to fertilization) affect biogeochemical cycling (particularly N, P, and C) in ponds.
  • Integrate nutrient dynamics with hydrological factors (surface and groundwater inputs, flow rates, and retention times, etc.).
  • Nutrient (both form and concentration) utilization by algae and subsequent bloom dynamics and HAB/toxin proliferation.

4.3.2 Vegetation

Managers commonly use vegetation around ponds to establish buffer zones, provide habitat for fauna, and enhance aesthetic value (e.g. Drescher et al. 2007, 2011a; Strosnider et al. 2007). Pond vegetation is broadly characterized as submerged, floating, or emergent (www.clemson.edu/extension/hgic/plants/other/landscaping/hgic1714.html; Rollins 2012; Rollins et al. 2012). Submerged plants (e.g., invasive Hydrilla, Eurasian watermilfoil) are rooted to the bottom of ponds and serve both management and ecological roles, such as stabilizing soils to prevent erosion and providing habitat for fish, turtles, and other macrofauna. Floating plants (e.g., alligatorweed, duckweed, water hyacinth) are common in S.C. and can rapidly cover a pond, thereby blocking sunlight to the benthos, potentially causing oxygen depletion within ponds. Monitoring and fish kill responses have shown positive associations between floating plants, such as duckweed, and hypoxia (Greenfield et al. unpubl. data). To control certain invasive vegetation, such as Hydrilla, non-native tilapia and/or triploid grass carp are commonly introduced to ponds (Rollins 2012). These fish species effectively graze several submerged and floating vegetation species, but they are comparatively less effective at controlling duckweed, except at salinities that are sub-optimal for rapid duckweed growth (N. Shea, pers. comm.). Consequently, herbicide application is the most commonly used method for duckweed management. Emergent vegetation, such as cattails and Phragmites, is the most commonly found pond vegetation category, and these species can sometimes cover expansive areas and mitigate shoreline erosion (N. Shea, pers. comm.). Beyond management, there is limited information on the ecological function(s) of vegetation in S.C. ponds, particularly how species (native and non-native plants) affect N and P cycling and faunal habitat. Studies in North Carolina have shown that pond and streamside vegetation can take up nutrients, thereby reducing levels in the water (Borden et al. 1997; Mallin & Wheeler 2000; Mallin et al. 2002), though removal efficiency is mediated by pond geometry (e.g., size, shape, depth) (Borden et al. 1997). Perimeter vegetation has also been shown to play an important role in carbon sequestration and increases macrofaunal diversity by creating vital habitat (Moore & Hunt 2012). Furthermore, the rhizosphere has been shown to be particularly important for N-cycling and removal through denitrification and ANAMMOX (ANaerobic AMMonium OXidation), particularly during summer (Song et al. 2014). However, the biogeochemical roles of plants in S.C. ponds are substantially understudied but likely highly important for nutrient cycling. Understanding the ecological role(s) of pond vegetation remains a notable gap in our knowledge of pond ecological function.

Research Gaps

  • Thorough assessments of native vs. non-native plant species in coastal S.C. ponds.
  • Assessments of nutrient uptake by plants (forms and rates).
  • Influence of seasonality, drought, and/or flooding on vegetation.
  • Evaluations of how soil type and maintenance (dredging, erosion control, construction, etc.) affect the growth and productivity of surrounding vegetation.
  • Importance of vegetation as critical macrofaunal habitat.
  • Relative importance of herbaceous vs. woody vegetation as perimeter species for providing habitat and/or erosion control.

4.3.3 Microflora and Fauna

Bacteria

Ponds commonly act as reservoirs for microscopic plankton (viruses, bacteria, algae, and zooplankton) because microorganisms often thrive in the water column and/or sediment environments that ponds provide (Lewitus et al. 2008; DeLorenzo & Fulton 2009; Greenfield et al. 2014b). No information was found on viruses in S.C. ponds during this review, but a Virginia study showed virus loads positively correlated with runoff (Williamson et al. 2014). Bacteria that pose concerns for public health, such as fecal coliform and pathogenic Vibrio spp., can reach high concentrations in S.C. ponds (Hathaway et al. 2009; DeLorenzo et al. 2012; Greenfield et al. 2014b, 2017b). In fact, Vibrio spp. levels in saline ponds can be enhanced by algal blooms and organic matter loading (Greenfield et al. 2014b, 2017b), suggesting that HABs at times act as a vector for Vibrio spp. proliferation. Bacteria, particularly fecal coliform, are also introduced to ponds by wildlife, such as the American alligator (Johnston et al. 2010), though other taxa, such as birds, are also likely contributors. Finally, some bacteria may be resistant to antimicrobial management, as DeLorenzo et al. (2012) found that 90% of 540 tested E. coli isolates collected from S.C. coastal ponds were attributed to pet waste and resistant to one or more antibiotics.

Algae

Pond algae have two major forms: filamentous and non-filamentous (phytoplankton). Filamentous algae are primarily cyanobacteria, such as fresh and saltwater species of Lyngbya. The frequency and geographic extent of filamentous cyanobacteria blooms is increasing dramatically across the East Coast (O’Neil et al. 2012; Paerl & Paul 2012). In addition, eukaryotic (often multi-nucleate) plant species (Viridiplantae) can produce substantial filamentous growth, and examples in S.C. ponds (usually freshwater) include Pythophora and Hydrodictyon. While well-studied elsewhere, research on filamentous algae is largely missing for S.C. ponds.

Phytoplankton are comparatively better studied because dense, pervasive, and often toxic blooms are a common feature of ponds (Lewitus & Holland 2003), and long-term monitoring has identified ponds as HAB hot spots, often linked with fish kills (Greenfield et al. 2017a). Blooms occur because ponds have poor flushing and accumulate runoff which makes them susceptible to eutrophication (Lewitus et al. 2003, 2008; Drescher et al. 2007b; DeLorenzo & Fulton 2009). HABs are most common from late spring through mid-fall (Hayes & Lewitus 2004; Hayes et al. 2008; Greenfield et al. 2009) and can present serious public health and environmental concerns. Blooms can cause fish kills and produce a range of toxins that induce gastroenteritis, dermal or respiratory ailments, and/or other health effects (reviewed in Chorus & Bartram 1999; Anderson et al. 2008; Heisler et al. 2008; O’Neil et al. 2012). Algal blooms and incidences of HAB species are more frequent in ponds surrounded by developed and/or managed land (e.g., residences, golf courses) than less developed land (Borden et al. 1997; Brock 2006; Reed 2014; Reed et al. 2016), likely due to higher nutrient inputs from fertilizers, pet waste, and wildlife (geese) (Lewitus et al. 2003, 2008; Serrano & DeLorenzo 2008; Siegel et al. 2011; Greenfield et al. 2014a). Although turf management and fertilizer application practices for areas draining into many pond networks are proprietary and consequently unavailable for this review, urea-based fertilizers comprise greater than 50% of nitrogenous fertilizers worldwide, and the use of urea in fertilizers is predicted to increase (Glibert et al. 2006). Thus, it is feasible that urea-based fertilizer application fosters HAB proliferation in S.C. ponds, but this hypothesis requires further testing. Overall, the capacity for microbes to utilize N and P, and their respective metabolic rates, is poorly studied but critical for understanding how nutrients are cycled within and among pond systems. A mechanistic understanding of how nutrients drive blooms is lacking, as are further studies involving bloom management.

HABs

Major HAB taxa in S.C. ponds often include cyanobacteria, raphidophytes, dinoflagellates, and diatoms, though prymnesiophytes and euglenophytes are also common. HABs have been associated with 27% of fish kills in S.C. ponds (Greenfield et al. 2017a). Cyanobacteria are responsible for the largest number of toxic HABs worldwide (Chorus & Bartram 1999), and they thrive in warm, nutrient-rich, stagnant conditions, such as ponds (Paerl et al. 2001; Lewitus et al. 2008; Greenfield et al. 2014a). Cyanobacteria HABs span the entire range of coastal salinities (fresh to marine) and can produce numerous toxins. Commonly found genera/species include but are not limited to Anabaenopsis, Microcystis aeruginosa, M. flos-aquae, Oscillatoria, Cylindrospermopsis, Aphanizomenon, Anabaena, and Pseudanabaena (Lewitus et al. 2003; Brock 2006; Serrano & DeLorenzo 2008; Siegel et al. 2011; Greenfield et al. 2013, 2014a; Wisniewski 2014; Dearth 2017). Cyanobacteria toxins detected in S.C. lakes and ponds have been associated with human health effects. For example, during the summer of 2014, an adult male suffered a rash and swelling of the arm in response to a bloom of Lyngbya sp., a filamentous cyanobacterium in a freshwater lake that tested positive for lyngbyatoxin (Greenfield et al. unpubl. data). In ponds, the most commonly-detected toxin is microcystin, a hepatotoxin (Sivonen & Jones 1999) that can induce gastroenteritis, liver failure, and/or death (Pouria et al. 1998; Falconer 2005). During June 2000, a 7-year old girl waded in a S.C. coastal pond with a highly toxic (greater than 1,000 μg L-1 microcystin) M. aeruginosa bloom and subsequently developed a severe body rash and respiratory ailment leading to a S.C. Department of Health and Environmental Control (DHEC) water posting (R. Ball, pers. comm.). Microcystin values exceeding 1,000 μg L-1, though rare, have been reported elsewhere in coastal S.C. during dense M. aeruginosa blooms (Brock 2006; Dearth 2017), but more typical values range 1 to approximately 100 μg L-1 (Brock 2006; Greenfield et al. 2013). For reference, the provisional guideline for drinking water issued by
the World Health Organization (WHO) of the most lethal and common toxin, microcystin-LR, is 1 μg L-1, and the
WHO classifies microcystin concentrations ranging from 10 – 20 μg L-1 and 20 – 2,000 μg L-1 as posing moderate and
high risks, respectively, for probable acute health effects in recreational waters (www.epa.gov/cyanohabs/world-healthorganization-
who-1999-guideline-values-cyanobacteria-freshwater).

Raphidophytes are commonly associated with fish kills, particularly the species Chattonella subsalsa, C. verriculosa, Fibrocapsa japonica, Heterosigma akashiwo, and Viridilobus marinus (Lewitus et al. 2003, 2004, 2008; Keppler et al. 2006; Reed 2014; Greenfield et al. 2015; Greenfield et al. 2017a). For example, raphidophytes were associated with 40 blooms and subsequent fish kill events between the months of June 2014 and August 2015 alone (Greenfield et al. 2015). Lysosomal destabilization (a sublethal effect) was observed in eastern oysters following exposure to C. subsalsa and F. japonica bloom water (Keppler et al. 2006), indicating that raphidophytes can exert deleterious effects on several animal taxa. In S.C. ponds, raphidophyte blooms are associated with low N to P ratios but high overall nutrient concentrations (Lewitus et al. 2004, 2008; Greenfield et al. 2015). Populations may be further regulated by microbes, as raphidophyte growth may be enhanced by certain bacterial assemblages (Liu et al. 2008a) or limited by algicidal species (Liu et al. 2008b). Dinoflagellates of the toxic genus Pfiesteria were documented in S.C. ponds and estuaries during the 1990s and 2000s, in some cases associated with fish kills (Lewitus et al. 2002, 2003; DeLorenzo & Fulton 2009). Pfiesteria blooms have not been reported in S.C. ponds since the previous DHEC State of the Knowledge report, but other dinoflagellate species, such as Karlodinium veneficum, Prorocentrum minimum, Protoperidinium quinquecorne, and Kryptoperidinium foliaceum, are common bloom-formers, either alone or as part of a multi-specific assemblage (Kempton et al. 2002a, b; Lewitus et al. 2003, 2008; DeLorenzo & Fulton 2009; Greenfield et al. 2014a). Finally, blooms of the diatom Pseudo-nitzschia have been observed with increasing frequency in ponds and across the S.C. coast (Goodson et al. 2011; Reed 2014; Reed et al. 2016; Sitta et al. 2018). This is cause for substantial concern because several species of this genus produce the neurotoxin domoic acid, which is responsible for amnesic shellfish poisoning as well as the largest harvest closure for razor clams and Dungeness crabs along the Pacific coast in history (summer to fall of 2015).

Despite increasing knowledge of the extent of HABs in ponds, major gaps exist in our ability to accurately detect and predict a bloom’s onset. A more thorough understanding of bloom processes, as well as the development and application of tools and technologies that rapidly characterize HABs, microbes, and other pathogens, is badly needed to safeguard public health. Current algal bloom management practices include but are not limited to aerators, chemical algicides, and vegetation (Halfacre et al. 2007; Greenfield et al. 2014a; Hehman 2014). A before and after study of aerator installments in S.C. ponds showed that aerators improved (increased) water dissolved oxygen levels but did not significantly affect nutrient concentrations or phytoplankton assemblages (biomass, community composition, or primary productivity rates) (Hehman 2014). During a combined field and laboratory study using pond bloom water, Greenfield et al. (2014a) showed that both copper and peroxide-based algicides controlled cyanobacteria numbers and biomass. However, toxins (microcystins) merely transitioned from intracellular to extracellular fractions as cells lysed due to chemical exposure. Moreover, total water toxin levels were not significantly different from controls (no algicide), suggesting that algicides may visibly control blooms, but they do not mitigate toxicity.

Examples of future directions should include the development of technologies enabling early-warnings and characterization, such as by molecular methods that detect and quantify HABs and other microorganisms (Greenfield et al. 2008; Doll et al. 2014; Dearth 2017). An emerging research priority and management need is to understand how molecular and genetic factors drive species assemblages, population structure, and responses to environmental conditions. With the emergence of novel genomic and related “omic” technologies (e.g., metabolomics, proteinomics, transcriptomics), scientists have the capacity to elucidate a detailed, mechanistic understanding of species populations and processes, at population, individual, and sub-cellular levels. Studies using genomics tools to answer questions about S.C. pond ecosystems are currently sparse.

Zooplankton

It is widely known that larger zooplankton (mesozooplankton, > 200 μm in length), such as copepods, play major ecological roles in aquatic ecosystems by exerting “top-down” grazing pressure on phytoplankton and microzooplankton (20 to 200 μm). Surprisingly, during this review we found almost no information on mesozooplankton assemblages and grazing for S.C. ponds. Since grazing may be central to bloom regulation, this is a significant information gap. Hayes et al. (2008) showed that microzooplankton consume phytoplankton during dilution experiments, but otherwise, information on the ecological role(s) of zooplankton in S.C. ponds, or how their distribution(s) vary along environmental gradients (e.g., salinity, connectivity with other ponds and tidal creeks), is lacking. However, studies of Maryland ponds have shown that zooplankton community diversity is affected by salinity and dissolved oxygen concentrations (Sokol et al. 2015), suggesting that ambient environmental conditions are important.

The vast majority of information cited above is derived from monitoring and event-response activities. Despite the critical importance of such programs for evaluating pond water quality and identifying incidences of public and environmental health concerns there remain numerous gaps in our understanding of microorganism processes. Virtually nothing is known about the role(s) of grazers on microbial and phytoplankton pond assemblages. New information shows that zooplankton grazing influences seasonal phytoplankton biomass and community composition in S.C. tidal creeks (Sitta et al. 2018), with microzooplankton grazing exerting a substantially greater influence on phytoplankton assemblages than mesozooplankton, but similar studies are lacking for ponds.

Research Gaps

  • Population dynamics of bacterial pathogens.
  • Understanding of controls on filamentous algal growth.
  • Evaluations of microbial interactions and regulatory pathways, such as viral or bacterial-mediated control of algae or allelopathy.
  • Inter- and intra-specific competitive interactions.
  • Influence of seasonality and climate on algal and microbial processes.
  • Molecular and “omics” tools and other technologies enabling rapid assemblage characterizations, ecophysiology studies, and predictions of HABs and other pathogens.
  • Transcriptomics to elucidate species and genetic diversity within and among systems.
  • Greater understanding of the drivers of environmental compound production by microorganisms and/or other processes.
  • Understanding of the interactions between viruses, bacteria, algae, and potentially other trophic tiers.
  • Assessments of zooplankton communities, both meso- and micro-zooplankton species.

4.3.4 Macrofauna

Invertebrates

Ponds in coastal S.C. are not primarily designed to create habitat; however, they nevertheless support insects, invertebrates, fish, reptiles, and birds through infaunal, demersal, and pelagic food webs. A major finding from this review was that information on invertebrate and vertebrate ecology, including parasites, within S.C. ponds is generally lacking, highlighting this as an area where considerable research is needed. Few studies on insects were noted, though available literature suggests that mosquito larvae thrive in the warm, stagnant conditions offered by ponds (e.g., Hunt et al. 2005; Gingrich et al. 2006; Ellis et al. 2014). However, managers have suggested that mosquito larvae are not a major issue in S.C. ponds because they are often and effectively controlled by fish (e.g., Gambusia) (N. Shea, pers. comm.). Additionally, frogs are well-known consumers of mosquito larvae, but no literature was available describing their role consuming larvae specifically in S.C. ponds. These separate lines of evidence underscore that further investigation into the potential role of ponds as mosquito larval habitat is needed.

Amphibians and Reptiles

Amphibians are well-known to be sensitive to urbanization because they require both terrestrial and aquatic resources, are sensitive to ecosystem disruption, and are often outcompeted by exotic species. Surveys have noted amphibians in S.C. ponds (Holt et al. 2015), but research is limited. However, other studies indicate that amphibians are impacted by pond construction. For example, frogs use ponds as breeding habitat (Brand & Snodgrass 2010), and ponds near Charlotte, N.C. were inhabited by up to 12 anuran species (Brix-Raybuck et al. 2010), including all but one species typically found in the region (Dorcas & Gibbons 2008). Presence of anurans in ponds is negatively associated with distance from the nearest riparian zone (Simon et al. 2009; Brix-Raybuck et al. 2010). Microcosm experiments comparing frog species tolerant of urbanization (American bullfrog, Bufo americanus) to non-tolerant species (wood frog, Rana sylvatica) exposed to pond sediments, containing elevated metals and chlorides, resulted in 100% mortality of R. sylvatica embryos compared to B. americanus, which only experienced sub-lethal effects (size reduction at metamorphosis) (Snodgrass et al. 2008). Although sediment characteristics may pose deleterious consequences, Massal et al. (2007) found that N-levels in Maryland ponds were not high enough to cause frog mortality. These studies underscore the importance of surrounding landscape characteristics for amphibian populations. Surveys of urban, golf course, and farm ponds in N.C. showed these systems host a range of turtle species, with relative abundances positively correlated with pond area, and nesting frequency negatively associated with extent of anthropogenic use (e.g., sidewalks, fairway) (Failey et al. 2007; Foley et al. 2012). Ponds are also commonly used as habitat for the American alligator (Alligator mississippiensis) (Johnston et al. 2010; N. Shea, pers. comm.).

Birds and Fish

The infaunal and demersal food webs are most likely the dominant components; however, ponds are often stocked with fish and attract wading birds and gulls (CIDEEP Report 2009; Rollins 2012; Ellis et al. 2014). In S.C., one unpublished study assessed bird usage of a select number of ponds (Nolan, pers. comm.). Similar research has been conducted in other states including Florida and Delaware (e.g., Wall 2006; Sloan 2008). Virtually nothing is known about how S.C. ponds affect bird habitat quality and use, particularly by migratory birds that may rely on S.C. coastal habitats during seasonal stop-overs. In one Maryland study, redwinged blackbirds (Agelaius phoeniceus) were found to nest in ponds located in commercial, residential, highway, and wetland (reference) areas, though nestling health was negatively impacted by sediment zinc concentrations (Sparling et al. 2004). Finally, it is well-known that ponds are commonly stocked with fish (Drescher et al. 2007a; Halfacre et al. 2007; Rollins 2012), often to mitigate growth of certain vegetation (described earlier), but little is known about how fish use ponds for breeding habitat. Some species, like the economically-important and catadromous American eel (Anguilla rostrata), have been collected from S.C. coastal ponds (S. Arnott, pers. comm.), though the extent to which eels use ponds as habitat is unknown.

Research Gaps

  • The presence and impact of parasites on host populations and physiology.
  • Influence of bivalves and other suspension feeders on pond water quality.
  • The diversity and richness of macro-flora and fauna, including reproductive and migratory patterns.
  • Assessments of benthic faunal communities (annelids, nematodes, and others) and their influences on sedimentary processes.
  • The role of amphibians for regulating insect populations, especially given the global decline of amphibians and potential impact of management practices, such as pesticide use for mosquito control, on these taxa.
  • How habitat impacts reptiles and thus the presence of higher trophic tiers.
  • Assessments of ponds as nursery habitats or refugia for fish.
  • Tidal influences on animal migratory patterns (e.g., fish and invertebrate larvae).
  • Habitat quality for nesting and migratory bird species, including waterfowl.
  • Using biomarkers to identify responses including, but not limited to, stress responses across trophic levels.

4.3.5 Coastal Landscape

The construction of ponds has replumbed the way water moves in coastal S.C., as stormwater that previously flowed across the land surface or infiltrated through the soil is now directed through stormwater infrastructure designed primarily to dampen flow but also to improve water quality by retaining sediments and other pollutants, as discussed in the preceding section. The ecosystems and habitats represented by ponds are a direct result of the development of human systems, as there are essentially no natural, open-canopy ponds or lakes in the southeastern coastal plain. The typical hydrology of this area is one dominated by groundwater flow through a shallow surface aquifer, with upland surface waters largely limited to forested wetlands draining to blackwater creeks. Coastal development generally leads to the loss of pervious surfaces that allow rainwater to infiltrate to groundwater. Construction of ponds increases the volume of surface water and thus increases the overall residence time of surface water on the landscape.

Very little is known about how these ponds fit within the landscape, including their relationships to soil type and drainage or proximity to receiving water bodies and other sensitive coastal habitats. Their connectivity with receiving waterbodies and how their placement within the watershed might impact their ecologic function within the landscape is largely unknown and likely complex. Their role in altering ecological function within the broader coastal landscape may occur either indirectly through changes in hydrology or material transport, or directly through the creation of habitat types that previously did not exist in this region.

Transport of materials and organisms in and out of ponds has a number of consequences for adjacent systems. For example, levels of dissolved nutrients, HABs, and pathogenic Vibrio spp. in ponds can mimic concentration trends recorded in receiving tidal creeks (Greenfield et al. 2009, 2014, 2017), suggesting that ponds can be sources of these and potentially other constituents. Conversely, DeLorenzo et al. (2012) found that fecal coliform levels were higher in receiving tidal creeks than associated ponds, suggesting that ponds may have been removing bacteria. This is partially due to direct deposit of fecal bacteria in creeks by wildlife. Many studies show that ponds can exert significant removal rates for suspended sediments and fecal coliform (Wu et al. 1996; Borden et al. 1997; Comings et al. 2000; Mallin et al. 2002; Hogan and Walbridge 2007; Hathaway et al. 2009; Krometis et al. 2009; Huda & Meadows 2010; Moore 2010; Hathaway & Hunt 2012; Smith & Peterson, unpubl. data). Pollutant removal capabilities are discussed in more detail in Chapters 2 and 3. Ultimately, ponds are constructed aquatic ecosystems designed specifically to meet the stormwater management requirements associated with development or other land disturbing activities.

Detention ponds are, by far, the most frequently used stormwater management practice in the S.C. coastal zone, and they link upland runoff to coastal aquatic ecosystems through distinct outlet structures. In larger development projects, ponds are often created in linear series along natural drainage paths or are interconnected through a network of underground pipes. Recent stormwater runoff modeling work has allowed quantification of the relative effects of soil type, surrounding land use, elevation, and potential changes in climate on runoff volume and timing (Blair et al. 2014). In the context of coastal development and associated ponds, this modeling demonstrates the large effect that position within the landscape (e.g., elevation gradient, soil type, vegetation) can have on the quantity and timing of stormwater inputs to ponds. For instance, the conversion of a forested coastal watershed to an urban watershed can, at a minimum, triple the volume of water flowing across the landscape that would have otherwise infiltrated or evapotranspired.

Of particular concern, from an environmental impacts perspective, are those ponds immediately adjacent to the land margin and having direct tidal exchange with coastal waters. The proportion of total water flow from uplands to the coastal zone that is collectively routed through ponds is currently unknown. Likewise, the extent to which groundwater exchange represents a significant biogeochemical link between ponds and other coastal aquatic ecosystems remains an open question. Coastal pond discharge is typically linked to tidal creek and open water subsystems, although in certain areas, especially the greater Myrtle Beach region, ponds can be connected directly to open beach environments through coastal swashes and discharge pipes. Coastal S.C. water bodies, especially tidal creek headwaters, are sensitive to upstream inputs (Holland et al. 2004; Sanger et al. 2015) and the potential relationship between ponds and the receiving waterbodies has been discussed above.

Furthermore, these relationships between ponds and the surrounding landscape likely vary spatially. As a general trend in S.C. and other flood-prone landscapes, higher areas less vulnerable to flooding are generally developed first, with subsequent development expanding into wetlands (e.g., downtown Charleston, New Orleans). As these areas become developed, newer development often occurs in the previously undesirable lower-lying areas (Vandiver & Hernandez 2009). These low-lying ponds are more likely to be impacted by marine processes including tidal exchange, influx of seawater, and rising sea levels. Influx of seawater to freshwater systems, for instance, has been observed to significantly alter nitrogen cycling (Aelion & Warttinger 2010). This presents a major potential shift in pond function and performance, which may be exacerbated by aging stormwater infrastructure (S.C. Sea Grant Consortium 2009).

The identification of these low-lying ponds, and an improved understanding of how their function may change over time as tidal connectivity increases, is needed to better understand, plan, and manage pond function in the coming decades. Many of these questions could be addressed by assessing present pond function along elevation gradients with various levels of tidal connectivity. Some of these findings could be scaled up by means of a detailed analysis of pond placement and associated soil, land use, and proximity to other habitats within the landscape.

Research Gaps

  • Understand the associations between ponds and surrounding landscape, especially coastal connectivity to ponds.
  • The landscape itself (elevation, flooding), and related consequences for biota.
  • Critical assessment of the impact of scale (pond size and related dimensions) on the ecological function.

4.4 The Future of Pond Ecosystems

Water temperature, salinity, connectivity, and rates of exchange are primary factors regulating ponds and associated organisms, and the ranges of these parameters will likely shift in the future. For much of the Atlantic coast, climate predictions entail longer, warmer summers (e.g., Cronin et al. 2003; Najjar et al. 2010), and precipitation events may be shorter and more intense (Meehl et al. 2007), affecting the rates and volumes of stormwater runoff as well as water column stability through buoyancy-increased stratification. Since ponds are prone to stagnation, these physical and climatic processes may enhance bottom hypoxia, particularly during summer, affect microbial processes rates, enhance HAB development, and increase the numbers of fish kills. In fact, numerous lines of evidence suggest that future climate scenarios favor algal bloom development. For example, the frequency and severity of HABs are increasing in response to warming and eutrophication (Paerl & Huisman 2009; O’Neil et al. 2012). Although nutrients, particularly N, are major drivers of bloom formation (Paerl et al. 2014), remarkably little is known about how climate and global change, alone or combined with nutrient enrichment, will impact HAB toxicity and the tendency for blooms to both form and persist. Moreover, the extent to which a changing climate will impact bacterial respiration rates and other higher trophic level processes, such as the reproductive cycles of macrofauna through earlier and warmer summers, is entirely unknown. These represent critical gaps in our understanding of ponds, and filling these gaps would provide critical information to inform effective coastal management strategies.

Changes in sea level, temperature, and precipitation patterns associated with climate change or other processes would be expected to further alter pond function and their relationships between adjacent ecosystems both upstream and downstream. Future management of these systems will need to address these changes, but in order to do so a number of unanswered questions remain regarding the associations between these pond systems and the surrounding landscape. Arguably, the most critical challenge facing researchers and managers in coastal low-lying regions, such as S.C., is how to adequately prepare for future climate and sea level scenarios. The relative influences of development patterns, landscape use, and climate change on pond processes are virtually unknown. Since sea level is predicted to continue to rise, and many coastal ponds are directly impacted by tides, salt water intrusion, and numerous other physical, chemical, and biological factors, this represents a key gap in our understanding of how managers and residents can prepare for future conditions. Related, a major information gap is understanding how climate change, including potential increases in the frequency and intensity of precipitation events, and sea level rise affect ponds.

Invasive Species

A serious concern is the potential for ponds to host exotic species and parasites, as they may alter aquatic food webs and overall ecosystem health. Invasive applesnails (Ampullairiidae spp.) are increasingly problematic across the Southeast due to high grazing rates on macrophytes (Horgan et al. 2014; www.dnr.sc.gov/invasiveweeds/snail.html). A survey of 200 coastal S.C. ponds found 18% were infested by applesnails, and population numbers are believed to be growing. Although their statewide distribution is not currently known, the extent to which applesnail, and other invasive species, populations may expand is an area of ecological concern.

Research Gaps

  • Knowledge of how climate change will impact biogeochemical cycling and remineralization rates within pond waters and sediments.
  • Inter- and intra-specific population variability and associated responses to environmental factors (e.g., nutrients, temperature, salinity, toxin production).
  • Influence of global change and sea level rise on pond function.

4.5 Summary and Recommendations

Despite advances in our understanding of S.C. pond function since the Drescher et al. (2007a) State of the Knowledge report was issued, numerous and significant gaps remain in our understanding of pond ecology, and how they function within a complex coastal landscape. A surprising outcome from this review was just how little is known about the processes and both biological and physical linkages among and between ponds. For example, do S.C. ponds impact higher trophic levels and, conversely, the role of higher trophic tiers and food web interactions in pond function? Many gaps can be broadly described as needs for integrative studies evaluating the interactions between environmental parameters that drive pond ecosystems, such as research that considers seasonal and/or climate trends. Other notable gaps include studies integrating basic ecological principles. In particular, studies focusing on competitive and/or allelopathic interactions, predator/prey relationships, species’ migration, breeding, and invasions, diversity assessments (both species and genetic/population) are almost entirely absent. Numerous ongoing monitoring efforts collect static and bulk measurements of various water quality parameters. These programs are highly valuable for assessing longterm trends and identifying incidences where water quality becomes problematic for public and/or ecological health. However, monitoring programs are not designed to evaluate processes nor do they adequately achieve that goal. In fact, process studies to date are limited and thus represent a major gap in our understanding of pond ecological function.

A review of S.C. pond hydrology is provided in Chapter 2, but the extent to which surface and subterranean water movement transports nutrients and other materials to/from ponds is of primary concern because pond connectivity influences ecological function. Improved understanding of connectivities between ponds is a significant and open question due to the complex hydrology of ponds from direct (pipe) surface networks, groundwater exchange, and tides. As many ponds are either directly connected to a receiving tidal creek or indirectly connected through daisychain links, elucidating hydrological linkages will inform biological and chemical linkages between and among ponds. Since ponds serve numerous ecological functions (nutrient remineralization, provide habitat for fishes, birds, reptiles, and invertebrates, etc.), it is crucial to not only build upon existing programs but continue to understand and address these gaps. Also key will be communicating these processes and their outcomes to residents, managers, and other stakeholders. As ponds generally increase the property value of adjacent plots, it is essential that residents gain a broader understanding of their ecological and associated economic value and thus enact informed management decisions.

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