Chapter 3 –An Assessment of Nonpoint Source Pollution in Stormwater Pond Systems in Coastal South Carolina
Mohammed Baalousha, Ph.D., Samantha McNeal, and Geoffrey I. Scott, Ph.D. Department of Environmental Health Sciences, Arnold School of Public Health, University of South Carolina, Columbia, S.C.
Corresponding Author: Geoffrey I. Scott, Ph.D. (firstname.lastname@example.org)
In this chapter, we seek to demonstrate why urban nonpoint source (NPS) pollutants carried by stormwater runoff are a major concern for human health and the health of coastal ecosystems, as well as the role ponds can play in mitigating this impact. Water quality in the United States is well monitored particularly around point sources of pollution. However, data from current monitoring programs indicate that NPS pollution is a growing problem and causes two-thirds of the total amount of pollution impacting water quality. This is attributed to various contaminants including metals, organic materials, and microbes (Wang 2015). Runoff results in approximately 40% of U.S. rivers, lakes, and estuaries falling below the basic standards for recreational uses, and South Carolina is no exception (Sparling et al. 2007). Contaminants are generated from various anthropogenic activities and are transferred from urban, industrial, commercial, agricultural, and rural areas into estuaries and surface waters via runoff (Good 2014). Ponds are engineered Best Management Practices (BMPs) that protect downstream systems, though low water circulation and the high pollutant load contribute to water quality problems within pond basins (Novotny 1995). The NPS runoff discharged into ponds may emanate from fertilizer and pesticide applications, fecal material input from humans, pets, and wildlife, and runoff from domestic wastewater, driveways, and roads (WHO 1999) (Figure 3.1).
Figure 3.1 Some major sources of NPS pollution loading within urban coastal areas of the U.S. that can contribute to runoff, some of which is discharged into ponds. Source
Urban runoff pollution can have many deleterious effects on plants, fish, animals, and humans. For example, untreated urban runoff from an auto recycling facility into waterways near Los Angeles, California over several years killed 20% or more of the exposed minnows (Swamikannu 1994). And a study in San Francisco Bay found higher levels of both legacy pollutants and Contaminants of Emerging Concern (CECs) in NPS runoff than in waste water treatment plant (WWTP) effluent (SFEI 2013).
Here, we review chemical and microbial pollution sources within coastal S.C. To assess the scope of each group of contaminants, data on concentrations and location (ponds vs. natural systems) in addition to a range of environmental parameters were extracted from the peer-reviewed literature and government sources (e.g., NOAA, USGS). A database was then developed using these data for each group of contaminants in surface waters and sediments, as appropriate. The topic sections that follow summarize information about the major pollutants found in ponds and what risks these contaminants may pose to environmental and human health.
3.2.1 Literature Review
Most pond chemistry data were taken from a study conducted by researchers at The Citadel and reported in a National Oceanic and Atmospheric Administration (NOAA), S.C. Sea Grant Consortium (SCSGC), and S.C. Department of Health and Environmental Control (DHEC) joint report by Weinstein et al. (2008) and later published in a series of peer-reviewed papers (Crawford et al. 2010; Weinstein et al. 2010a, 2010b). This work constituted a detailed analysis of pond sediments from 16 of the 112 ponds reported on in the DHEC 2007 state of knowledge report (Drescher et al. 2007). Samples were taken from two locations: the pond inlet and pond center. Here, the different pond types reported in these studies were converted into the current pond classification system developed by Smith et al. (Chapter 1) and include the following urban site types: residential, mixed land use, golf course, and commercial ponds. We also retained a fifth type, which was reference ponds located in rural or undeveloped areas (two of these ponds were sampled in the report). According to the 2013 inventory, these rural ponds comprise 57% of all ponds in the eight
coastal counties (Chapter 1). Microbial pollutant data from South Carolina Urbanization and Southeastern Estuarine Systems (USES), and the Land Use Coastal Ecosystem Study (LUCES) summarized in Scott et al. (2008), along with studies by Johnston et al. (2009), Serrano & DeLorenzo (2008), Berquist et al. (2010), and DeLorenzo et al. (2012), are discussed in the context of the older studies by Drescher et al. (2007) and Messersmith (2007). A number of references on contaminant levels in estuarine and tidal creek sites in S.C. are also included and used for comparison with pollutant levels reported in ponds. Data – including location, land use types, chemical, or microbial pollution data, and any other relevant information from each study – was entered into a database. In total, our citations include 111 references as 66 primary literature sources, nine books, four conference presentations, 30 technical reports, and two student theses.
An environmental risk assessment was conducted using methods described by the Environmental Protection Agency (EPA 2004), Holland et al. (2004), and the Interstate Shellfish Sanitation Conference (ISSC) (2004). Simply put, NPS contaminants quantified in runoff, surface waters, and sediments were ordered from lowest to highest concentration to generate a cumulative frequency exposure curve for each pollutant type. This curve was plotted versus environmental standards or guidelines in water or sediments for each contaminant, including published EPA water quality criteria (Acute and Chronic Water Quality Criteria) and Sediment Quality Guidelines (SQGs) (Long & Morgan 1990; Long et al. 1995; MacDonald et al. 1996; Long & MacDonald 1998) (Table 3.1). Both Probable Effect Level (PEL), the concentration above which adverse effects on survival or growth are expected to occur frequently, and the Threshold Effect Level (TEL), the concentration below which adverse effects are expected to occur only rarely, are reported for each contaminant listed. In our discussion we mainly use the designations of Effects Range Low (ERL), the threshold concentration where adverse effects begin to occur, and Effects Range Median (ERM), the median concentration where adverse effects are more certain to occur. These threshold levels correspond to the Incidence of Adverse Effects (IAE), which is the estimated toxicity observed in sediment toxicity tests by Long et al. (1995). The IAE is expressed as a percentage (%IAE), and this value corresponds to the number of data entries within each concentration range in which biological effects were observed in benthic indicator species, divided by the total number of entries within each range.
|SQG Concentrations1||% Incidences of Adverse Effects|
|Chemicals||TEL||PEL||ERL||ERM||< ERL||> ERL and < ERM||> ERM|
|Trace metals (mg/kg dry weight or ppm)|
|Pesticides/PCBs (μg/kg dry weight or ppb)|
|PAHs (μg/kg dry weight or ppb)|
|SUM LMW PAH2||NG||NG||552||3,160||13%||48.1%||100%|
|SUM HMW PAH3||NG||NG||1,700||9,600||10.5%||40%||81.2%|
1SQG= Based on NOAA’s national Sediment Quality Guidelines
2SUM LMW PAH = Concentration of all low molecular weight PAHs
3SUM HMW PAH = Concentration of all high molecular weight PAHs
4NG = No guideline values
Table 3.1 The Sediment Quality Guidelines (SQGs) concentrations and Incidences of Adverse Effects as shown by Long et al. (1995) and used in this report to predict adverse effects on benthic and epibenthic estuarine/marine fauna.
For legacy pollutants and fecal coliform bacteria (FCB), we have both SQGs and Water Quality Criteria to compare reported concentrations and to provide an initial assessment of risk and hazards. For select CECs, there are reported best available risk assessment guidelines provided by Anderson et al. (2012), as analyzed by an expert panel in California (CA). These are useful for a first order approximation of risks (comparison of maximum measured concentration with estimated bioeffects). These benchmark environmental metrics were used as thresholds to predict bio-effects criteria for each chemical contaminant and to indicate when shellfish harvest or contact recreation guidelines were exceeded. Determinations were then made as to the extent and magnitude of expected effects that will occur within the environment for each major class of contaminants for which there are existing environmental quality guidelines.
For microbial pollutants we utilized the South Carolina Estuarine and Coastal Assessment Program’s (SCECAP) water quality index to evaluate the concentrations of bacteria found in stormwater runoff, pond surface samples, and estuarine water samples (Sanger et al. 2016).
Finally, we compared our literature findings with a 2016 NOAA National Status and Trends (herein, “NOAA NST”) Program report. In this report NOAA published data from 161 estuarine sites in coastal S.C. and Georgia (GA). These included evaluations for metals, pesticides, and polycyclic aromatic hydrocarbons. Where applicable we have included a summary table comparing data from the published works described above with those from this NOAA study.
3.3 Chemical Contaminants
3.3.1 Trace Metals
Metals in the environment at trace (or low) concentrations include any metal that can leech from sediments into the environment naturally (e.g., aluminum and iron) or is discharged into the environment from a point source or as an NPS pollutant. Anthropogenic sources include but are not limited to: mining, leechate from solid waste sites, energy production (e.g., coal-fired power plants), emissions from transportation sources (cars, trucks, and trains), animal and human wastes, and paint pigments. The top 20 substances from the federal Comprehensive Environmental Response, Compensation, and Liability Act (CERCLA) Priority List of Hazardous Substances include five metals of concern often found in trace concentrations at Superfund sites across the U.S.: arsenic (As), lead (Pb), mercury (Hg), cadmium (Cd) and chromium (Cr). Trace metals are a major cause of water quality impairments in the U.S., causing more than 11,000 impairments nationwide. Some of the most common metal impairments were: mercury (Hg; 4,476 impairments), iron (Fe; 774), copper (Cu; 620), arsenic (As; 593), lead (Pb; 485), zinc (Zn; 326), aluminum (Al; 297), cadmium (Cd; 241), and nickel (Ni; 61) (EPA 2015). Additionally, in 2010, trace metals were responsible for more than 21% of the 1,103 water quality impairments in S.C. (Mehta 2010). Mercury is one of the major causes of fish advisories in 40 out of 50 states, including S.C. And these trace metals are of environmental and/or public health concerns in shellfish (Scott et al. 1984). While Al and Fe are largely derived from natural sources, other trace metals including Cu, Pb, and Zn are commonly discharged from anthropogenic sources. All of these metals were included in our analysis of pollution in ponds. The elemental symbols are used throughout the following text.
Figure 3.2 Cumulative frequency plot for Cu concentrations in pond sediments and estuarine/tidal creek sites in S.C., where different colored symbols represent different pond types versus tidal creek sites shown in gray. To interpret the figure: recall that for concentrations above the ERL point the %IAE, or estimated potential for toxicity in benthic indicator species, is 29 to 83.7%, while above the ERM the likelihood for negative biological effects is greater than 83.7%. Additionally, the cumulative frequency demonstrates that ~80% of all samples fall below the ERL threshold while < 5% exceeded the ERM. However, we see that more pond samples fall above the ERL relative to the tidal creek samples, indicating that ponds are hot spots for these chemical contaminants. This plot is generally representative of results for Cd and Zn as well, which exceeded SQGs, and concentrations were generally higher in ponds than estuarine sites.
Results of sediment trace metal chemistry analyses (Table 3.2) and environmental risk assessments showed that a number of samples from S.C. exceeded SQGs for eight of the nine trace metals assessed (Table 3.3). Metal concentrations in S.C. ponds can be classified into three groups: 1) metals that do not exceed ERL or ERM, 2) metals that exceeded ERL, but not ERM, and 3) metals that exceeded both ERL and ERM levels. A total of eight ERL exceedances were observed for As, Cd, Cr, Cu, Hg, Ni, Pb, and Zn, with 8.7 to 65% of all samples exceeding the respective ERL. Similarly, three ERM exceedances were observed for Cr, Cu, and Zn, though only 1.1 to 3.7% of all samples exceeded the respective ERM. For three of the eight ERL exceedances (Cd, Cu, and Zn) and two of the three ERM exceedances (Cu and Zn), the highest concentrations were observed in ponds. The second highest concentration for chromium was observed in ponds. These results suggest that ponds serve their design function and remove some trace metals from runoff before it reaches downstream systems.
ERL = n/a
ERM = n/a
ERL = 1.2 mg/kg dw
ERM = 9.6 mg/kg dw
ERL = 81 mg/kg dw
ERM = 370 mg/kg dw
|Reference sites||9324 ± 4604||< 1,181 – 17,491||4||1.32 ± 0.76||< DL – 2.68||4||< DL||< DL||4|
|Residential||8,265 ± 2,746||926 – 25,512||10||2.08 ± 0.35||< DL – 4.14||11||11.1 ± 7.4||< DL – 58||10|
|Golf course||10,549 ± 4,134||1,006 – 26,977||6||2.22 ± 0.38||1.61 – 2.79||6||< DL||< DL||6|
|Mixed land use||10,596 ± 3,170||1,947 – 20,010||6||2.19 ± 0.21||1.73 – 3.12||6||< DL||< DL||6|
|Commercial||7,882 ± 1,707||1,830 – 19,693||10||2.11 ± 0.62||< DL – 2.62||10||< DL||< DL||10|
|Creeks/Estuaries||3,417 ± 679||0.02 – 22,488||114||0.43 ± 0.05||< DL – 5.50||153||98.8 ± 22.6||4 – 2,873||152|
|67.4 ± 3.04||4 – 270||150|
ERL = 34 mg/kg dw
ERM = 270 mg/kg dw
ERL = n/a
ERM = n/a
ERL = 46.7 mg/kg dw
ERM = 218 mg/kg dw
|Reference sites||9.72 ± 5.60||< DL – 20||4||4,471 ± 3,122||1013 – 13,848||4||3.37 ± 0.69||2.17 – 5.17||4|
|Residential||154.54 ± 48.82||< DL – 465||10||4,128 ± 1,396||721 – 15,050||10||1.82 ± 1.71||< DL – 6.48||10|
|Golf course||89.13 ± 25.59||< DL – 161||6||5,031 ± 1,851||682 – 12,061||6||0.91 ± 0.64||< DL – 3.81||6|
|Mixed land use||15.27 ± 4.36||< DL – 32||6||5,831 ± 1,709||520 – 10,833||6||4.05 ± 1.72||< DL – 10.74||6|
|Commercial||181.28 ± 77.30||< DL – 589||10||4,301 ± 1,035||1,118 – 11,549||10||5.00 ± 1.30||< DL – 13.81||10|
|Creeks/Estuaries||22.28 ± 2.06||1 – 271||154||5,203 ± 3,385||0.42 – 320,799||114||29.74 ± 2.02||1.67 – 153.6||148|
ERL = 150 mg/kg dw
ERM = 410 mg/kg dw
|Reference sites||40.52 ± 2.36||< DL – 89||4|
|Residential||37.64 ± 10.89||< DL – 81||10|
|Golf course||30.04 ± 10.94||< DL – 61||6|
|Mixed land use||91.93 ± 37.65||< 2.37 – 244||6|
|Commercial||224.0 ± 58.14||< 6.07 – 573||10|
|Creeks/Estuaries||92.21 ± 5.03||8.60 – 422||148|
Table 3.2 Comparison of concentrations of metal in sediments (units = mg/kg dry wt.) from ponds, tidal creeks, and estuaries in South Carolina. Where # = number of samples; SE = standard error; < DL = below detection limit; Shipyard Creek = Superfund site (placed on the EPA list in 2000); Reference sites = pristine sediments. The ERL and ERM concentrations (in mg/kg dry weight) are provided as references; n/a = no ERL or ERM established for the given metal.
|Trace Metals||% Samples compared to SQG|
|< ERL||> ERL||> ERM|
Table 3.3 Summary of sediment trace metal chemistry results in ponds and tidal creeks compared to SQGs. Note ponds had highest concentrations of cadmium, copper, and zinc that exceeded ERMs in some samples. Gray shading indicates ponds have highest concentration, and bolding indicates ponds have second highest concentrations.
Some trace metals were detected at concentrations that exceeded SQGs in both ponds and tidal/estuarine sites in coastal S.C. There were ERL exceedances for arsenic (65% of all samples), Cd (8.2%), Cr (30.7%), Cu (14.7%), Hg (9.5%), Ni (30.9%), Pb (11.4%), and Zn (12.4%) in all S.C. sites. Pond sediments had the highest levels of Cd, Cu, and Zn as compared to any natural site; concentrations of Zn were especially elevated in sediments from commercial ponds (Weinstein et al. 2008). There were no reported sediment contaminant levels in stormwater ponds for three trace metals: As, Hg, and Ni, indicating a current data gap that future monitoring studies should address. Overall, metals have been correlated with various pond characteristics including pond age, pond surface area, extent of drainage area, and the percent of surrounding impervious cover (Weinstein et al. 2008). With the exception of cadmium, metal concentrations in ponds were positively correlated with percent clay content of sediments and total organic carbon (TOC). In general, metals associate with fine-grain sediments (e.g., clay). The accumulation of copper in pond sediments has been previously linked to known applications of Cu-based algaecides, especially in golf course and residential ponds (Weinstein et al. 2008). Also, it was found that for several metals, concentrations were higher at the pond inlet as compared to the center, while other metals were concentrated at the pond center (Weinstein et al. 2008), indicating sedimentation does not necessarily occur uniformly throughout the basin.
The levels of trace metals in sediments may pose significant ecotoxicological risk to benthic organisms in ponds, with the predicted %IAE for trace metals ranging from 6 to 95, including for Cd (6 to 36%), Cr (21 to 95%), Cu (29 to 84%), and Zn (47 to 70%). These risks would be primarily concentrated within the endemic benthic fauna within ponds; though many of the ponds in coastal S.C. do have tidal exchange with nearby creeks. Additional impacts of flow from these ponds to estuaries and tidal creeks may occur when there is significant discharge from the ponds, which would involve increased suspended sediment loads. This occurs during heavy precipitation events or tropical disturbances. It is worth noting that some of the ponds in S.C. (those with higher salinity) are natural habitats for benthic/epibenthic organisms such as oysters, crabs, and shrimp. Residents may fish these organisms for food, resulting in human exposure which may be a potential public health concern. We compared exceedances of SQGs trace metals summarized in this report versus levels reported by NOAA NST at 161 estuarine sites randomly sampled and analyzed in S.C. and GA. This comparison indicated very similar results for some trace metals (Hg and Ni) between the states, while S.C. had much higher levels of exceedance for As, Cd, Cr, Cu, Pb, and Zn (Table 3.4).
|Trace Metals||NOAA NST||S.C.||NOAA NST||S.C.|
|> ERL||> ERL||> ERM||> ERM|
Table 3.4 Comparison of trace metal ERL/ERM exceedances of SQGs for estuarine/tidal creeks and pond sites in S.C. (this study) versus estuarine sites in the southeastern (SE) U.S. (S.C. and GA). Gray shading indicates ponds have highest concentration, and bold text indicates ponds have second highest concentrations (out of 186 samples).
Most measurements in ponds have been of total metal concentrations with little attention given to metal speciation. Metals can occur as free ions, complex with ligands such as organic matter, and anions sorbed to nanoparticles, colloids, and particulate matter, or occur as nanomaterials (more details in sub-section 3.5.3). The environmental and health risk of metals depend on their speciation, with free/labile metals generally being the most toxic species. Therefore, future research should consider:
- Speciation of metals in pond ecosystems and the impact of metal speciation on their environmental behavior, fate, sequestration in sediments and potential for later release to surface waters, and cumulative mixture toxicity.
- Processes should be studied in the context of land use and the variability of storm events.
3.3.2 Pesticides and Polychlorinated Biphenyls (PCBs)
Worldwide, more than 1,000 organic chemicals are registered for use as active ingredients in pesticides, which are often ubiquitously present in the environment (Aprea et al. 2002). Approximately one billion pounds of active ingredient pesticides are used in the U.S. each year (Nowell et al. 1999). Because of this widespread use, there are several monitoring programs put in place by governmental agencies, including the Food and Drug Administration (FDA), the EPA, and the Department of the Interior.
Inputs are derived from residential coastal development, such as termiticide treatment of dwellings, residential lawn and turf management, golf courses, or antifouling agents, as well as industrial, municipal, and groundwater inflows. In fact, one third of the pesticide used in the U.S. is attributed to urban use, and this is reflected by concentrations of elevated organochlorine pesticides in urban influenced streams, which frequently exceed samples from intensive agricultural sites (Paul & Meyer 2001). Urban areas were found by the EPA to have the greatest exceedance of chronic Aquatic Life Benchmarks (ALBs) for Pesticide Registration based on toxicity levels for various taxa. The ALBs are national guidelines of toxicity values for pesticides based on scientific studies for freshwater species. From 2002–2011 throughout the U.S. nearly two-thirds of agriculture land-use streams, 50% of mixed land-use, and 90% of urban streams exceeded ALBs (Figure 3.3). Primary pesticides found to exceed ALBs were fipronil, metolachlor, malathion, cis-permethrin, and dichlorvos, which exceeded chronic ALBs in more than 10% of streams sampled. Impervious surfaces in developed areas exacerbate the issue by increasing the direct loading of pesticides into waterways (Paul & Meyer 2001).
Figure 3.3 Results of national stream and river pesticide assessment reported by USGS (Stone et al. 2014). This study includes pesticide monitoring data from sediments, grain size, and total organic carbon measurements.
Federal monitoring programs are supplemented by regional and local studies, such as USES, LUCES, and SCECAP. Wickliffe (2013) reported that more than 80 of the 100 pesticides most frequently sold by Lowes and other retail outlets in S.C. are used on residential lawns and golf courses. This application was more than four times greater than the number of pesticides used for other coastal pesticide applications. These data are significant given the high levels of urbanization in coastal S.C., and underscore the importance of ponds to control NPS runoff that may contain pesticides used in urban areas.
Pesticides are manufactured so that they are toxic to a general pest, such as weeds, insects, fungi, and rodents. If not properly applied according to the label, pesticides may then affect non-target species, including aquatic life, wildlife, and humans, because most pesticides are not necessarily species-specific but are designed to be toxic to a broad array of organisms (Aprea et al. 2002). Pesticides have been found in the tissues of a wide variety of aquatic species, including oysters, shark, finfish, tuna, sea turtles, alligators, sea birds, and clams (O’Connor 1991; Mathews 1994; Lauenstein et al. 2002; Gelsleichter et al. 2005). Coastal wetland ecosystems are of particular interest for any coastal application of pesticides, but treatments upland will eventually be transported to these systems via runoff and stream flow (Clark et al. 1993; EPA 2000). Wetlands and their sediments act as a repository for pesticides. The major concern for adverse effects of pesticides in the environment is based on the transport of these compounds by NPS runoff following precipitation events. Spray drift from local mosquito control applications may also be a source of pesticide exposure. Several studies have linked land use and the occurrence of pesticides in associated runoff and receiving waters in S.C. (Scott et al. 1994; Corbett et al. 1997; Sanger et al. 1999a, b; Graves et al. 2004; Weinstein et al. 2008; Crawford et al. 2010). Studies have also demonstrated chronic and acute toxicity for pesticides in low concentrations (from ng/L to mg/L) depending upon the pesticide class and the species exposed (Clark et al. 1993).
There were no published literature sources for PCBs in S.C. ponds at time of this report. Though there were no PCBs measured in pond sediments, 11.3% of the estuarine and tidal creek sites in this study exceeded ERLs. Similarly, NOAA NST reported that there were ERL exceedances for PCBs at 9.9% of the 161 estuarine sites surveyed in S.C. and GA, similar to the levels reported in S.C. tidal creeks (Sanger et al. 1999b).
Legacy pollutants refer to chemicals produced by industry that remain in the environment long after they are introduced. Chlordane was phased out by EPA in 1989 as a termiticide, the year Hurricane Hugo inundated much of the coast. This termiticide was detected in 11.1% of ponds, with concentrations ranging from DL (detection level) to 1.5 ng/g (average = 0.16 ng/g). Chlordane was only detected in residential ponds (range = DL – 15; average = 0.3 ng/g) and commercial ponds (range = DL – 1.3 ng/g; average = 0.22 ng/g). Sanger et al. (1999b) reported much higher concentrations of chlordane in S.C. tidal creeks, with highest concentrations in mixed land use residential areas (mean concentration ≤ 4.5 ng/g), and lower concentrations in agricultural, residential (urban lower density), and industrial areas (mean concentrations ≤1 ng/g). Concentrations in more pristine reference sites were much lower (< 0.75 ng/g).
Contemporary-Use Pesticides (CUPs)
Chlorpyrifos was the most frequently detected pesticide in S.C. pond sediments; nine out of 16 ponds (or 56% of total ponds surveyed) had detectable concentrations ranging up to 27.7 ng/g for a single high-density residential pond. Concentrations of chlorpyrifos in pond sediments, sampled from the center of ponds, were not significantly different among pond types (golf course = 9.46 ± 2.75 ng/g; low-density residential = 8.17 ± 2.01 ng/g; commercial = 10.97 ± 5.45 ng/g; Figure 3.4a). Chlorpyrifos sediment concentrations were found to be dependent on pond surface area (Weinstein et al. 2008). Holland et al. (2008) found lower concentrations of chlorpyrifos in S.C. tidal creeks (Figure 3.4b). The lowest levels were found in sediments from reference creeks (0.2 ng/g), while first order urban tidal creeks had the highest concentrations, as they were most closely associated with development (0.8 ng/g). DeLorenzo and Fulton (2009) found similar trends for the herbicide atrazine in estuarine surface water and ponds; maximum annual atrazine levels in ponds (1.00 μg/L) were nearly an order of magnitude higher than in tidal creeks (less than 0.05 μg/L). Weinstein et al. (2008) found several other pesticides in S.C. ponds including: endosulfan = 5.6% of ponds (6.2 ng/g in commercial ponds); fonofos = 5.6% of ponds (10.0 ng/g in low density residential ponds); and dichlorvos = 5.6% of ponds (25.9 ng/g in high population density mixed land use) ponds (Table 3.5).
Figures 3.4a and 3.4b A comparison of average sediment concentrations (± one standard deviation when replicates available) of the organophosphate insecticide chlorpyrifos in a) S.C. ponds (Weinstein et al. 2008) and b) S.C. tidal creeks (Holland et al. 2008). Ponds were distinguished by land use classes into: golf course, low density residential (LD Res), high density residential (HD Res), and commercial (Comm) ponds. First order creeks are those in closest proximity to the terrestrial or built environment.
|Site Description||Chlordane (ng/g)||Endosulfan (ng/g)|
|Mean ± SE||Range||#||Mean ± SE||Range||#|
|Residential||0.3 ± 0.3||< DL – 1.5||5||ND||ND||5|
|Mixed land use||ND||ND||2||ND||ND||2|
|Commercial||0.22 ± 0.22||< D L – 1.3||5||1.03 ± 0.22||< DL – 6.2||6|
|Creeks/Estuaries||< 4.5a||< DL – 4.5||NR||NC||< DL – 0.28b||NR|
|Fonofos (ng/g)||Dichlorvos (ng/g)|
|Mean ± SE||Range||#||Mean ± SE||Range||#|
|Residential||2.00 ± 2.00||< DL – 10||5||ND||ND||5|
|Mixed land use||ND||ND||2||12.95||< DL – 25.9||2|
Table 3.5 Sediment concentrations of pesticides as reported by where # = number of samples; SE = standard error; < DL = below detection limit; ND = not detected; NA = not assessed; NC = not calculated. Data taken from: a = Holland et al. (2008); b = Leight et al. (2005) and all other from Weinstein at al. (2008).
The presence of legacy contaminants in pond sediments indicates that ponds do sequester these chemicals. Pesticides can have a long half-life when particle bound, and the anaerobic conditions in pond sediments may increase the life of these chemicals, leading to lower degradation rates (Bondarenko & Gan 2004). NOAA NST reported that there were ERL exceedances for DDT, another legacy pollutant, at 19.8% and for DDE, a breakdown product of DDT, at less than 1% of the 161 estuarine sites surveyed in S.C. and GA (Table 3.6). However, despite the persistence of these legacy pollutants, concentrations measured in ponds are lower when compared with levels of contemporaryuse pesticides such as chlorpyrifos. This chemical replaced chlordane as a termiticide in residential areas until this usage was phased out by the EPA in 2001 due to risks to human health; this chemical is still utilized in commercial applications.
Chlorpyrifos has caused pervasive ecotoxicology issues in aquatic ecosystems. Stone et al. (2014) found that the organophosphate insecticide chlorpyrifos was one of several insecticides found in extremely high concentrations in urban areas and was one of the major causes of exceedances of ALBs nationally. There are no SQGs established by the EPA for this compound. However, Weinstein et al. (2008) found that sediments from nine ponds (50%) had levels of chlorpyrifos which exceeded ecological life criteria screening level benchmarks (derived by EPA 1996 and Jones et al. 1997), but none of the ponds had levels that were above the benchmarks for human health concern. Chlorpyrifos was identified by the state of CA Expert Panel on CECs (Anderson et al. 2012) as one of 12 Dirty Dozen Contaminants of Concern recommended for future monitoring and was reported by USGS (Stone et al. 2014) as one of several problematic insecticides, particularly in urban areas. Finally, this pesticide is known to be synergistically toxic to aquatic invertebrates when found in the presence of the herbicide atrazine. The aquatic life criteria is 5.6 ng/L for the combination of these pesticides.
|> ERL||> ERL|
Table 3.6 Comparison of DDT and PCBs to the SQG assessments of estuarine sites and ponds in S.C. (this study) versus estuarine sites reported by NOAA NST.
Only limited data exists on pesticide levels in ponds, and PCB data is largely lacking. In particular, Contemporary Use Pesticides such as chlorpyrifos and pyrethroids should be monitored given their likelihood for acute and chronic toxicity to invertebrates (Maruya et al. 2013). Continued monitoring of these pesticides is warranted to evaluate the future hazards these pesticides pose to aquatic organisms. Specifically:
- Levels of chlorpyrifos, bifenthrin, fipronil, and pyethroids such as permethrin should be monitored in ponds as recommended by the State of CA Expert Panel on CECs, based on their environmental risk to aquatic organisms relative to measured concentrations.
- Mosquito control pesticides should be further monitored, given the emphasis on prevention and control of West Nile and the Zika virus, as most of the recommended insecticides used for mosquito control have very narrow margins of safety.
- Chlorpyrifos and atrazine should continue to be monitored due to their synergistic toxic effects on crustaceans and their widespread occurrences in ponds.
3.3.3 Polycyclic Aromatic Hydrocarbons (PAHs)
Polycyclic aromatic hydrocarbons are a ubiquitous group of chemical compounds that are predominately formed during the incomplete combustion of organic materials. PAHs may be formed three ways in the natural environment: high temperature pyrolysis of organic materials; low to moderate temperature diagenesis of sedimentary organic material to form fossil fuels, and direct biosynthesis by plants and microbes (Neff 1979). Fires occurring in nature are the largest contributing factor for natural sources of PAHs released in the atmosphere (Eisler 1987). Although PAHs can occur naturally in the environment, they are also considered an important group of pollutants associated with development and found in urban runoff (Sanger et al. 1999b; Walker et al. 1999; Van Dolah et al. 2005, 2006). PAHs are defined by their fused ring structure of two or more benzene rings. Those with two to five benzene rings are, in general, of the greater concern for human and environmental health (EPA 1999). These PAHs differ in their distribution and behavior in the environment and can be divided into low molecular weight PAHs and high molecular weight PAHs. Solubility tends to decrease with increasing molecular weight. Low molecular weight PAHs (two to three rings and include fluorene and anthracene) are more likely to have significant acute toxicity on aquatic organisms. High molecular weight PAHs (four to seven rings) are known to be carcinogenic. The EPA identifies 34 PAH compounds as being the most abundant and frequently measured in sediments (EPA 2003) and of these, 16 are on an EPA priority pollutant list due to the fact these compounds are characterized as mutagenic or carcinogenic (EPA 2002a, b).
Figure 3.5 PAH concentrations measured in marsh sites adjoining major highway types in coastal areas of S.C. (taken from Van Dolah et al. 2005). The line indicates the ERL concentration. Note that highest PAH levels were found on moderate-use roads and not more heavily traveled roads due to road design and the distance to the adjoining marsh from the roadway. However, in all cases, concentrations were below the ERL.
PAHs released in fires tend to adhere to suspended particulates and enter the terrestrial and aquatic environments when the particulates fall out (Eisler 1987). Based on their application and routes of disposal, PAHs will eventually be discharged into aquatic ecosystems. PAH concentrations are often elevated around roadways due to the significant discharge of auto emissions. Van Dolah et al. (2005) assessed these for S.C. coastal highways (Figure 3.5) and found the highest PAH levels originated from moderate-use roads. Because practically all incineration processes generate PAHs, several studies from S.C. compared PAH concentrations across a range of watershed-level land uses (Sanger et al. 1999b; Menzie et al. 2002; Brown & Peake 2006).
Results of sediment PAH risk assessments indicated that SQGs were exceeded for 12 of the 13 PAHs evaluated.
Figure 3.6 Fluoranthene concentrations (μg/kg dw) measured in sediments from ponds, tidal creeks, and other coastal estuarine locations in S.C. Note how pond concentrations exceeded the ERLs and ERMs and were generally higher than estuarine tidal creeks. Similar results were seen for phenanthrene and pyrene.
Therefore, PAHs can be classified as those that exceeded ERLs or those that exceeded both ERLs and ERMs (e.g., Figure 3.6). Of the 12 ERL exceedances, 1.5 to 65% of all samples exceeded the respective ERL. Similarly, a total of eight ERM exceedances were observed for anthracene, fluorene, phenathrene, benzo(a)anthracene, benzo(a)pyrene, chrysene, fluoranthene, and pyrene, with 2.3 to 16.4% of samples exceeding their respective ERM (Table 3.8). For three of the 13 ERL and ERM exceedances (phenanthrene, fluoranthene, and pyrene) highest concentrations were measured in ponds. These results indicate that ponds may act as potential sinks for at least some PAHs. All four urban pond types: commercial, golf course, mixed land use, and residential had the highest concentrations measured for most contaminants as compared to ponds from reference (pristine) areas or with tidal creeks and estuarine ecosystems (Figure 3.6; Tables 3.7 and 3.8).
|Mean ± SE||Range||#||Mean ± SE||Range||#||Mean ± SE||Range||#|
|Reference Sites||91.5 + 19.8||31.6 – 161.6||9||57.9 ± 19.1||< DL – 125.8||9||364.1 ± 153.7||< DL – 1178.4||9|
|Residential||91.4 ± 10.8||50.3 – 197.7||21||15.5 ± 6.1||< DL – 83.9||21||30.6 ± 9.7||< DL – 162.0||21|
|Golf Course||191.1 ± 46.4||62.2 – 588.3||13||30.5 ± 10.0||< DL – 130.3||13||9.4 ± 5.4||< DL – 67.0||13|
|Mixed Land Use||132.0 ± 13.3||55.1 – 197.6||13||95.1 ± 48.5||< DL – 491.6||13||13 48.3 ± 26.8||< DL – 284.1||13|
|Commercial||424.8 ± 126.9||54.2 – 1993.8||21||318.7 ± 97.0||0 – 1491.4||21||21.3 ± 10.0||< DL – 147.0||21|
|Creeks/Estuaries||16.6 ± 5.7||3.0 – 273.0||55||42.6 ± 17.2||2.0 – 705.0||54||38.7 ± 32.1||1.00 – 1783.0||55|
|Mean ± SE||Range||#||Mean ± SE||Range||#||Mean ± SE||Range||#|
|Reference Sites||40.0 ± 4.8||27.8 – 64.8||9||1.9 ± 1.9||< DL – 17.0||9||290.8 ± 94.8||11.6 – 658.3||9|
|Residential||42.7 ± 4.6||30.7 – 110.8||21||44.7 ± 16.3||< DL – 225.3||21||282.6 ± 13.0||12.0 – 1591.2||21|
|Golf Course||81.5 ± 16.1||30.3 – 178.4||13||72.6 ± 16.5||4.8 – 193.3||13||528.2 ± 147.3||35.6 – 1612.4||13|
|Mixed Land Use||49.1 ± 5.9||28.8 – 103.1||13||79.2 ± 24.1||13.6 – 241.3||13||513.7 ± 210.4||39.8 – 2306.8||13|
|Commercial||247.1 ± 88.3||DL – 1432.1||21||11.3 ± 26.6||DL – 353.5||21||3701.2 ± 1111.0||19.9 – 16,345.1||21|
|Creeks/Estuaries||16.7 ± 7.6||1.0 – 386.0||54||31.5 ± 5.7||5.0 – 227.0||57||97.1 ± 47.8||2.0 – 2534.0||54|
|Mean ± SE||Range||#||Mean ± SE||Range||#||Mean ± SE||Range||#|
|Reference Sites||18.7 ± 6.8||< DL – 50.7||9||6.6 ± 2.6||< DL – 20.9||9||1.8 ± 1.2||< DL – 9.5||9|
|Residential||207.8 ± 98.8||3,160 – 1323.9||21||500.8 ± 178.0||5.3 – 2253.2||21||17.0 ± 7.1||< DL – 94.8||21|
|Golf Course||104.2 ± 45.5||3.3 – 555.0||13||532.3 ± 176.0||8.5 – 1972.9||13||9.3 ± 4.1||< DL – 51.3||13|
|Mixed Land Use||682.9 ± 3112.8||5.2 – 3013.8||13||728.2 ± 239.2||14.2 – 2444.1||13||76.0 ± 39.8||< DL – 421.3||13|
|Commercial||1888.4 ± 515.2||< DL – 7126.5||21||2,335.5 ± 525.7||< DL – 7393.6||21||209.9 ± 60.3||< DL – 954.3||21|
|Creeks/Estuaries||310.0 ± 128.2||5.0 – 7328.0||63||172.6 ± 64.5||1.0 – 3171.0||54||28.9 ± 14.9||1.0 – 818.0||59|
|Mean ± SE||Range||#||Mean ± SE||Range||#|
|Reference Sites||53.0 ± 17.3||2.7 – 118.4||n/a||192.7 ± 66.4||< DL – 448.4||9|
|Residential||156.8 ± 492.3||11.2 – 6600.5||n/a||797.5 ± 386.7||< DL – 5140.3||21|
|Golf Course||445.4 ± 193.6||17.2 – 2329.5||n/a||385.7 ± 175.3||< DL – 2110.7||13|
|Mixed Land Use||2214.3 ± 997.4||5.4 – 9359.2||n/a||1871.6 ± 869.5||2.33 – 8311.0||13|
|Commercial||9652.9 ± 2687.5||9.8 –|
|< DL –|
|n/a||792.5 ± 451.5||2.00 –|
Table 3.7 Concentrations of polycyclic aromatic hydrocarbons in sediments (μg/kg dry wt.) from ponds; creeks/ estuaries = all tidal creeks and estuaries in S.C. where # = number of samples; SE = standard error; < DL = below detection limit; Reference Sites = pristine sediments.
|% Samples compared to SQG|
|Contaminant||< ERL||> ERL||> ERM|
Table 3.8 Summary of sediment PAH chemistry results in ponds and tidal creeks compared to SQGs. All ponds and many tidal creeks have contaminant concentrations > ERL. Gray shading indicates ponds have highest concentrations.
Many PAHs in sediments were detected at levels that exceeded SQGs in both ponds and tidal creek/estuarine sites within coastal S.C. Ponds had the highest levels of several PAHs including phenanthrene, fluoranthene, and pyrene. Overall, PAHs were not highly correlated with clay or TOC content of the sediments, suggesting a different deposition pathway from metals, likely long-term atmospheric deposition (De Luca et al. 2005). Though PAHs from some ponds carried a distinct fuel signature (based on ratio of low to high molecular weight), the primary source for PAHs in all 16 ponds were pyrogenic sources (Weinstein et al. 2008). Commercial ponds had the highest PAH concentrations, and these were especially elevated in older ponds that had likely not been dredged (Fernandez & Hutchinson 1993). Residential and golf course pond sediments carried the highest motor oil signatures. Sediment-bound PAHs may pose significant ecotoxicological risk to benthic organisms within ponds, as the predicted %IAE for PAHs range from 10.3 – 73. The high %IAE values for all PAHs tested for suggest that at the reported concentrations there is potential for adverse effects in benthic organisms within coastal S.C. ponds.
Comparison of PAH exceedances summarized in this report versus levels reported by NOAA NST indicated very similar results for some PAHs (2-methylnapthalene, napthalene, benzo(a)anthracene, and dibenzo(a,h) anthracene), while S.C. had higher levels of acenapthene, fluorene, phenanthrene, benzo(a)pyrene, chrysene, fluoranthene, and pyrene (Table 3.9). NOAA reported that the mean estimated impact by PAHs in S.C. and GA based upon %IAE was 14.6 (at the 95% confidence level (CL) = 10.0 – 19.3%) while experimental sediment toxicity tests (based upon amphipod development and reproduction in sea urchins) provided estimates of 0.3 to 21.3%. Thus, the ranges of predicted and actual toxic effects overlapped between the two approaches. Comparisons of the potential for impacts predicted in S.C. similarly provide areal estimates of 5.8% (at the 95% CL = 0 – 18%) of tidal creek/estuarine sites and 10.5% (95% CL = 3 – 17%) of pond samples, which have the highest concentrations of legacy pollutants. These predictions of ecotoxicological effects from legacy pollution in our S.C. study compares very favorably with NOAA results for the following watersheds: Charleston, Winyah Bay, and North Edisto River (in S.C.), and the Savannah and St. Simons Island (in GA).
|NOAA NST||S.C.||NOAA NST||S.C.|
|Contaminant||> ERL||> ERL||> ERM||> ERM|
Table 3.9 Comparison of PAH SQG Assessments of estuarine and pond sites in S.C. (this study) versus estuarine sites reported by NOAA NST. Gray shading = ponds had the highest concentrations. Note the higher PAH levels in S.C. when compared to other parts of the southeastern U.S.
Because of the potential toxic effects for aquatic organisms of the various contaminants sequestered in ponds, whether these are potentially hazardous to wildlife needs to be better constrained. Specifically, we need to better understand:
- Ecotoxicological effects of PAHs discharged from ponds into estuarine ecosystems and the spatial extent of any downstream impacts.
- Cumulative mixture ecotoxicological effects of different PAHs and other contaminants.
3.4 Microbial Pollutants
Microbial pollution in coastal surface waters is a problem affecting recreational and commercial uses of rivers, beaches, and estuaries throughout the U.S. and is one of the major causes of water quality impairments in S.C. As development alters landscapes and watershed hydrology, the major sources of bacterial pollution are human waste from inadequately maintained septic tanks and waste from wildlife, livestock, and pets. The standard solutions to this problem are to construct a central sewer collection system to reduce estuarine inputs from individual septic tank systems (Jolley 1978) and to construct stormwater BMPs to manage bacteria in NPS runoff.
In 2012, commercial fisherman in S.C. collected 9.9 million pounds of shellfish valued at $17.5 million in landing revenue (NMFS 2014). Microbial contamination can result in the closure of shellfish harvesting waters due to the presence of pathogenic bacteria and viruses (Scott et al. 2006). Bed closures can also decrease property values near the affected tidal creeks. Roughly 30 million tourists visit the state each year, and the majority of these visitors come to the coast. Any beach closures can have an economic impact on coastal S.C. It was estimated that in 2016, two million fewer tourists visited Myrtle Beach than in the previous year, and one of the reasons cited was the perception of bacteria-laden waters (Myrtle Beach Chamber of Commerce). Therefore, there is a strong need for tools that can help resource managers make timely and accurate decisions regarding closures, from both a public health and economic standpoint.
|Limits in CFUs/100 mL1|
|Monthly Average||Daily Maximum||Type||Indicator Used||Class||Approved Use|
|35||104||saltwater||Enterococci||SA2||recreational; heavy use|
|35||501||saltwater||Enterococci||SB3||recreational; infrequent use|
1CFU = colony forming units; 2Salt Water A rating; 3Salt Water B rating.
Table 3.10 Current bacteria classifications and standards for waters of the State taken from DHEC’s Bureau of Water R.61-68 Water Classifications and Standards (effective June 27, 2014). Waters of the State as defined by the S.C. Pollution Control Act, and in compliance with the federal Clean Water Act, does not currently include stormwater conveyances, although there was some traction starting in 2014 at the federal level to include wetlands and other tributaries (such as wet ponds) as waters of the U.S. (see the Clean Water Rule in Chapter 5).
FCB concentrations are used to evaluate long-term water quality status of shellfish harvesting waters, shellfish tissues, and waters used for recreational contact. The coliform group is broad and consists of all aerobic and anaerobic, gram negative, non-spore forming, rod-shaped bacteria that grow and produce gas in broth when incubated at 35 – 37 °C (mammalian body temperature) within 24 hours. In general, FCB are associated with contamination from warmblooded animals, and the FDA approves the measure of their abundances as an indicator of fecal pollution to protect shellfish consumption and recreational contact (Cabelli et al. 1983; Ruple et al. 1989; FDA 1989). FCB counts can indicate Escherichia coli, which is one of the more commonly measured bacteria in human feces (Kator & Rhodes 1996) and occurs with greater frequency in urban areas versus more pristine sites (Vernberg et al. 1996). In shellfish tissues, E. coli is the principal FCB present when water temperatures are less than 22°C (Cook & Ruple 1989), while at higher temperatures, Klebsiella species will dominate over other FCB. Vernberg et al. (1996) reported that K. pneumoniae was the most dominant member of the FCB group after E. coli in the evaluation of estuarine surface waters and oysters from pristine and urban estuaries in S.C. Klebsiella species are usually not a threat to human health (Paille et al. 1987). Enterococcus and E. coli have been shown recently to more accurately predict acute gastrointestinal illness from contact by swimming than total levels of FCB. Therefore, DHEC now uses Enterococcus rather than FCB as the indicator for contact recreation in saltwater, while E. coli is the indicator used for recreational contact in freshwater. Note that limits based on Enterococcus concentrations are lower than those for E. coli or total FCB as this genus represents just 1% of the flora in saline systems (Layton et al. 2010).
Several studies of microbial pollution in S.C. ponds were identified, including the state of knowledge report on 112 ponds (Drescher et al. 2007). To recap, in this comprehensive review, most ponds were described as being freshwater with salinities less than 1 ppt. Approximately 20% had low dissolved oxygen concentrations (less than 4 mg/L), and some 23% of the ponds also had high FCB levels, greater than 400 colony-forming units (CFUs)/100 mL in surface water. The highest FCB levels were found after rain events.
|Fecal Coliform Bacterial Density (cfus/100 ml)|
|Source Water||<43||>43 – <200||>200 – <400||>400|
|Estuarine Surface Water||34%||50%||13%||13%|
Table 3.11 Comparison of fecal coliform levels (% of samples exceeding specific criteria) in NPS runoff in pond surface waters and in estuarine surface waters (after Scott et al., 2008). Geometric mean MPN values of <14-43 colony forming units (cfus)/100 ml are used as the standards for safe shellfish harvesting. MPN values of >43 cfus/100 ml are used as an indicator of shellfish harvest standard exceedances. Fecal coliform standards of 200cfus/100ml have been used as Class SA Water standards and were formerly the safe contact recreation standard. EPA Whole Body Water Contact Standards in marine waters today use an Enterococci level = 104cfus/100ml (single sample max). Fecal coliform standards of 400cfus/100ml have been used as Class SB Water Quality standards.
It is important to compare FCB levels in ponds to levels measured in adjoining estuarine ecosystems, as ponds are intended to intercept and remove pollutants. Scott et al. (2008) summarized results from two studies: DHEC-OCRM (2002) and the USES study. A comparison of FCB concentrations from NPS runoff, ponds, and estuarine surface waters indicated that the overall densities differ (USES 2004; Table 3.11), with FCB levels in runoff far exceeding that of ponds and estuarine waters. The USES data indicated that FCB densities in runoff source water from urbanized watershed in Murrells Inlet, S.C. ranged from 430 – 11,000 CFUs/100 mL (average = 4,610 CFUs/100 mL) versus densities ranging from 4,600 -11,000 CFUs/100 mL (average = 8,867 CFUs/100 mL) at the pristine NOAA National Estuarine Research Reserve (NERR) site in North Inlet, S.C. (Kelsey et al. 2003). Peak FCB densities of approximately 11,000 CFUs/100 mL measured in runoff were similar to peak levels at the NERR site and Murrells Inlet (USES 2004). These data far exceeded thresholds for recreational contact and shellfish harvesting standards (see Table 3.10 above for current water quality classifications). While overall there were no statistical differences detected between the two sites when data were pooled, monthly comparisons showed FCB levels were higher in the NERR system than Murrells Inlet (USES 2004). Similarly, the DHEC-OCRM study (2002) reported peak FCB densities as high as 35,000 CFUs/100 mL in NPS runoff flowing into estuarine areas of the East Cooper River near the Isle of Palms in Charleston County, S.C. Densities exceeding 10,000 CFUs/100 mL were measured in surface waters at eight other stream locations that receive runoff in the East Cooper river region, including Shem Creek.
Figure 3.7 Data from coastal S.C. ponds and natural sites (Scott et al. 2008) and evaluated based on the S.C. Estuarine and Coastal Assessment Program (SCECAP; Sanger et al. 2016) criteria of poor (>400 CFU/mL), fair (43 < x ≤ 400 CFU/mL), and good (≤ 43 CFU/mL) water quality.
Comparison of FCB levels in surface waters of ponds and estuarine samples were more comparable and generally indicate the importance of reducing levels in NPS runoff through sedimentation and other processes. Scott et al. (2008) analyzed FCB data from ponds in S.C. and compared these data with levels in estuarine surface waters and tidal creeks in S.C. These were plotted based on SCECAP’s water quality criteria (Figure 3.7; Sanger et al. 2016). Surface waters containing NPS runoff that flows into ponds generally contain very high levels of FCB (often >1000 CFUs/100 mL), which exceed shellfish harvest standards and standards for contact recreation in more than 99% of the samples. FCB levels were higher in residential and commercial ponds than in industrial and golf course ponds. In a comparison of a freshwater pond in the Charleston area with the receiving tidal creek, Serrano and DeLorenzo (2008) showed the tidal creek often had elevated FCB levels. The average creek levels were 1519 CFU/100 mL as compared to 181 CFU/100 mL in pond surface waters. Pond concentrations were often below the threshold for recreational contact while creek values far exceeded those allowed for contact and shellfish harvesting. Levels of FCB in waters decreased with decreasing water temperatures, with the exception of trends driven by introduction of bacteria from migratory birds in fall and winter. More recent S.C. pond studies found that FCB levels in surface waters of seven ponds of differing land use ranged from 1 to 2,600 CFUs/100 mL (average = 505.5 ± 361.5 CFUs/100 mL; Johnston et al. 2009), and in two freshwater ponds from 86.2 to 469.2 CFUs/100 mL (average = 238.5 ± 41.3 CFUs/100 mL; DeLorenzo et al. 2012).
More than 99% of FCB loadings to coastal waters occur during storm events due to direct deposition by runoff and sediment resuspension. During storms, FCB levels in downstream systems are most affected by erosion of sediments in pipes and channels that convey stormwater. Collectively, studies show the wide range of FCB measured in surface waters; some of this variance is due to differing rainfall amounts that drive increased bacterial loadings. In general, NPS runoff levels are higher than estuarine surface waters, and the mean estuarine surface water FCB levels were typically reduced by 10 – 100 fold. Levels in the middle and outer regions of most estuaries had much lower levels of FCB (1-100 CFU/100 mL range) due to tidal mixing and dilution from offshore seawater that is tidally exchanged.
Berquist et al. (2010) demonstrated the negative correlation between FCB levels and salinity in ponds and tidal creeks. Holland et al. (2004) found significant correlations between FCB levels in surface waters of estuarine tidal creeks and the extent of impervious surfaces and land use. A mechanistic model linked to flow (Holland et al. 2004) showed that when impervious cover exceeded 10 to 20% of a watershed we see increased FCB loadings to waterways. Peak flows in rain events (Cubic Feet per Second = cfs) and impervious surfaces were highly correlated with highly urbanized areas; where 40% cover had 271 cfs versus 175 cfs in moderately urbanized areas (defined as greater than 10% and less than
40% impervious area). In areas with low urbanization (less than 10% cover), peak flows were 100 cfs so the ratio of peak flows in high to low urbanized areas was 2.71. This value is very similar to enrichment levels measured by Van Dolah et al. (2008): 2.85 for chemical contaminants and 3.64 for FCB in highly urbanized areas. This underscores the importance of using BMPs to reduce chemical and microbial pollution in urban runoff and to reduce bacterial loadings to estuarine receiving waters used for shellfish and contact recreation. However, the North Inlet vs. Murrells Inlet comparison study and the results from Serrano & DeLorenzo (2008) demonstrate the potential for seasonal contributions from migratory waterfowl and wildlife at estuarine sites (Siewicki et al. 2007). These high FCB levels at even pristine sites demonstrate the need for management strategies to reduce bacterial source loadings with Poop-a-Scoop programs (e.g., pet waste reductions) and effective wildlife management strategies (e.g., Canada geese deterrents).
Overall, the higher bacterial concentrations found in NPS runoff near pond inflow sites compared to near pond outlets indicates that ponds do reduce FCB levels. Chapter 4 discusses some biogeochemical processes that occur in ponds which may remove some pollutants. In the 2007 state of knowledge report, Drescher et al. reported that FCB reductions in 112 coastal ponds ranged from -47 to 99% and Total Suspended Solids (TSS) were reduced from 79 to 91%. This close agreement between FCB and TSS reductions reinforces the importance of bacterial association with sediments. The EPA finds that 90% of materials settle out of the water column in ponds between rain events. Though the primary means of pollutant removal in ponds is via sedimentation, modeling indicates that other processes must also be involved (Borden et al. 1998). Multiple ponds connected in a series provided greater treatment potential than single pond designs as the cumulative sedimentation led to greater overall removal efficiencies (Messersmith 2007).
Weinstein et al. (2008) demonstrated that bacterial levels in ponds were positively correlated with the size of the pond’s drainage area, pond surface area, concentrations of total organic carbon, and percent clay particles. For ponds with low turbidity, significant UV treatment from solar radiation may also occur in the surface waters, which may further reduce FCB levels.
Overall these studies suggest that different mechanisms control and predict FCB densities during dry weather versus storm periods and that the geometry of the stormwater conveyance system is an important control on sediment resuspension during storm events. Therefore, pond design should include features that prevent the disturbance of sediments during rain events. Otherwise ponds can discharge both chemical and microbial pollutants and become significant sources downstream.
In general, risks from human sources are considered to be more dangerous than non-human sources due to greater chance for antibacterial resistance (DHEC-OCRM 2002; USES 2004). In the majority of studies, bacteria tested from ponds had low levels of antibiotic resistance, indicating wildlife sources. However, the co-occurrence of high levels of trace metals in sediments and high levels of FCB may result in the development of antibiotic resistance in the ambient communities. As reported by Luo (2015), surface waters in China were impacted by heavy metal pollution with the same trace metals we have found elevated in retention ponds sediments (sub-section 3.3.1). Bacteria can become resistant to antibiotics naturally through a series of mechanisms (Aminov 2009; Davies & Davies 2010), and resistance can be conferred by direct exposure to a number of environmental factors, including high temperatures, heavy metals (Baker-Austin et al. 2006; Seiler & Berendonk 2012), and exposure to antibiotics. With respect to antibiotic resistance, particular pathogens of concern include Vibrio species. Vibrio survive in brackish (1- 5 ppt) to full strength seawater (greater than 35 ppt) and may be introduced through tidal exchange, thus they likely reside in many coastal S.C. pond environments. Acquisition of resistance is significant, as V. vulnificus is responsible for more than 85% of deaths from seafood consumption. Additionally, between 1985 and 2008 the U.S. saw a greater than 100% increase in the rate of infection, and Vibrio was one of only three bacterial infections in the U.S. that saw increased rates of illness (Newton et al. 2012). Baker-Austin et al. (2008) isolated 350 V. parahaemolyticus strains from both water and sediments at three locations along the Atlantic coast of GA and S.C. and found that 99% of strains were resistant, with 24% of the isolates demonstrating resistance to 10 or more antibiotics. The increase in illnesses attributed to Vibrio strains are primarily due to these increased occurrences of wound infections (Weiss et al. 2012) caused by highly antibiotic resistant strains.
Elevated FCB sediment levels in ponds can pose a risk to estuarine environments where molluscan shellfish are harvested if these sediments are resuspended and channeled during storm events. The sequestration of both metals and bacteria in ponds may have the potential to result in increased levels of antibiotic resistance in bacteria residing in these environments. Topics that should be further studied include:
- Alterations in genes of pond bacteria associated with exposure to different antibiotics and chemical contaminants.
- Potential for lateral gene transfer among different bacterial species including E. coli (shellfish), Enterococcus (contact recreation), and Vibrio bacteria (shellfish and wound infections).
- Potential for ponds to be hospitable environments for the propagation of antibiotic-resistant microbes, and the role, if any, legacy contaminants may play.
3.5 Contaminants of Emerging Concern (CECs)
CECs are those pollutants that are not currently included in mandated routine monitoring programs, therefore there are no SQGs or water quality guidelines for these contaminants. However, CECs are a growing environmental issue due to their high-volume usage, potential for toxicity in non-target species, and increasing occurrence in the environment. These include pharmaceuticals and personal care products (PPCPs), nanomaterials, and industrial chemicals such as flame retardants and poly- and perfluorinated alkyl substances (PFAs) (Shaw & Kannan 2009; Howard & Muir 2010; Scott et al. 2012). CECs could pose a significant risk, requiring future regulation, depending on their potential for ecotoxicological and other health effects.
The principal focus of the CEC research effort to date has been on diffuse agricultural runoff, while in terms of urbanization, research has almost exclusively addressed wastewater effluents and drinking water, with relatively little attention paid to runoff sources and discharges (Scott et al. 2012; Scott et al. 2016). The San Francisco Estuary Institute (SFEI) reported results of a collaborative study between the NOAA NST Program and the Southern California Coastal Water Research Program which found higher levels of CECs in runoff than in WWTP effluent (SFEI 2013). In addition, the co-occurrence of CECs with legacy pollutants such as trace metals, PAHs, and pesticides provides additional complexity in terms of mixture exposure to aquatic organisms. The flushing of CECs from impervious surfaces during wet weather conditions may be an important source, given the potential variety of everyday materials that contain or sequester xenobiotic pollutants (e.g., solvents in wood preservatives, garage services forecourts, industrial yards, discarded recreational drugs, drug syringes, leaching from weathered plastic materials).
The fate of CECs in the environment depends on their physicochemical properties, such as solubility in water and volatility; interaction with environmental components (e.g., sorption to nanoparticles, suspended sediments); degradation processes such as photolysis and microbial degradation; and environmental characteristics such as redox conditions (Anderson et al. 2012; SFEI 2013; Shaw & Kannan 2009). Sorption to organic matter and nanoparticles may result in the release of CECs to surface waters, while the affinity of CECs to associate with suspended sediments implies their potential for removal from the water column and subsequent concentration in pond sediments. Microbial degradation may then result in the formation of new metabolites, which may themselves be benign or harmful.
3.5.1 Pharmaceutical and Personal Care Products (PPCPs)
Pharmaceuticals are being used in large quantities by the medical community and in veterinary practices. After application, these compounds and their metabolites are released into the environment via various pathways such as WWTPs (primarily human use) or field runoff (veterinary or farm sources). As stated in several reviews (e.g., Daughton & Ternes 1999), the amount of PPCPs are equivalent to the mass quantities of agrochemicals, and a further complication of PPCPs in the environment is the constant introduction of new products. By design, PPCPs are biologically active, although in many cases the direct mode of action is unknown, so it can be assumed that organisms in the environment with similar targeted receptors are also affected. The task of predicting potential ecological effects is made even more difficult due to the fact that a single compound may have additional modes of actions at other target sites for which it was not intended. In fact, single PPCPs at low concentrations in the environment have been shown to cause measurable effects due to combination with other compounds (Daughton & Ternes 1999; Silva et al. 2002; Khetan & Collins 2007). Further challenges arise from the ability to observe adverse changes in ecosystems. Although acute toxicity is usually only reached at high, environmentally non-relevant concentrations, chronic effects should not be neglected, as these may be just as detrimental to a population in the long term (Fent et al. 2006; Vieno et al. 2007).
PPCPs have also been measured in effluent from sewerage treatment plants in Charleston, S.C. and other locations around the U.S., with dominant contaminants measured including caffeine, triclosan, ibuprofen, and acetaminophen, all in the parts per trillion range (USGS 2002; Scott et al. 2016). Therefore, we may expect to measure these compounds in ponds where there is secondary treated effluent applied to land surfaces (e.g., golf courses and other spray irrigation locations). PPCPs pass through WWTPs without major degradation; current disinfection practices do not remove these compounds. Nationwide antibiotics were found at 48% of the sites sampled by USGS (2002). Cooper (2007) reported ponds reduced pharmaceuticals from sewage treatment plant effluent near Kiawah Island, S.C. No PPCPs were measured in sediments in S.C. ponds in any of the literature sources we reviewed for this study, but the antibiotic oxytetracycline was measured in surface waters in ponds (Scott et al. 2016). Oxytetracycline was measured in ponds surrounding the fairway on a golf course where secondary treated sewerage was applied to the lawn (Figure 3.8). However, concentrations at these sites were well below those that pose a risk to the environment, as established for the broad class of tetracyclines (see Scott et al. 2016).
Figure 3.8 Oxytetracycline concentrations (μg/L) measured in secondary WWTP effluent land applied to a golf course in coastal S.C. (summarized in Scott et al. 2016). Note the detectable levels in the holding pond adjacent to the green and levels found in adjoining tidal creeks draining the golf course.
3.5.2 Environmental Nanomaterials (ENMs)
Nanotechnology is a rapidly growing, innovative technology that exploits the novel properties of ENMs to develop new products and enhance the performance of existing products (Christian et al. 2008; Baalousha et al. 2014). Outdoors, ENMs can be found in a wide range of applications such as scratch resistant surface coatings, fuel additives to enhance fuel efficiency, superior paints, hydrophobic surfaces to repel water and dirt, high-performance tires, antireflection layers for road signs and panels, and road markings (Adachi 2006; Van Broekhuizen 2009; Lee et al. 2010; Van Broekhuizen et al. 2011; Dylla & Hassan 2012; Yu et al. 2012; Ugwu 2013). As with any other technology, outdoor urban nanotechnologies are likely to create pollution and waste in receiving waters due to the degradation and erosion of nano-enabled products, and these streams may further aggravate pre-existing urban stressors (Wiek et al. 2013). Because ponds store sediments, this may lead to hot spots of ENMs in ponds and potentially in the sediments of receiving water bodies.
So far, most attention has been given to the characterization of occurrences and transformations of ENMs in WWTPs, as they are predicted to be one of the major conveyances of ENMs to the environment (Gottschalk et al. 2009, 2013). However, little attention has been given to the occurrence and transformations of ENMs and incidental nanomaterials, those unintentionally generated as a side product of anthropogenic processes, in urban runoff.
No ENMs were measured in S.C. ponds in the published literature. Therefore, the authors of this chapter used transmission electron microscopy to conduct an exploratory analysis of nanomaterials in three Charleston ponds: a residential, a mixed land use, and a golf course pond. This revealed different types of nanomaterials including silica, iron oxides, and titanium dioxide in pond sediments (Figure 3.9). Silica NMs rich in chromium were found in the mixed land use pond (Figure 3.9a); titanium nanoparticles were identified in the residential (Figure 3.9b) and golf course ponds (Figure 3.9c). Further research is required to identify the different types of nanomaterials in ponds, their association with other contaminants such as trace metals and organics, the potential for these contaminants to be discharged into receiving water bodies, and their effect on these ecosystems.
Figure 3.9 Results of the transmission electron microscopy method for nanomaterials, which measures the difference in energy between the inner and outer electron shells, a characteristic of the atomic structure of the emitting element. Nanomaterials (NM) were detected in a) mixed land use pond in S.C. (Park Circle) associated with chromium and iron; b) a residential pond (Lake Edmond) associated with iron and titanium; and c) a golf course pond (Shadow Moss) associated with titanium and iron.
3.5.3 Flame Retardants
Flame retardants are derived from anthropogenic sources and can be found in many applications. Currently, the largest source of these pollutants is brominated flame retardants that can be toxic and persist in the environment for long periods of time. At this time there is little information on how these chemicals impact microorganisms or their bioaccumulative effects within the food web (Segev et al. 2009).
Weinstein et al. (2008) measured polybrominated diethylethers (PBDEs) flame retardants in sediments, which were detected in one out of 18 ponds (5.5%) and sampled at concentrations ranging from 30 to 72.8 ng/g. Higher concentrations have been measured in sediments (212 ng/g) and tissues (13,300 ng/g lipid weight in invertebrates) from natural samples (Fair et al. 2007). The limited data has shown that generally PBDE concentrations in ponds are below analytical detection limits in most ponds, and higher concentrations were detected in estuarine sediments.
With increased coastal flooding predicted due to sea level rise in the future (Union of Concerned Scientists, 2014), concerns have been raised about the potential for ponds to be impacted by sewerage overflows which may lead to additional CEC pollution sources within ponds.
- Identify, characterize, track, and quantify ENMs and NMs in all urban environmental compartments (i.e., air, impervious surfaces, and surface waters).
- Investigate the fate, pathways, and transport of urban NMs into surrounding non-urban environments.
- Develop analytical tools and methodologies to differentiate and quantify natural, engineered, and incidental NMs to determine the sources of NMs in the environment.
- Assess the interaction and toxicity of NMs (and mixtures of NMs including microplastics carried by NMs) to all sectors of the biosphere within the urban environment.
- Explore ways to mitigate and/or eliminate the most adverse bio-disrupting aspects of ENM and NM interactions within the urban environment.
- Develop or revise regulations to protect environmental and human health in light of findings from these research areas on the occurrence of ENMs in the urban environment.
- Re-investigate the geochemical cycle of metals, taking into account the significant occurrence of these metals in the form of NM components in the urban environment.
- Assess the biological impacts of contaminant mixtures and their potential impact on human and ecosystem health in both ponds and estuarine habitats.
3.6 Summary and Recommendations
Ponds in urban environments are often used as amenities by the public for boating, fishing, and swimming, and therefore have more uses other than designated BMPs for flood and water quality control. Risk assessment and routes of exposure to chemical, microbial, and emerging contaminants in ponds should receive more attention, as these BMPs are used throughout populated areas along the coast. Focus should be placed on management activities for both the environmental and public health risks posed by chemical and microbial contaminants. This review has shown there are environmental contaminants in ponds that may warrant action in terms of sediment loads, water quality concerns, and human exposure. For several chemical compounds, pond sediments had the highest concentrations when compared to concentrations measured in tidal creeks and estuarine ecosystems, indicating that they were somewhat effective in removing chemical pollutants and reducing loadings to estuarine ecosystems. Land use type was an indicator of contamination levels; all four types of urban ponds generally had higher contaminant concentrations for 29.1% of the contaminants assessed when compared to ponds from more pristine areas or to tidal creeks and estuarine systems. These urban ponds had the highest concentrations of 23% of PAHs and both legacy and contemporary use pesticides. Sediment contaminant levels varied among the urban pond types, with levels in commercial and residential ponds often having higher levels as compared to mixed land use and golf course ponds. Residential areas and golf courses use over 80 different pesticides that have a wide range of safety, including some which should not be used near the water or in combination (e.g., chlorpyrifos and atrazine).
Fecal coliform levels in NPS runoff were generally higher than levels in both ponds and estuarine surface waters. Based on the available data, ponds can be effective in treating microbial contaminants in runoff via sedimentation processes. The high levels of chemical contaminants found in ponds may enhance antibiotic resistance in potentially pathogenic bacteria such as E. coli or Vibrio spp. The levels of antibiotic resistance in E. coli are three to four times higher in urban areas than in pristine waters, and future studies are warranted given the high rate of antibiotic resistance and increased rates of illness reported for Vibrio in coastal waters. Increased resistance may be conferred by interactions with trace metals or the loading of pharmaceuticals into our urban waterways. Only limited monitoring of CECs, including pharmaceutical products, has been conducted in S.C. ponds. Flame retardants were measured in one of 18 ponds sampled, and only at very low concentrations. Antibiotic oxytertacycline was found in golf course ponds where sewerage was applied via irrigation. Limited monitoring in S.C. showed the presence of iron, titanium, and chromium nanomaterials in ponds. We conclude that future studies should consider CEC monitoring; a potential program model would be the Southern California Coastal Water Research Program. Importantly, the cumulative ecotoxicological effects of contaminants concentrated in ponds on estuarine flora and fauna is largely unknown.
With regards to the question of whether ponds are effective structural BMPs for controlling downstream water quality, our review found ponds can be successful detention pools for particle-associated contaminants. Pond design can have much to do with the success of the pond at retaining these contaminants. Effective designs include: vegetated forebays that trap coarse sediments and allow small particles to be removed gradually in the main pond, vegetated littoral shelves, and multi-pond series as opposed to single pond designs. Additionally, soil type plays a role in pollutant removal. Though small particles may take longer to be removed from the water column, research indicates that some pollutants are disproportionately found to be associated with small particles; percent clay content was correlated with storage of several pollutant classes in S.C. pond sediments. The Lowcountry region of S.C.’s coast (Beaufort and Jasper counties) is typified by fine Class D soils, comprised of clay or silt with low infiltration rates, while in the Grand Strand region larger Class A soils that are sandy and have high infiltration rates dominate. Therefore, regional studies may prove beneficial for identifying how various stormwater BMPs perform under these diverse hydrological soil conditions. Because ponds are primarily designed to promote sedimentation and 90% of this occurs between storm events, we also need to consider the impact of increasing storm intensities and frequencies predicted as a result of climate change.
As shown in this review, pond sediment contaminant levels exceeded several SQGs, indicating the potential for toxicity to benthic and aquatic life if these sediments were discharged into the estuarine environment. Pond bottoms are especially vulnerable to resuspension during storms, and evidence from recent events, namely the 1,000-year flood in October 2015 and Hurricanes Matthew and Irma (2016 and 2017, respectively), have shown that stormwater systems can be overwhelmed by periods of intense rainfall. Therefore, pond designs should be required that effectively retain larger volumes, by increasing surface area to drainage volume, designing a meandering flow path, and allowing for depths sufficient to prevent sediment resuspension by winds. As discussed in Chapter 2, retention time is vital for ensuring ponds serve as both flood and water quality control structures. Additional potential for environmental or human exposure to pond sediments occurs during and following dredging operations. According to DHEC guidelines, ponds should be dredged when the volume of the permanent pool is reduced by 25% to ensure the basin can serve the designed flood control. However, at present, there are no specific regulations in S.C. pertaining to the disposal of pond sediments, and dredged sediments are used for fill material. Thus, effective pond management strategies must evaluate the hazards posed by current sediment loads. This may include regulations that treat pond sediments as potentially hazardous wastes, because ponds can be “hot spots” for metals, pesticides, and other contaminants that have the potential to pose human health risks.
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