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Fate and transport of enteric microbes from septic systems in a coastal watershed.


During the 2007-2010 surveillance periods, the U.S. National Waterborne Disease and Outbreak Surveillance System reported that more than half of drinking water-associated disease outbreaks were associated with untreated or inadequately treated groundwater, indicating that contamination of groundwater remains a public health problem (Hilborn et al., 2013). Fecal contamination from humans and animals is one of the primary factors contributing to microbial pollution of both groundwater intended for drinking (Geary & Whitehead, 2001; Hagedorn, Mc Coy, & Rahe, 1981; Scandura & Sobsey, 1997; Whitehead & Geary, 2000; Yates, 1985) and coastal surface waters (Bechdol, Gold, & Gorres, 1981; Carroll, Hargreaves, & Goonetilleke, 2005; Lipp, Farrah, & Rose, 2001; Rose, Griffin, & Nicosia, 1999). Microbial contamination of groundwater continues to be a public health concern as nearly 2.7 million North Carolinians rely on private groundwater wells for drinking water (North Carolina Groundwater Association, n.d.), and approximately 60% of residences use onsite wastewater treatment systems (OSTW) in coastal North Carolina (North Carolina National Estuarine Research Reserve, 2004). After analyzing data from 2011 made available by the U.S. Environmental Protection Agency (U.S. EPA, 2012), the National Resources Defense Council reported the third highest number of beach closing and advisory days in the U.S. in 22 years. Sixty-nine percent of these beach closings/advisories were due to increased bacteria levels exceeding beach water quality standards, indicating the presence of human or animal feces in the water (Dorfman & Rosselot, 2012). From 2010 to May 2013, over 230 proclamations of polluted waters were released (not including individual closures), resulting in temporary closures of shellfish waters in North Carolina (North Carolina Department of Environmental and Natural Resources, 2012).

OWTS discharge septic effluent into the subsurface and are frequently reported as a source of groundwater contamination, resulting in environmental and public health risks (Carroll et al., 2005; Hagedorn et al., 1981; Yates, 1985; Yates & Yates, 1989). Along with high densities of OWTS, groundwater contamination can occur as a result of improper construction and maintenance of septic systems, leading to their malfunction (Ahmed, Neller, & Katouli, 2005; Geary & Gardner, 1998; Geary & Whitehead, 2001; Lipp et al., 2001; Whitehead & Geary, 2000; Yates, 1985). An average of nearly 1,500 septic systems in coastal North Carolina hydraulically malfunction each year (Humphrey, 2010), creating significant impacts on groundwater and adjacent surface waters (Ahmed et al., 2005).

This two-year study was conducted to evaluate the impact of two household septic systems on shallow groundwater and adjacent surface water quality in coastal North Carolina.

Microbial indicators of fecal contamination were studied to help improve understanding of the nature and extent of potential impacts of OWTS discharge to aquifer systems and nearby surface waters. Previous researchers have recommended using a suite of indicator microbes for better assessments of water quality, including E. coli, enterococci, and Clostridium perfringens (Griffin, Lipp, McLaughlin, & Rose, 2001). Molecular microbial source tracking for human- and animal-specific markers was also evaluated to provide additional evidence indicating whether OWTS are a source of groundwater contamination. Microbial source tracking using Bacteroides gene targets and mitochondrial DNA has been reported to identify human (Haugland et al., 2010; Shanks, Kelty, Sivaganesan, Varma, & Haugland, 2009) and animal (Caldwell & Levine, 2009; Schill & Mathes, 2008) waste sources in surface water, but application of such microbial source tracking tools to groundwater investigations has been less frequently reported. Microbial source tracking detections can be assessed in combination with fecal indicator data to evaluate whether septic systems are associated with groundwater contamination.


Site Selection and Study Design

In August 2009, seven sites in coastal North Carolina were evaluated for inclusion in this two-year study by methods previously described (Deal et al., 2007). These sites included residential homes served by OWTS, two of which met inclusion criteria (depth to groundwater and distance to surface water were in accordance with current state regulations). Year one sampling at sites 1 and 2 occurred between October 2009 and May 2010. Year two sampling was conducted from January 2011 to October 2011 at site 1 only, with the exception that the septic tank at site 2 was also sampled in year two. Site 2 groundwater was monitored during year one only due to funding constraints. Site 1 housed a single compartment septic tank, while the septic tank at site 2 contained a baffle wall dividing the tank into two compartments. Pipes were installed midway down into each tank to facilitate connection to a peristaltic pump for sample collection; septic tank lids at site 1 were modified to connect two pipes in the single compartment (one near the inlet side and one near the outlet side), while one pipe was installed in each compartment at site 2.

Piezometer development and preliminary groundwater testing was done by methods previously described (Humphrey et al., 2013). An OhmMapper TR1 electrical resistivity mapper was used to estimate the orientation of the septic plumes, while direction of groundwater flow was estimated based on the hydraulic gradient as determined from a three-point problem solution at each site (Humphrey, Deal, O'Driscoll, & Lindbo, 2010). Piezometers were installed upgradient and downgradient of OWTS flow paths for groundwater sample collection and monitoring. Piezometers were driven into boreholes at depths ranging from 1.3 to 3.7 m for collection of groundwater samples. Nineteen piezometers were installed at site 1, and 14 piezometers were installed at site 2. Most piezometers were installed adjacent to and downgradient of drainfields, and several were installed upgradient from the OWTS to assess background groundwater conditions. After installation, piezometers were purged to remove sediment and were also purged prior to each sampling event.

Sample Collection

During year one, monthly samples were collected from septic tanks while piezometers and estuary surface water samples were collected bimonthly (November 2009, January 2010, March 2010, and May 2010). Year two included bimonthly (February, April, June, August, and October 2011) sampling of piezometers at site 1, along with septic tank samples at both sites (Figures 1 and 2). Septic tank samples were collected in sterile 500-mL bottles at the outlet location followed by the inlet location. Ground and surface water samples were collected in sterile 1-L bottles. All samples were shipped on ice to the Centers for Disease Control and Prevention in Atlanta for analysis within 24 hours.

Sample Analyses

All samples were processed by membrane filtration for quantification of bacterial indicators using 0.45-[micro]m mixed-cellulose ester filters (Millipore) and subsequent culture on agar plates. Indicator microbes were cultured using the following methods: E. coli by membrane filtration and mTEC agar according to Method 1603 (U.S. EPA, 2009); enterococci by membrane filtration and mEI agar using Method 1600 (U.S. EPA, 2002); and C. perfringens by membrane filtration and mCP agar according to U.S. Geological Survey (USGS) Ohio Water Science Center method (USGS, 2007). During year 2, E. coli was enumerated in septic tank samples using Standard Method 9223B for Colilert (American Public Health Association, American Water Works Association, & Water Environment Federation, 2005).

Microbial Source Tracking

DNA Extraction

During year 2, DNA was extracted from septic tank, ground, and surface water samples at site 1 only. Samples were filtered onto 0.45-[micro]m pore-sized cellulose nitrate membrane filters and DNA extraction was performed using a soil DNA isolation kit. DNA extraction was modified by placing membrane filters, instead of a soil sample, into the provided bead solution tubes. The homogenization step was also modified using a mini-beadbeater for three minutes.

Real-Time Polymerase Chain Reaction (PCR) Five molecular targets were assayed using real-time PCR including human waste biomarkers (Bacteroides 16S rRNA [Bacteroides HF183] [Bernhard & Field, 2000; Haugland et al., 2010] and Bacteroides hypothetical human-specific protein [Bacteroides HumM2] [Shanks et al., 2009]) and animal waste biomarkers (mitochondrial cytochrome b of dog and deer hosts [Schill & Mathes, 2008] and the mitochondrial NADH dehydrogenase subunits 5 and 2 of cat and avian hosts [Caldwell & Levine, 2009]).

Statistical Analyses

Mann-Whitney tests for E. coli, enterococci, and C. perfringens were performed using Minitab 16 statistical software to determine if significant differences existed between septic tank and drainfield groundwater concentrations. Septic tank and pooled data for drainfield groundwater concentrations at sites 1 and 2 were compared to pooled data for background groundwater at both sites to help assess the impacts of OWTS on shallow groundwater.


Source Characterization of Septic Tank Wastewater

E. coli

E. coli concentrations for septic tank wastewater at site 1 ranged from 2.4 x [10.sup.3] to 9.8 x [10.sup.4] CFU/100 mL (geometric mean 2.3 x [10.sup.4] CFU/100 mL). Since the two samples collected from the septic tank at site 1 were from a nondivided compartment, geometric means reflect concentrations within the whole septic tank; no appreciable differences existed between first and second compartment data. Between the last two sampling rounds, site 1 was flooded during Hurricane Irene, which resulted in the family having to leave their home. Samples from October 2011 following the hurricane resulted in lower-than-normal levels of all indicators, with E. coli levels at 258 CFU/100 mL. E. coli concentrations in domestic wastewater typically range from [10.sup.4] to [10.sup.6] CFU/100 mL (Humphrey, 2010; Lowe et al., 2009). Site 2 had higher fecal indicator concentrations overall than site 1 during the two-year study, with first compartment levels ranging from 2.1 x [10.sup.4] to 6.8 x [10.sup.5] CFU/100 mL (geometric mean 2.0 x [10.sup.5] CFU/100 mL), and second compartment levels ranging from 1.4 x [10.sup.4] to 6.1 x [10.sup.5] CFU/100 mL (geometric mean 1.9 x [10.sup.5] CFU/100 mL).


As observed for E. coli, enterococci concentrations in the septic tank at site 2 were higher than concentrations at site 1. The geometric mean at site 1 was 2.4 x [10.sup.4] CFU/100 mL, ranging from 1.7 x [10.sup.3] to 3.7 x [10.sup.5] CFU/100 mL. The final sampling round (in which the residents were not living in the house) resulted in a lower-than-normal concentration of 490 CFU/100 mL. Enterococci concentrations in domestic wastewater are typically slightly lower than E. coli, ranging from [10.sup.4] to [10.sup.5] CFU/100 mL (Humphrey, 2010; Lowe et al., 2009). Site 2 first compartment levels ranged from 9.1 x [10.sup.3] to 3.1 x [10.sup.6] CFU/100 mL (geometric mean 5.7 x [10.sup.4] CFU/100 mL). The geometric mean concentration in the second compartment was 7.5 x [10.sup.4] CFU/100 mL, ranging from 1.1 x [10.sup.4] to 3.1 x [10.sup.6] CFU/100 mL.

C. perfringens

C. perfringens at site 1 was inconsistently detected, with only 15 detections out of 26 samples collected and concentrations ranging from 30 to 700 CFU/100 mL (geometric mean 197 CFU/100 mL). More consistent detections and higher concentrations of C. perfringens were found at site 2, ranging from 65 to 3.1 x [10.sup.4] CFU/100 mL (geometric mean 924 CFU/100 mL) for first compartment samples and 100 to 4.2 x [10.sup.4] CFU/100 mL (geometric mean 829 CFU/100 mL) in the second compartment.

Microbial Source Tracking

All wastewater samples collected from the site 1 septic tank were positive for both Bacteroides human-specific genetic markers, with average crossing threshold values of 24.1 and 31.0 for HF183 and HumM2 markers, respectively. Animal-specific genetic markers were not detected.

Microbial Characterization of Shallow Groundwater

E. coli

E. coli concentrations in groundwater generally decreased with increasing distance down-gradient from the septic tanks and drainfields at sites 1 and 2 (Figure 3). Median concentrations in groundwater beneath the drainfield (89 CFU/100 mL) were similar to those found within 15 m of the drainfield (median 95 CFU/100 mL). E. coli levels in groundwater samples taken 15-30 m away were a median of 7.9 CFU/100 mL, while median E. coli concentrations decreased to 4.0 CFU/100 mL in samples taken 30-40 m from the septic tank and 1.0 CFU/100 mL in samples collected >40 m away. Median concentrations in background piezometers (10 CFU/100 mL) were similar to those taken >15 m and beyond. Median E. coli concentrations in septic tank samples (4.0 x [10.sup.4] CFU/100 mL) were significantly greater than levels observed in background piezometers (p < .001). Drainfield and groundwater piezometers within 15 m also had significantly higher median concentrations of E. coli than those observed in background groundwater, with p-values of .031 and .012, respectively. Median E. coli levels in piezometers 15-30 m downgradient from the drainfield were also appreciably higher than those in the background piezometers, though not significantly (p = .4867). Regression analysis revealed a log-linear relationship ([R.sup.2] = .77) between median E. coli concentrations in groundwater and distance from drainfield, with outliers shown by hollow circles (Figure 4). Generally, after 20 m downgradient from the drainfield, median concentrations of E. coli in groundwater were below median background water levels.


Enterococci concentrations in groundwater samples followed similar trends that were observed with E. coli (Figure 5). Drainfield samples contained a median of 224 CFU/100 mL, and groundwater samples collected within 15 m of the drainfield had a slightly lower median density of 195 CFU/100 mL. Piezometer samples collected 15-30 m and 30-40 m had decreasing median concentrations of 47 and 12 CFU/100 mL, respectively. Groundwater samples collected >40 m from the drainfield had a low median concentration of 1.0 CFU/100 mL. Enterococci concentrations in septic tank samples were significantly greater (median 3.6 x [10.sup.4] CFU/100 mL) than median levels of 20 CFU/100 mL observed in background groundwater (p < .001).

Drainfield and piezometer samples within 15 m also had significantly higher median concentrations than those observed in background piezometers (p = .047 and .018, respectively). Median levels in groundwater downgradient 15-30 m from the drainfield were also appreciably higher than background samples, though not significantly (p = .1249). Enterococci concentrations exhibited a similar log-linear relationship as was observed for E. coli, with a consistent decrease in concentration with increasing distance from the septic tank ([R.sup.2] = .66), with the exception of outliers at site 2 that had elevated concentrations (data not shown). Generally, in groundwater 25 m downgradient from the drainfield, median enterococci concentrations fell below median background levels.

C. perfringens

C. perfringens concentrations were relatively elevated in background groundwater samples versus downgradient groundwater samples, compared to differences observed for E. coli and enterococci (Figures 3, 5, and 6). Drainfield groundwater samples, as well as those collected <15 m and 15-30 m downgradient from the drainfield, all contained similar median concentrations of 37, 32, and 40 CFU/100 mL, respectively. C. perfringens concentrations in piezometers 30-40 m and >40 m from the drainfield (median 16 and 6 CFU/100 mL, respectively) were found to be significantly lower than median background levels (median 100 CFU/100 mL) (p = .005 and .001, respectively). Median C. perfringens levels in septic tank samples (224 CFU/100 mL) were higher than levels observed in background groundwater, although not significantly (p = .273).

Microbial Source Tracking

Human fecal markers were detected in groundwater at multiple piezometers at site 1. Repetitive detections of human fecal markers occurred at piezometer locations 4 and 5 (Figure 1), which were within the drainfield and <15 m from the septic tank, respectively. At piezometer location 4, human fecal markers were detected a total of three times during two sampling events in February and April of 2011. During one of the two sampling events, only the Bacteroides HF183 marker was detected, while during the other sampling event both Bacteroides markers were detected in the same sample. Piezometer location 5 included single detections of both Bacteroides markers, each from samples collected during separate sampling events. Seven additional downgradient groundwater samples (locations 8, 9, 10, 13, 14, 15, and 17) and one surface water sample had single detections of the Bacteroides HF183 marker. The Bacteroides HF183 marker was also detected once in background piezometer at location 1. An animal biomarker (the MitoDog assay) was also detected once in the location 1 piezometer (October 2011), the only animal marker detected in any piezometer during the study.


The data from this study indicate that the OWTS at sites 1 and 2 contributed significantly to concentrations of fecal indicators in shallow groundwater. E. coli and enterococci concentrations in septic tank wastewater at the two sites were significantly elevated relative to all other sampling points (Figures 3, 4, and 5). Groundwater beneath the drainfields and within 15 m of the OWTS contained the next highest median concentrations, followed by groundwater >15 m from OWTS, background water samples, and the estuary.

OWTS at both sites contributed to elevated concentrations of E. coli and enterococci in shallow groundwater, with significant contributions (relative to background) identified in piezometers within 15 m of the septic tanks. When pooling the E. coli and enterococci data for sites 1 and 2, similar spatial trends were observed for both indicators. More specifically, the highest concentrations of E. coli and enterococci were found in septic tank effluent, followed by groundwater beneath the drainfield, groundwater within 15 m of the OWTS, groundwater >15 m, background groundwater, and finally, samples 30-40 m and beyond. The largest declines in E. coli and enterococci were observed between the septic tank and drainfield groundwater at both sites. A decline between the drainfield and groundwater downgradient also occurred but was more subtle. E. coli levels fell at or below background concentrations in groundwater 15-30 m from the drainfield, while enterococci levels fell below background levels closer to 30-40 m from the drainfield. E. coli concentrations in groundwater were lower at site 1 relative to site 2, possibly because of higher salinity in groundwater due to the influence of the nearby (<45 m) estuary, as indicated by elevated chloride levels (data not shown). Enterococci are generally more persistent in the environment than E. coli and are also more tolerant to higher saline conditions. Also, it is possible that groundwater at site 1 >40 m downgradient from the system was most influenced by the estuary and thus diluted E. coli and enterococci concentrations to lower than background wells. An aquitard (confining layer) was discovered at site 2 approximately 5 m below the surface. This aquitard may have promoted lateral, rather than vertical, movement of groundwater, thus preventing deeper groundwater contamination.

In the U.S., geometric mean ambient water quality criteria for primary contact with surface water for enterococci and E. coli in freshwater are 33 CFU enterococci and 126 CFU E. coli/100 mL, respectively (U.S. EPA, 2003). Concentrations of both fecal indicators exceeded these surface water contact standards in groundwater from both the drainfield and piezometers downgradient from the OWTS.

Concentrations of C. perfringens in background groundwater were not significantly different among those found in the septic tank, drainfield, and downgradient piezometers. C. perfringens was not consistently detected in septic tank samples at either site, and concentrations were highly variable when they were detected. C. perfringens is known to be present at lower concentrations in feces and domestic sewage than E. coli and enterococci (Vithanage, Fujioka, & Ueunten, 2011). Data from our study indicate that C. perfringens concentrations in the surficial aquifer decreased with closer proximity to the estuary, however with no apparent influence by the OWTS. The relatively higher background level of C. perfringens was not unexpected because C. perfringens is a spore-forming bacterium whose spores can survive in the environment for years (Fujioka & Shizumura, 1985), compared to E. coli and enterococci vegetative cells, which die off in the environment relatively rapidly.

The molecular source tracking results supported the fecal indicator culture data indicating that the site 1 septic system affected the microbial quality of shallow groundwater. Human waste markers were detected in background groundwater once, but were consistently detected in the septic tank and in drainfield groundwater samples three times and twice in a piezometer <15 m from the drainfield. The only animal fecal marker (for dogs) was detected once in one background piezometer. The use of five different genetic targets to detect five different potential sources increased the likelihood of identifying fecal pollution sources for this study. While both Bacteroides HF183 and HumM2 markers were detected in septic tank wastewater, the HF183 marker was detected more often in groundwater samples, suggesting that it may be a more reliable marker to use as a source tracking tool related to OWTS.

Background piezometers were upgradient from the OWTS, but background groundwater quality could have been influenced by other human and animal waste sources in the neighborhood, such as other OWTSs and animals. Such influence was also suggested by the detection of the dog mitochondrial biomarker in the background location 1 piezometer.

North Carolina requires a separation distance of 45 cm from OWTS trenches to seasonal high water table. Prior studies have shown, however, that in order to reduce the likelihood of elevated groundwater microbial concentrations, a separation distance of 60 cm or more may be needed (Humphrey, O'Driscoll, & Zarate-Bermudez, 2011; Stall, Amoozegar, Graves, & Rashash, 2014). On several occasions groundwater was less than 60 cm from the trench bottom of the OWTS at both sites.

Study Limitations

Most of our study was conducted during periods of below-average rainfall; the area was under drought conditions during the spring and summer months. The site experienced two major storm events (tornado outbreak on April 16 and Hurricane Irene on August 27), however, in which two piezometers near the estuary were torn from the ground. The hurricane caused flooding of site 1, forcing the family to leave their home while it was rebuilt. This was apparent in fecal indicator data collected from the septic tanks during the final sampling round two months after the hurricane.

Chloride concentrations and specific conductivity were also elevated during the final sampling round in October 2011, indicating that brackish floodwaters had impacted the septic tank (data not shown). These extreme weather events likely contributed to increased variability during the study.

The funding level for our study did not allow for inclusion of additional households. Field and laboratory work for such intensive sampling would make such a study very expensive, and funding only allowed for inclusion of two households.


Data from our study showed that OWTS at both sites contributed to significantly elevated concentrations of E. coli and enterococci in groundwater beneath the drainfields relative to background groundwater concentrations. Groundwater from impacted piezometers contained E. coli and enterococci at concentrations exceeding ambient water quality criteria. In addition, molecular source tracking data demonstrated that human fecal markers were most often associated with piezometers downgradient of the site 1 septic system and could be detected as far away as 40 m from the drainfield. The data indicate that molecular fecal source tracking assays can be a useful addition to the "toolbox" approach to detect and identify sources of fecal contamination in water samples. While data from our study indicated that the OWTS impacted the microbial quality of shallow groundwater, general trends from sites 1 and 2 indicated that E. coli and enterococci concentrations decreased with increasing distance downgradient from OWTS (toward the estuary). Median concentrations of E. coli and enterococci in shallow groundwater dropped to below median background concentrations after approximately 30 m downgradient of the OWTS.

This study provides evidence of OWTS influence on groundwater quality and indicates the benefits and constraints for a diverse set of fecal indicator and fecal source tracking methods for groundwater quality studies. In particular, for the study location in North Carolina, the findings of our study provide information that can be used to assess North Carolina OWTS setback and separation distance regulations. Since the study sites were adjacent to an estuary, results suggest that current OWTS setbacks of 15-30 m may not be sufficiently protective to prevent elevated microbial concentrations in shallow groundwater from reaching nearby surface water and adjacent waterways (e.g., shellfish harvesting areas). Future studies are warranted to evaluate the potential impacts of OWTS on groundwater quality in different hydrogeological areas and how observed impacts relate to regulated setback and separation distances for these systems.

Acknowledgements: We acknowledge the contributions of Dr. David Lindbo from North Carolina State University during the planning and site evaluation phases of the study. Funding for this study was provided in part through a project grant from the National Environmental Health Association. The findings and conclusions in this report are those of the authors and do not necessarily represent those of CDC. Use of trade names and commercial sources is for identification only and does not imply endorsement by CDC or the U.S. Department of Health and Human Services.


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Chandra L. Schneeberger

National Center for Emerging and Zoonotic Infectious Diseases Centers for Disease Control and Prevention IHRC, Inc.

Michael O'Driscoll, MS, PhD

Department of Geological Sciences Institute for Coastal Science and Policy/Coastal Water Resources Center East Carolina University

Charles Humphrey, MS, PhD, REHS

Environmental Health Sciences Program East Carolina University

Keaton Henry, MS

Department of Geological Sciences East Carolina University

Nancy Deal, MS, REHS

North Carolina Department of Health and Human Services

Kathy Seiber, MS

National Center for Emerging and Zoonotic Infectious Diseases Centers for Disease Control and Prevention Association of Public Health Laboratories Emerging Infectious Disease Fellowship

Vincent R. Hill, PhD, PE

National Center for Emerging and Zoonotic Infectious Diseases Centers for Disease Control and Prevention

Max Zarate-Bermudez, MSc, MPH, PhD

National Center for Environmental Health Centers for Disease Control and Prevention

Corresponding Author: Vincent Hill, National Center for Emerging and Zoonotic Infectious Diseases, Centers for Disease Control and Prevention, 1600 Clifton Road, NE, Mailstop D-66, Atlanta, GA 30329.

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Author:Schneeberger, Chandra L.; O'Driscoll, Michael; Humphrey, Charles; Henry, Keaton; Deal, Nancy; Seiber
Publication:Journal of Environmental Health
Article Type:Report
Geographic Code:1U5NC
Date:May 1, 2015
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