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Chapter 2 – Environmental Factors and Design Features that Control Stormwater Transport and Contaminant Fate in Ponds


Vijay M. Vulava, Ph.D., Barbara A. Beckingham, Ph.D., and Timothy J. Callahan, Ph.D., Department of Geology and Environmental Geosciences, College of Charleston, Charleston, S.C.

Corresponding Author: Vijay M. Vulava, Ph.D. (

2.1 Background

Stormwater ponds are recognized as serving both water quantity and water quality functions. Traditionally engineered with the objective to control flooding, they act as reservoirs to collect water and eroding soil from the surrounding landscape and to dampen the storm pulse to receiving water bodies. This function is important for the integrity of downstream structures, and ecosystems therein (Persson & Wittgren 2003). Over time, it has been shown that ponds can also serve other functions; they provide ecological habitat for birds and aquatic life (Hassall & Anderson 2015), either trap or act as a gateway to the transport of various environmental pollutants to receiving bodies (Van Metre et al. 2000; Thapalia et al. 2010), sequester carbon and provide cultural and ecosystem services (Moore & Hunt 2012), and can serve as a valued, aesthetic feature of a community (Ghermandi & Fichtman 2015). By averting and storing runoff on-site, ponds open up the possibility for rainwater use, such as for landscape irrigation, saltwater intrusion barriers, drinking water aquifer recharge, augmentation of potable water reservoirs, non-potable water use in buildings, and baseflow augmentation to improve freshwater habitat and recreational use depending on water quality and need (Grebel et al. 2013). Thus, how ponds perform hydrologically and geochemically influence how they are viewed as either environmental assets or liabilities.

2.2 Framework

The primary focus of this chapter is the fate and transport of environmental contaminants in ponds. Due to the complex interdisciplinary nature of the problem, this chapter includes multiple objectives:

  1. Hydrology: to describe the roles that surface water and groundwater play in transporting water and contaminants to and within ponds in different geographical and geological settings across the Southeastern U.S., with a focus on the conditions found in the eight coastal S.C. counties.
  2. Nature of contaminants: to identify key contaminants of concern associated with runoff and how their properties can influence their fate in pond environments.
  3. Biogeochemical processes: to describe the relevant mechanisms and reactions that influence contaminants and water quality in stormwater and pond systems.

This study includes an extensive review of appropriate studies that outline current state of knowledge from a hydrological and contaminant fate and transport perspective. Our citations include 194 references as 159 primary literature sources, 12 books, four conference presentations, nine technical reports, and six student theses. Gaps in our current knowledge are identified and specific research priorities are offered.

2.3 Physical Hydrology of Stormwater Systems

2.3.1 Climate, Geography, and Geology

The water balance in S.C., derived from the most recently available 30-year climatological average, is dominated by precipitation (average annual precipitation is 48 inches [1,200 mm]) and evaporation and transpiration, often expressed as evapotranspiration, or ET (average annual ET is 34 inches [865 mm]) (Wachob et al. 2009). Only a few site-specific studies address seasonal and longer-term impacts of hydrologic processes (e.g., runoff, sedimentation) on ponds (Drescher et al. 2007). In the lower gradient areas of the Southeastern Coastal Plain (SECP), referred to as the Lowcountry, hydrology is strongly influenced by the flat topography and shallow depths to the water table (Tufford & Marshall 2002; Callahan et al. 2012; Hitchcock et al. 2014). Dry detention basins are recommended for higher permeability sandy soils compared to wet detention basins (ponds described here) in lower-permeability soils and/or shallow water table conditions, yet there appear to be no studies in the lower SECP on how the conveyed stormwater flows into and through different ponds. Identifying the geographic distribution of ponds is a major effort of this state of knowledge report (Chapter 1), but an additional set of indices could be useful for exploring their hydrologic functioning, such as comparing water balance details (duration and fluctuations of pond water level) with underlying soils and geology data for mapped ponds. There may be important links between the substrate and the vulnerability of ponds to flooding and contaminant accumulation (Fletcher et al. 2013).

2.3.2 Water Balance and Flow

Hydrological processes are better understood in the lower SECP as a result of studies conducted over the past few decades (e.g., Sun et al. 2000, 2002, 2006; Messersmith 2007; Tufford & Marshall 2002; La Torre Torres et al. 2011; Callahan et al. 2012). These studies have collectively confirmed the importance to the overall water balance of (a) forest health and productivity through ET as a water demand process and (b) permeable cover in the watershed for groundwater recharge. In S.C. and across the SECP, topographic slope, soils, and land use/land cover are important site factors controlling water flow. Runoff includes overland flow (also called sheetwash) and shallow subsurface flow (soil interflow), which ultimately delivers stormwater to receiving water bodies. If those systems are dynamic – if the overland flow and interflow are focused into a channel – it is referred to as concentrated flow and moves across the landscape as small rills, moderate-sized creeks, and larger streams and rivers.

In the built environment, ditches, culverts, and other engineered systems are designed to move this water away from infrastructure to reduce flooding hazards. Construction of ponds and the associated engineered infrastructure such as ditches and culverts serve to intercept and sequester surface runoff (Fennessey et al. 2001; Fletcher et al. 2013). These developed and developing landscapes in the Southeast U.S. have altered water balance because of reduced ET caused by deforestation, and decreased groundwater infiltration caused by larger coverage of impervious surfaces (hardscapes such as roofs, parking lots, and roads). The net result is increased surface water discharge. This surface water has traditionally been managed through engineered structures such as drainage ditches and storm sewers, and in the past few decades, through routing the stormwater to detention ponds, lagoons, and basins. Urbanized settings have a new water balance that is now dominated by runoff rather than ET or infiltration (Garrett et al. 2012; Fletcher et al. 2013). Ponds and dry basins are structures that are designed to buffer natural receiving waters from runoff (Wachob et al. 2009).

Subsurface-surface connections, as affected by shallow water table conditions, can contribute groundwater inflow into the pond and/or reduce infiltration and recharge due to the small hydraulic head differences between the pond water and the adjacent water table. However, there is a lack of detailed research on the importance of subsurface – surface water exchange in the modern landscape that contains thousands of ponds. Evidence is emerging that groundwater may be a conduit for the movement of constituents in/out of ponds (Bunker 2004; Wisniewski 2014), and potentially from ponds to aquifers (Bunker 2004).

2.3.3 Pond Design and Hydrology

Another classification of stormwater basins is those in which plants play an integral function, such as buffers between the sources and the pond, as well as engineered bioretention cells or constructed wetlands built and permitted to receive runoff. Addition of vegetation within the stormwater pond’s watershed acts to increase demand through plant transpiration and increase surface roughness to reduce runoff velocity (Bettez & Groffman 2012; Hitchcock et al. 2014). Separate from ponds (wet detention basins), bioretention cells are typically very shallow, landscaped, depressional areas and thus do not classify as a pond. While outside the scope of the current review, bioretention cells are a potentially effective low impact development (LID) practice. Constructed wetlands are also shallow vegetated depressions, but with variations in elevation to create environments for slow water flow, a diversity of vegetation, and biogeochemical conditions that can assist in water quality improvements. They are also subject to design and regulation criteria that govern wet detention ponds (Ellis et al., 2014). In a S.C. Department of Health and Environmental Control-Office of Ocean and Coastal Resource Management (DHEC-OCRM) study of 511 ponds, wet detention was the most frequently used type (Drescher et al. 2007), and is the focus of this entire report accordingly.

Ponds are generally classified according to the way they store and route runoff. In the broadest terms, stormwater basins are either wet or dry – having a permanent pool of water or not. Whether a pool is permanently wet or dries up between rain events is a function of connectivity with surface and ground waters, soil properties, and design of outlet structures. For instance, under similar geologic settings, a permanent pool of water can be retained by setting outlet structures at higher elevations than for a dry detention basin (Persson & Wittgren 2003; Weiss et al. 2007). For wet basins, the pond is sized to have a temporary storage volume above a permanent pool of water (DHEC 2005). For dry basins, stormwater peak flow is attenuated, with the runoff draining over the course of at least 24 hours and typically less than 48 hours (Weiss et al. 2007; Ellis et al. 2014).

Nuisance flooding is a hazard to property near ponds (Fennessey et al. 2001) that may result from a lack of understanding of the hydrology of ponds in the coastal zone. Property can flood following the construction of a large adjacent impervious area when a pond is constructed directly upslope, as is done for in-fill residential or commercial development. This may instigate a local increase in the water table if the pond is not lined with low permeability clay or geotextile to minimize groundwater seepage, or if an extreme event causes pond overflow or bank failure. These hazards can occur when the site plan over-predicts the predevelopment runoff of a watershed and/or uses property boundaries to define the catchment area for the ponds, as is commonly done (Fennessey et al. 2001). The ultimate result of improperly sized stormwater basins may be an additional financial burden on community residents impacted by flooding or increased maintenance.

The role of engineering design on water detention with respect to size and volume is well-studied. However, in the dynamic and rapidly-developing coastal zone of S.C., combining site details such as topography, pond design details, and runoff predictive models that include soil data such as the Curve Number approach (Fennessey et al. 2001) would inform best management practices for stormwater managers to plan for and mitigate stormwater inputs. A needed area of investigation is the design “lifetime” of ponds with respect to sedimentation. Rapid urbanization and increase in impervious cover are likely to result in increased sedimentation in ponds. A site-specific study at a development on Daniel Island, Berkeley County near Charleston, S.C. found that in just a few years a single stormwater pond lost 36% of its usable volume and a series of linked ponds lost 15% of the system’s total usable volume (Messersmith 2007). It is not known if this is a widespread issue for the entire region or a site-specific exception. Existing studies have identified potential threats to pond life associated with these sediments and other particles as a mode to convey pollutants into ponds (Zoumis et al. 2001; Weinstein et al. 2010a; Frost et al. 2015).

Research Gaps

  • Investigate the role of soils and geology of the SECP on pond design and function.
  • Understand which management practices are important to help avoid sedimentation problems in ponds.
  • Examine the time scales required for stormwater retention to protect water quality.
  • Determine the implications of reduced pond holding capacity and potential surface–groundwater interactions due to sedimentation.
  • Evaluate the impacts of pond design elements (e.g., lined vs. unlined, pond size relative to catchment area) on stormwater management efficacy.
  • Examine the capacity of ponds to effectively function in the future under scenarios of changing precipitation patterns and temperature.

2.4 Transport of Contaminants in Runoff

2.4.1 Particulate Matter

Suspended particulate matter (PM) makes up a significant portion of surface runoff contaminant load. Major nonpoint sources that result in PM in runoff include roads, construction and urban developments, soil erosion, deforestation, agricultural activities, and aerial deposits. Urban areas contribute significantly to runoff pollution load to ponds due to higher human density (Wong et al. 2012). Particles can have a distinct geochemical or anthropogenic composition (e.g. mineral, organic, or synthetic) depending on the source (Frost et al. 2015). Size and morphology of these particles in runoff are highly variable depending on the intensity of rainstorms as well as the catchment type (Characklis & Wiesner 1997; Sansalone & Buchberger 1997a, b; Sansalone et al. 1998; Kim & Sansalone 2008).

Emerging contaminants such as nanomaterials (e.g., Das et al. 2012) and microplastics (e.g., Teuten et al. 2007; Andrady 2011; Cole et al. 2011) are becoming ubiquitous in estuarine and coastal environments from anthropogenic sources, but these types of particles have been understudied thus far in ponds (Chapter 3). Larger-sized materials including garbage and trash are also often part of runoff in highly developed areas, leading to clogged water conveyances or compromised pond function. In addition to PM itself being a potential hazard to aquatic life, the nature of the suspended particles in stormwater and the sediment bed in stormwater ponds influences contaminant transport, fate, and effects.

2.4.2 Inorganic Chemicals

Trace metals are common contaminants in urban environments and can be present in dissolved form or associated with PM in runoff (Wong et al. 2006; Fletcher et al. 2013). The most common trace metals in dissolved forms reported in runoff include lead, copper, cadmium, and zinc (Characklis & Wiesner 1997; Sansalone & Buchberger 1997a, b; Davis et al. 2001; Kamalakkannan et al. 2004; Brown & Peake 2006; Bentzen & Larsen 2009; Karlsson et al. 2010; Frost et al. 2015). Specific sources reported include building siding and roofs (e.g., Quek & Förster 1993) and automobile brakes and tires (Laxen & Harrison 1977; Ellis et al. 1987; Sansalone & Buchberger 1997b; Davis et al. 2001). Dry and wet deposition can be significant sources of trace metals in urban areas (Nriagu 1989; Sabin et al. 2005; Fulkerson et al. 2007; Eckley & Branfireun 2008). Radioisotopes and stable isotopes of trace metals are often employed in source-tracking of trace metals found in various environments, including soils and lake sediment, and may apply to ponds (Quast et al. 2006; Thapalia et al. 2010; Louchouarn et al. 2012; Wiederhold 2015).

Other inorganic chemicals in stormwater include major elements and ions such as sodium (Na), potassium (K), calcium (Ca), magnesium (Mg), chloride (Cl), and sulfate (SO4), as well as nitrate (NO3) and phosphate (PO4), which are important nutrients. These chemicals are often present in relatively high concentrations and may be derived from the weathering of inorganic minerals present in dry deposits, road dust, eroded soils, road applied salts (in the case of inclement weather), and fertilizer applications (Göbel et al. 2007). In addition to dissolved forms, certain inorganic chemical species are often associated with sediment particles, which can be major vectors in transport of these chemicals (Bryan & Revitt 1982; Sansalone et al. 1995; Characklis & Wiesner 1997; Sansalone & Buchberger 1997a, b; Kamalakkannan et al. 2004; Jartun et al. 2008).

2.4.3 Organic Chemicals

Fate and transport of organic chemicals is more complex compared with that of inorganic chemicals and sediment due to their chemistry, which can include transformations and distribution between solid, dissolved, and gaseous forms (Jaffé 1991). Common organic chemicals in urban areas that are ubiquitous in runoff are derived from automobile and other combustion-derived oil and gas residues, asphalt, and coal tar-coated pavement and roofs (Fam et al. 1987; Van Metre et al. 2000; Mahler et al. 2012). These organic mixtures can be predominately comprised of polycyclic aromatic hydrocarbons (PAHs), a class of nonpolar organic chemicals which exhibit relatively low vapor pressure and low aqueous solubility (Hoffman et al. 1985; Stout et al. 2001; Menzie et al. 2002; Kamalakkannan et al. 2004; Brown & Peake 2006; Hwang & Foster 2006; Boving & Neary 2007; Göbel et al. 2007; Jartun et al. 2008; Bentzen & Larsen 2009; Watts et al. 2010; Birch et al. 2012; Mahler et al. 2012). Because of these properties, PAHs are more likely to be associated with PM in surface runoff (Kamalakkannan et al. 2004; Jartun et al. 2008; Bentzen & Larsen 2009; Birch et al. 2012).

Other non-polar organic chemicals, such as polychlorobiphenyls (PCBs), polychlorinated dibenzodioxins (dioxins), and polychlorinated dibenzofurans have similar chemical properties as PAHs, are anthropogenic in origin, and are often found in industrial-urban runoff (Eriksson et al. 2007; Cornelissen et al. 2008; Howell et al. 2011a, b). These chemicals are also more likely to be associated with sediment in runoff (Zgheib et al. 2011). Fire retardant chemicals such as polybrominated diphenyl ethers (PBDEs) are another class of chemicals that are being increasingly found in urban runoff associated with organic-rich particles (Oram et al. 2008; Gasperi et al. 2014). Herbicides such as atrazine, alachlor, and glyphosate are relatively common in agricultural runoff (e.g., Leonard et al. 1979; Thurman et al. 1991; Logan et al. 1994), but herbicides are also commonly used in urban areas to control unwanted vegetation among commercial and residential developments and along public roads and highways (e.g., Revitt et al. 2002; Huang et al. 2004). Pesticides and insecticides such as diazinon, carbaryl, and malathion are also increasingly found in urban runoff (e.g., Hoffman et al. 2000; Weston et al. 2009). Concentrations of pesticides in ponds in coastal S.C. have been positively correlated with high temperatures and rainfall, and also correlated with concentrations in adjacent tidal creeks (e.g., pyrethroids, imidacloprid), which demonstrates potential connections between runoff, ponds, and receiving waters (Serrano & DeLorenzo 2008; DeLorenzo et al. 2012).

Emerging contaminants such as pharmaceutical chemicals, hormones, and sunscreen/ultraviolet filters are now ubiquitous in runoff (Richardson & Ternes 2011). These chemicals are structurally very complex and are increasingly found in natural waters resulting in ecosystem disruption (Kümmerer 2004), yet studies on the transport of these chemicals to ponds are needed.

2.4.4 Pathogens

Pathogenic organisms are increasingly common in runoff (Geldreich 1996; Mallin et al. 2000; Grant et al. 2001; Han et al. 2006; Arnone & Walling 2007; Mallin et al. 2009; Parker et al. 2010; Rich & Maier 2015). These organisms include bacteria, enteric viruses, protozoa, and parasites and come from a variety of sources, including humans (e.g., septic system failure), pet animals, farm animals, and wildlife. Because of the range of origins for these organisms in the environment, there are a variety of source-tracking methods available to identify specific sources (Whitlock et al. 2002; Meays et al. 2004; Webster et al. 2004; Siewicki et al. 2007; Parker et al. 2010; Rock et al. 2015). Due to the wide variety of pathogenic organisms present in these environments, common fecal-sourced organisms such as fecal coliforms, E. coli, and enterococcus are often used as indicators of contaminated surface runoff (Arnone & Walling 2007; Gerba 2015b). Enteric viruses such as enterovirus and adenoviruses are more difficult to track, but specific viruses may also be used as indicators of compromised water quality (Fong & Lipp 2005). Studies have shown that environmental pathogens are often naturally associated with soil in the landscape and are mobilized by runoff (Ahn et al. 2005; Surbeck et al. 2006; Parker et al. 2010). Survival of these organisms is greatly dependent on the environmental conditions such as water temperature, UV light penetration, pH, dissolved salts, organic matter, and biological factors such as existing microflora (Gerba 2015a). Adding to the complexity, dynamic communities of bacteria and viruses are highly variable in individual ponds, even those in close proximity (Saxton et al. 2016).

Research Gaps

  • Understand the nature of contaminants carried by runoff and the sources of those contaminants in order to
    better design and manage ponds.
  • Understand the interaction of contaminants in ponds, which is crucial considering trends in increasing
    urbanization in the coastal region.

2.5 Transformation and Fate of Contaminants in Ponds

2.5.1 Biological Processes

Biological processes leading to biological uptake and transformation of chemicals alter the budget of contaminants in ponds. These processes are highly dependent on the specific contaminant as well as other conditions in the water column (e.g., pH, dissolved oxygen, salinity, organic carbon content, redox status, temperature) and biological community structure (Mihelcic & Luthy 1988; Mueller et al. 1989; Leahy & Colwell 1990; Das & Chandran 2010). Different microorganisms can adapt to contaminated environments, resulting in the creation of specific conditions that biodegrade complex organic chemicals (Das & Chandran 2010; Maier & Gentry 2015) or alter the bioavailability of organic and inorganic chemicals, including toxic metals (Roane et al. 2015).

Contaminants may also be removed from the water column via uptake into plants near or within ponds by a process commonly referred to as phytoremediation (Alkorta & Garbisu 2001; Susarla et al. 2002; Pulford & Watson 2003); however, for chemicals that are simply stored on or within the plants to be effectively removed from the system and not re-released following plant turnover, vegetation needs to be harvested at the end of the growing season. Vegetation can also aid contaminant removal in ponds by reduction of water velocity and entrainment of sediment, or adhesion to plant surfaces. Although ponds have been observed to remove bacteria from runoff, in-pond conditions favorable for bacterial growth could lead to a proliferation of harmful species and higher concentrations in the effluent of ponds (DeLorenzo & Fulton 2009; Hathaway et al. 2009; Grebel et al. 2013). Ponds may also attract wildlife, leading to inpond sources (Hathaway et al. 2009), which may show seasonality depending on wildlife migration patterns (Phillips 2014).

Biological processes contributing to contaminant removal depend on the habitat the pond provides. Ponds in coastal S.C. that are tidally influenced have been observed to range from low brackish to marine salinities, and this affects specific bacteria assemblages (DeLorenzo & Fulton 2009). Benthic invertebrate communities will also be affected by salinity tolerance. Action by these invertebrate communities to bioturbate sediments in detention ponds could impact hydraulic conductivity (and therefore water infiltration through sediment bed), oxygen availability in sediment pore waters, and sediment-water pollutant exchange (Nogaro & Mermillod-Blondin 2009).

2.5.2 Chemical Transformations

Values for chemical transformation rates under specified conditions may be found for a small subset of potential pollutants in the literature (e.g., wastewater treatment, environmental chemistry, and engineering) or may be modeled based on chemical structure (e.g., U.S. EPA EPI Suite). Chemical transformation in stormwater treatment structures has been modeled using first-order kinetic decay assumptions, with removal mechanisms (e.g., physical, chemical, or biological) not treated individually (Elliott & Trowsdale 2007). However, pond design characteristics, such as dimensions, that influence environmental conditions (e.g., temperature and redox conditions) should also affect the transformation rate, and therefore these rates should not be applied universally. Stratification and eutrophication can impact the availability of oxygen in pore waters and subsequently the redox conditions, which control the processing of nutrients, speciation of metals, and degradation of organic compounds. The depth of the pond permanent pool will influence light penetration for the photochemical transformation of certain organic contaminants. Shallow openwater treatment cells have been designed for photolytic degradation of certain organic contaminants, but may not be applicable to ponds due to low light penetration to sediment bottoms and growth of shading, floating vegetation such as duckweed (Jasper & Sedlak 2013). For contaminants with relatively high vapor pressures for which volatilization is a potential removal pathway from pond waters, important pond features are the surface area of the pond and surrounding open area generating wind across the pond surface, since these factors will influence evaporation of water and temperature which affect volatilization. While there are few studies that measure rates of photodegradation and volatilization from ponds, lessons can be applied from studies that have focused on the treatment of chemicals in wastewater treatment ponds and constructed wetlands (Oulton et al. 2010; Jelic et al. 2011).

2.5.3 Sorption Processes and Role of Particles in Contaminant Removal within Ponds

The basic environmental geochemistry that explains sorption and complexation processes is well established (Stumm & Morgan 1996; Langmuir 1997; Schwarzenbach et al. 2003; Appelo & Postma 2005; Luoma & Rainbow 2008). Sorption processes include surface complexation, ion exchange, fixation, surface precipitation, and partitioning (Schwarzenbach et al. 2003; Bradl 2004; Brown & Calas 2012). Because of the versatility of some of the different sorption mechanisms, chemical sorption occurs in a wide range of sorbent material, from suspended materials to bed sediment, and including particulate and dissolved organic matter, mineral phases, and tissue of living organisms (McCarthy & Zachara 1989; Schwarzenbach et al. 2003; Brown & Calas 2012). Dissolved and particulate organic carbon fractions of natural organic matter (NOM) from disparate sources are likely to be found in runoff and ponds. While these forms of NOM are not primary contaminants in water, increased presence in runoff can behave as vectors for other contaminants or eventually impact contaminant fate in streams and ponds, due to sorption and complexation of sparingly soluble chemicals (e.g., Chiou et al. 1986; Allen-King et al. 2002). “Black” carbon, particles formed by incomplete combustion such as soot, also impacts contaminant fate due to strong sorptive interactions.

Trace metals can form strong complexes to specific components in sediment such as clay minerals, mineral oxides (iron [Fe], aluminum [Al], or manganese [Mn] oxides), and organic fractions. Relative sorption control or strength in the matrix depends on factors such as relative availability of the different sediment components, sorption kinetics, bulk solution conditions (e.g., pH, ionic strength or salinity, redox conditions) and particle surface characteristics (Stumm & Morgan 1996; Langmuir 1997; Appelo & Postma 2005; Du Laing et al. 2009; Frost et al. 2015). Ionic forms of trace metals, inorganic ions, and also organic ions in aqueous environments readily complex with other ions present in water and to a variety of inorganic mineral or organic surfaces.

Nonpolar and/or neutral organic chemicals strongly sorb to PM surfaces and/or partition into nonpolar phases including soils, suspended organic matter and sediment, and organisms (Karickhoff 1984; Schwarzenbach et al. 2003). The exact nature of sorption depends on the chemical structure of the sorbate in addition to the material sorbent and is typically related to aqueous solubility (Allen-King et al. 2002). In addition to the sorption mechanisms, reaction kinetics play a significant role in the distribution of chemicals in dynamic environmental systems, such as stormwater ponds. Temperature, salinity, and dissolved humic substances can additionally strongly impact the sorption of nonpolar and neutral organic molecules (Tremblay et al. 2005).

Pathogens are also more likely to be sorbed, or attached as biofilms, to PM in aquatic environments (Ahn et al. 2005; Cizek et al. 2008; Jenkins et al. 2012). This association with sediment increases the survival of the pathogenic organisms and lowers predation by other microorganisms (Gerba & McLeod 1976; Schillinger & Gannon 1985; Evanson & Ambrose 2006; Jeng et al. 2005; Bonilla et al. 2007; Jenkins et al. 2012;). The type of sediment (presence of certain mineral oxide surfaces and organic matter), aqueous geochemical conditions (e.g., pH, salinity, dissolved oxygen, water temperature, alkalinity, and redox conditions), and seasonal conditions also play a significant role in survival of pathogens in the sediment (Van Donsel et al. 1967; Scholl & Harvey 1992; Jimenez-Sanchez et al. 2015). These geochemical conditions in stormwater ponds can be highly variable and can impact storage versus further transport of pathogens out of ponds and downstream. Stormwater runoff has been shown to transport enough fecal bacteria loads to close coastal S.C. shellfisheries, and high levels of bacteria have been recorded in runoff and S.C. ponds following heavy rainfall events (Drescher et al. 2007).

Physical and chemical characteristics of particles, such as size, mineral makeup, organic content, surface area, and bulk density, impact mobility and sorption and therefore also impact transport of contaminants associated with particles (Zanders 2005). For example, small and less dense particles (compared with quartz mineral grains, whose density is approximately 2.65 g/cm3) are often more mobile as they are not easily trapped by vegetation and can be carried by flowing water at lower velocities. Smaller-sized particles (e.g., less than 300 μm diameter) are often associated with high concentration of contaminants (Vaze & Chiew 2004; Zanders 2005). Silt and clay fractions are two common classes of inorganic suspended particles in streams and ponds (Slattery & Burt 1997). While silt-sized fractions can be a combination of quartz and other silicate minerals, clay-sized fractions (less than 2 μm diameter) tend to be predominantly aluminosilicates (Langmuir 1997). Specific mineralogy is reflective of the soil mineralogy of the region subject to erosion by surface runoff (Wall & Wilding 1976; Carter et al. 2003; Taylor & Owens 2009). These properties also control the affinity and kinetics of interactions between the contaminant and solid phase (via chemical interactions or biological niche in the case of biological agents), as well as particle settling potential. Contaminants may originate on PM, become particle-bound during transport, or sorb once within the ponds. And given the sorption processes described, PM phases play a significant role in transporting, sequestering, or mobilizing contaminants in freshwater, estuarine, and coastal waterbodies (Turner & Millward 2002; Walling 2005; Fries et al. 2006; Taylor & Owens 2009; Crompton 2015; Frost et al. 2015).

2.5.4 Sedimentation Processes

Once runoff reaches ponds, the water velocity slows, allowing ponds to serve as basins that trap particles by sedimentation and filtration. As described above, it is the ability of ponds to trap particles and associated contaminants that largely determines whether ponds improve water quality. According to the International Stormwater Best Management Practice (BMP) Database, median pollutant removal rates range from 17 to 96%, depending on the pollutant type (Clary et al. 2017). Pond performance is often better with particulate than dissolved pollutants, and ponds are better at removing suspended solids than some other types of BMPs (Barrett 2008). The removal efficiency of suspended solids, total nitrogen (N), and total phosphorus (P) may be greater in eutrophic than mesotrophic ponds (Borden et al. 1997; Winston et al. 2013), potentially due to phytoplankton N and P uptake.

Sedimentation of particles in ponds is an important process (U.S. EPA 1999; Bentzen & Larsen 2009; Hathaway et al. 2009) and is a function of the density, size, and buoyancy of particles and the flow regime in relation to the size, depth, and pathways that water takes in the pond. Thus, where particles and associated contaminants drop out or are filtered out is dependent on both particle source and pond design traits affecting hydrodynamics. Large and dense particles are more likely to be deposited in ponds and near inlets since they typically have faster-settling velocities. However, other conditions such as fluid density and viscosity, laminar vs. turbulent flow conditions, particle shape, and current flow have to be explicitly considered when making quantitative predictions of transport of sediment and associated contaminants in water (Julien 2010). Organic particles are generally slower to settle and less effectively removed, depending on pond design. Pollutant distribution between dissolved and particulate phases and the ability of ponds to capture particles strongly control pollutant removal within ponds. For instance, Crawford et al. (2010), in a survey of 16 coastal S.C. ponds, found that levels of Al and Cd were higher in pond inlets, while levels of Cu, Pb, and Zn were higher in the center of the ponds. This suggests the affinity for certain metals to adsorb to finer clay-sized particles and thus travel further from the inlet before sedimentation occurs. Clay and organic carbon content also were generally seen to be higher near the center of ponds than near the inlet (Crawford et al., 2010).

Although some removal occurs during the dynamic period when runoff enters the pond, the U.S. EPA reports that 90% of pollutant removal occurs during the quiescent period between rainfall events when particles settle to the bottom of the detention pond (U.S. EPA 1999). Generally, hydrocarbon removal is related to suspended sediment removal and, in one study, ponds had a median removal rate of 80 – 90%, again illustrating the importance of sedimentation (Schueler 2000).

2.5.5 Role of Pond Design in Contaminant Fate

Hydraulic residence time (RT) is a parameter that controls sedimentation and biogeochemical transformation. It is determined by pond dimensions and water discharge. Residence time has been cited as the factor that is most limiting to water quality improvement in ponds (although other factors, such as redox environment and plants are notable for certain constituents) (Mallin et al. 2002; Weiss et al. 2007, 2008). For instance, Comings et al. (2000) found that a pond that retained stormwater seven times longer (RT = 1 week) than a reference pond allowed for superior pollutant removal. Carpenter et al. (2014) found that adding a sluice gate to the outlet of a detention pond allowed for a longer RT, which resulted in improvement of removal efficiency for total suspended solids from 39 to 90%, ammonianitrogen from 10 to 84%, and zinc from 20 to 42%. It has also been suggested that stormwater structures with low flow conditions maximize protection of downstream assets from pesticides (Howitt et al. 2014).

Although RT in S.C. ponds is generally long according to some previous studies (Lewitus et al. 2003; Bunker 2004; Brock 2006), performance assessments of ponds elsewhere in VA indicated that the holding capacity of these systems were often less than 24 hours (Hancock et al. 2010). However, this is an area of needed research to substantiate the conditions leading to vulnerable conditions. While many of the above studies have shown ponds to be broadly effective at removing and retaining pollutant loadings, specific removal efficiencies among studies are highly variable. The reasons for this variability are not well understood but are likely at least partially driven by pond geometry (Wu et al. 1996; Borden et al. 1997; Mallin et al. 2002), proximity to rural vs. urban areas (Tufford & Marshall 2002), and hydrology (Persson & Wittgren 2003; Huda & Meadows 2010). Size, depth, and shape are important design criteria that have been observed to affect pollutant removal in ponds. In a study on Daniel Island, S.C. researchers found a series of linked ponds had longer RT and superior pollutant removal capabilities as compared to a single pond on the site (Messersmith et al. 2007). Comings et al. (2000) found that the larger a pond is in relation to the area it drains, the better it performs at removing pollutants, and furthermore, that pollutant removal can be improved by moving weirs to direct flow in a tortuous pathway through treatment cells. As shown in this study and other work on the basic functioning of water treatment facilities in general, the path of stormwater through a pond should not be in a direct line, which promotes short-circuiting. Ellis et al. (2014) and S.C. DHEC (2005) designate the criteria for pond dimension as at least a ratio of 1.5:1 for pond length to width, although a ratio of 3:1 affords better water quality treatment. Pond surface area was able to explain trapping efficiency for PAHs in coastal S.C. ponds, although whether the effects were due to larger ponds being able to settle solids more effectively or larger size reducing the susceptibility of sediment bottom to scour during high flow events was undetermined (Weinstein et al. 2010a). Adequate wet detention pond depth allows for effective sediment gravity settling (Weiss et al. 2007). An additional advantage is that the sediment bottom, which potentially stores high levels of nutrients and sediment-bound contaminants, is buffered against washout of sediments during the next storm event by the overlying water column, or “permanent wet pool.” A pool depth of 4 to 6 feet is ideal as per S.C. DHEC (2005). Vegetated ponds and stormwater wetlands have the additional advantage of plants providing a buffer to slow water flow and structure to hold sediments in place through the growth of root systems. However, periods of high flow create turbulence or shear force, which leads to sediment resuspension and can cause the release of contamination to receiving waters (Julien 2010).

Hydrophobic organic contaminants with strong sorption and low water solubility have been observed to concentrate in sediments (Weinstein et al. 2010a; Roinas et al. 2014); however more polar and water soluble organics will be increasingly mobile and may pass through the pond depending on how water is routed (Neary & Boving 2011; Roinas et al. 2014). Pond design often includes a forebay to trap the majority of the coarse suspended solids fraction prior to flow entering the main body of the pond, although finer sediments will pass through.

Pollutant Wet Detention Pond
(S.C. DHEC, 2005)
Wet Retention Pond
(International Database)
Total Suspended Solids 65-80% *75%
Total Copper 40-65% *52%
Total Zinc 50-75% *56%
Metals 35-75% **17-75%
Total Lead 60-85% *67%
Total Phosphorus 50-70% *55%
Total Nitrogen 35-45% *23%
Bacteria 45-75% *54-96%

*Indicates that the median removal capability of wet ponds was found to be significant (p < 0.05; non-parametric Mann Whitney test) (Clary et al. 2017).
**With the exception of arsenic, all metals were found to be removed at significant rates.

Table 2.1 Average pollutant removal capabilities summarized from S.C. DHEC (2005) and the International BMP Database (Clary et al. 2017), which classifies basins with a permanent pool as retention ponds. The S.C. DHEC manual cites the range of average removal capabilities while the International database provides median values. For “metals” the range is provided because the removal efficiency differed depending on the type (e.g., nickel, chromium) and the range for “bacteria” represents differing measures (e.g., E. coli, fecal coliform).

The overall performance of ponds in removing or sequestering contaminants is very site-specific – there is no clear formula for how all ponds behave. For example, S.C. DHEC (2005) lists the large ranges of pond removal capabilities for a number of common stormwater pollutants (Table 2.1). The International BMP database, which synthesizes data from hundreds of pond samples, also shows there is much variability in removal capabilities, which is difficult to constrain due to site differences (Clary et al. 2017).

Regular removal of the accumulated sediments from ponds is necessary to minimize the risk of contaminated sediments affecting benthic habitat quality, in-pond and downstream ecosystem health, and groundwater quality, and to maximize the operational efficiency of the pond (Durand et al. 2005). Dredging of sediments is recommended when the permanent pool volume is significantly reduced beyond useful design specifications, but there are no state regulations requiring dredging (Rollins & Powell 2013). Further illustrating the role of sediments to capture contaminants in ponds, Weinstein et al. (2010b) report that, depending on local and state regulations, excavated pond sediments are typically treated as either solid waste or hazardous waste and transported to either landfills or hazardous waste facilities for disposal, respectively. Currently, in S.C. there are no requirements that pond sediments be tested for chemical or biological contaminants prior to sediment removal (Ellis et al. 2014). This topic is covered in more detail in Chapter 3 of this report.

2.5.6 Modeling of Stormwater Processes at the Landscape Scale

Knowledge of transport, transformation, and storage has been translated into models that aim to predict pollutant concentrations and loads in stormwater systems to inform engineering and management (Elliott & Trowsdale 2007; Vezzaro et al. 2011, 2012; Fletcher et al. 2013). Integrated modeling of stormwater quality involves simulating source generation, the flush of pollutants from catchment surfaces and how they are channeled through surface water drainage systems, and final treatment in ponds. These steps are modeled as functions of hydrologic factors and pollutant characteristics, emphasizing that hydrology and biogeochemistry are fundamentally linked. The majority of existing models on stormwater treatment include contaminant settling in ponds using sediment settling theory, first or second order decay, removal fractions or output concentrations, or other variables such as flow rate (Elliott & Trowsdale 2007). It has been shown that improvement of models will require more monitoring data for calibration and field testing and improved understanding of fate processing for a wider selection of environmental contaminants (Bertrand-Krajewski 2007; Elliott & Trowsdale 2007; Scholes et al. 2008). A study of micropollutant removal abilities by stormwater ponds found the most sensitive model parameters were those related to physical processes such as flow and settling/resuspension (e.g., total suspended solids) (Vezzaro et al. 2011). However, at present, research has not identified an optimum set of parameters for employing stormwater quality models (Bertrand-Krajewski 2007). Increasingly, models are also considering stormwater as a resource, though the performance of stormwater technologies in restoring water quality remains poorly quantified. As models improve, they will also consider the uncertainty associated with climate change that requires all water management systems to be resilient to change (Fletcher et al. 2013).

Research Gaps

  • Understand the major controls that influence contaminant fate in pond systems, specifically for the Southeastern coastal region.
  • Assess nutrient and biological exchanges between ponds and tidal creeks.
  • Integrate knowledge on hydrologic function with engineering principles, contaminant behavior, and biological interactions to advance our understanding of pond systems.
  • Explore the role of groundwater for nutrient transport, and concurrent influence of soil type and landscape on subterranean movement.

2.6 Summary and Recommendations

Significant portions of the SECP have been altered due to development, affecting the hydrology and water balance through re-routing of water from infiltration and evapotranspiration to surface runoff and storage in wet detention ponds constructed in the landscape. This review summarizes various processes and mechanisms that are likely to occur in ponds using our best understanding of hydrology and biogeochemical principles. Hydrology and contaminant behavior in ponds is very complex and, as summarized in the sections above, there is still more research that needs to be done to more fully understand the dynamic behavior of these systems. For instance, we were unable to clearly establish contributions of ponds to groundwater contamination, though it is of growing concern and requires investigation (Walsh et al. 2005; Roy & Bickerton 2012). While a basic understanding of pond hydrology is fundamental to our conceptual understanding of runoff and pond function, no systems-level studies that address all relevant mechanisms were identified over the course of our literature review.

Since the implementation of federal control regulations and guidelines in the 1990s to manage stormwater, careful calculations of pond design have been enacted in the field based on straightforward engineering principles. However, while it is widely recognized that ponds can reduce contamination of the ultimate receiving water bodies for certain classes of contaminants (S.C. DHEC 2005), ponds were not specifically designed for post-construction phase water quality improvements (Ghermandi & Fichtman 2015). Pond designs that include elements that actively reduce water contamination could help address the growing need for contaminant reductions of nonpoint sources to final receiving water bodies (Walsh et al. 2005; Clary et al. 2008; Collins et al. 2010). Pond water quality and the ability to remove pollutants and protect downstream water bodies are affected by pond location in the landscape. We need a better understanding of how diverse engineered features impact hydrological and biogeochemical processes in order to inform policies and/or regulations that govern pond design and management practices (Chiandet & Xenopoulos 2016; Johnson & Sample 2017). Because of the variability in pollutant removal efficiencies, an audit of design details to measured performance is needed, as is an analysis of how changing precipitation patterns in the future may impact functionality. Studies that categorize the strengths, weaknesses, and vulnerabilities of stormwater ponds in the southeastern coastal zone are progressing, but more comprehensive work is needed. Therefore, we recommend concerted, stepwise approaches to integrate and improve the understanding of stormwater pond functioning as it pertains to the unique biological, hydrological, and geochemical settings in each region of coastal S.C.

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