Occurrence, Concentrations, and Risks of Pharmaceutical Compounds in Private Wells in Central Pennsylvania – Kibuye – 2019 – Journal of Environmental Quality


Groundwater is an important supply of drinking water globally. It is estimated that half of the population accesses potable water from groundwater aquifers (Smith et al., 2016). In the United States, approximately 13 million households use private wells as a drinking water source (USEPA, 2008a). Furthermore, homeowners with private wells commonly have septic tanks on their property to treat their wastewater (Schaider et al., 2014). About 25% of domestic wastewater in the United States is treated with septic systems, whereas some rural and forested areas in the northeastern United States have up to 85% of residents relying on septic systems (Swartz et al., 2006). When these systems are functioning properly, treated wastewater effluent is dispersed into subsurface absorption fields for further treatment before recharging underlying groundwater.

Although septic tanks and absorption fields are ideally installed downgradient of domestic supply wells, domestic wastewater can still degrade well water quality, especially if a septic tank is not maintained, was improperly installed, or has passed the designed lifespan. Per regulations, a horizontal distance of at least 30 m is required between a septic absorption field and a domestic supply well (US Department of Housing and Human Development, 2012). In Pennsylvania, a minimum vertical distance of 1.2 m is required between the bottom of an absorption unit and a downward limiting zone of bedrock or a seasonal high water table (Pennsylvania Code, 1997).

In addition to commonly regulated contaminant issues in private wells, such as fecal coliform, Escherichia coli, and nitrate (Swistock et al., 2013), emerging contaminants, including pharmaceuticals, ingredients in personal care products, and other organic wastewater compounds, pose potential threats to groundwater quality. Most emerging contaminants are known to incompletely degrade in a wastewater treatment system including public wastewater treatment plants (Verlicchi et al., 2012) and septic systems (Godfrey et al., 2007). Consequently, compounds that persist or are incompletely degraded may travel with wastewater plumes and contaminate groundwater, thus making septic systems important pollutant sources to surrounding domestic groundwater sources (Swartz et al., 2006; Carrara et al., 2008; Katz et al., 2010; Yang et al., 2016). Additionally, impacted groundwater can act as a source of pharmaceuticals and other classes of emerging contaminants to nearby surface water sources (Standley et al., 2008).

Nearly 10 to 20% of septic systems nationally are functioning poorly (USEPA, 2008b), thereby increasing the risk of groundwater contamination both by pollutants regulated by the USEPA’s drinking water standards, as well as unregulated emerging contaminants, such as pharmaceuticals. In a statewide study of 701 private wells in Pennsylvania, Swistock et al. (2013) found that 41% of private water wells failed at least one drinking water standard. Drinking water standards for nitrate (10 mg N L−1), pH (6.5–8.5), arsenic (0.01 mg L−1), and lead (0.015 mg L−1) were exceeded in up to 20% of sampled private wells. Because the USEPA’s Safe Drinking Water Act does not regulate private well sources, the responsibility of ensuring safe drinking water falls to the private well owner. Few homeowners regularly sample their wells to determine if the USEPA’s primary drinking water standards are met (Focazio et al., 2006).

Occurrences of pharmaceuticals in private well sources can indicate septic system failure in treating these compounds, as they are good markers of human wastewater impacts in domestic groundwater (Barnes et al., 2008; James et al., 2016) and surface water sources (Kolpin et al., 2002; Focazio et al., 2008). Furthermore, it is important to measure the levels of pharmaceuticals in drinking water sources to better understand exposure levels and associated human health risks. Conventional drinking water treatment reduces levels of pharmaceuticals in drinking water sources, albeit with differing removal efficiencies depending on treatment technology and the specific pharmaceutical compound (USEPA, 2010; Glassmeyer et al., 2017). However, communities relying on private wells and springs for domestic water supply generally perform minimal water treatment prior to use; therefore, source water concentrations are likely representative of the concentrations consumed in drinking water.

Given the prevalence (approximately one‐third) of residents in the Commonwealth of Pennsylvania that use onsite wastewater treatment systems for domestic wastewater disposal (Pennsylvania Department of Environmental Protection, 1995) and private wells and springs for domestic drinking water supplies (Penn State Extension, 2007), documenting the occurrence of pharmaceuticals in groundwater is important for informing public health decisions related to septic systems and their potential impacts on well water quality. Previous studies have monitored the impact of septic systems on domestic wells in targeted locations in Cape Cod, MA (Swartz et al., 2006; Schaider et al., 2014, 2016), New York, and New England (Phillips et al., 2015); however, no such study has been performed in Pennsylvania.

The goal of this project was to screen for the presence and concentrations of selected pharmaceuticals in private domestic groundwater sources in central Pennsylvania. The compounds selected for this study by request of the funding agency (Pennsylvania Sea Grant) were six commonly used prescription antibiotics (ampicillin, ofloxacin, sulfamethoxazole, and trimethoprim), analgesics (acetaminophen and naproxen), and a stimulant (caffeine). The targeted compounds have a wide range of physicochemical properties and are therefore expected to be representative of a broader range of pharmaceuticals of interest. A solute transport modeling approach (Harman et al., 2011) was used to predict vadose zone movement of pharmaceuticals compounds from an absorption unit to a groundwater limiting zone. The model was used to explore how the physicochemical characteristics of the pharmaceutical compounds contributed to their presence or absence in the 26 groundwater sites that were sampled. Additionally, the groundwater well concentrations were used to calculate human health risk quotients to better understand the potential risk that the presence of these compounds in the drinking water supplies may pose to residents.

Materials and Methods

A total of 26 homeowners who use groundwater for potable water supply (24 private wells and 2 springs) volunteered to participate in the study. The homeowners were recruited through the Master Well Owner Network (MWON), an Extension program through the Pennsylvania State University (https://extension.psu.edu/programs/mwon). All sites were located in the West Branch subbasin of the Susquehanna River watershed (Fig. 1). Reported well depths ranged from 12 to 130 m, and 96% of the households had active septic systems within the property (Supplemental Table S1). All samples were collected in winter 2017 (29 January–9 March). During the same sampling period, surface water samples were collected in January and March from the West Branch of the Susquehanna River (Fig. 1) to enable comparisons of pharmaceutical concentrations between surface and groundwater. All water samples were analyzed for seven pharmaceutical compounds (Table 1) selected to represent a wide range of physicochemical parameters.

Map of private groundwater well sampling locations and surface water sampling location in the West Branch of the Susquehanna River basin.

Table 1.
Physicochemical properties of selected compounds.
Compound type Pharmaceutical compound Chemical formula† Molar mass† pKa log(KOW)† Excreted Half‐life in soil
g mol−1 %
Antibiotics Ampicillin C16H19N3O4S 349.40 2.5; 7.3 1.35 30–60‡ 1 d†
Sulfamethoxazole C10H11N3O3S 253.28 1.6; 5.7 0.89 30§ 39 d¶
Ofloxacin C18H20FN3O4 361.37 5.97; 9.28 −0.39 70–98# 4 yr††
Trimethoprim C14H18N4O3 290.32 7.12 0.91 80§ 62 d¶
Analgesics Acetaminophen C8H9NO2 151.17 9.38 0.46 <5‡ <1 d‡‡
Naproxen C14H14O3 230.26 4.15 3.18 <1§§ 2 d‡
Stimulant Caffeine C8H10N4O2 194.19 10.4 −0.07 35 h¶¶

Sample Collection

Sampling kits containing two 250‐mL trace‐cleaned amber glass bottles, one 250‐mL bottle of deionized water for creating a field blank, a pair of latex laboratory gloves, sample collection and handling procedures, ice packs, a brief survey about the sampled well, and a prepaid return shipment label were assembled and mailed to each household. Participants collected their raw groundwater sample prior to any existing treatment in the household into one 250‐mL trace‐cleaned amber glass bottles. Participants then poured the shipped deionized water into the second 250‐mL sample bottle to produce the field blank sample. This field blank was collected to understand any potential contamination that may have occurred at the time of sample collection. Any detected compounds in field blanks were used to censor groundwater concentrations at each site (Supplemental Table S2). Detection frequencies in field blanks were generally lower than those in groundwater samples. Samples were shipped overnight on ice to the USDA‐ARS laboratory in University Park, PA, and stored at 4°C until processing, which occurred within 48 h of collection.

Sample Analysis

The analysis and quantification of pharmaceuticals was done using a high‐resolution accurate mass (HRAM) Q Exactive Orbitrap mass spectrometer (ThermoFisher Scientific), interfaced to the chromatography system through a heated electrospray injection (HESI) source. Analytical methods have been described in detail by Kibuye et al. (2019). In brief, water samples were filtered through a 0.22‐μm polyethersulfone (PES) syringe filter after which samples were concentrated from a 500‐μL volume to a 20‐μL volume using an inline concentrator column (Hypersil Gold aQ 20 × 2.1 mm 12 μm, ThermoFisher), then injected onto a 100‐mm × 2.1‐mm 3‐μm Hypersil Gold analytical column. For all tested analytes except ofloxacin, the method detection limit (MDL) was 0.01 μg L−1 (signal‐to‐noise ratio of 3 over background), and the method quantification limit (MQL) was 0.1 μg L−1 (signal‐to‐noise ratio of 10 over background). The MDL and MQL for ofloxacin were 0.3 and 3 μg L−1, respectively. Reporting limits for the measured analytes were set at one‐half of the MQL as specified by the USEPA Method 301 (USEPA, 2017) guidelines for determination of analytes below the quantification limit. The recoveries were determined from 12 samples for each compound for a calibration range of 0.1 to 500 μg L−1. The percentage recoveries were as follows: 90 to 95, >95, >85, 80 to 85, 70 to 80, and >75% for acetaminophen, caffeine, trimethoprim, sulfamethoxazole, naproxen, and ofloxacin, respectively.

Modeling Approach

As contaminants travel through the soil, they are subject to degradation and transformation. Half‐lives for the selected pharmaceuticals range widely, with some shorter than 1 d (caffeine and acetaminophen, Table 1) and others as long as several years (ofloxacin, Table 1). Compounds with long half‐lives are persistent in soil and may contaminate groundwater. However, it is the combination of sorption to the soil, degradation during transport, and dilution that affect occurrence at a concentration above the MDL in groundwater.

Contaminant travel rate in the vadose zone can be estimated through the determination of retardation factors, which account for the diminished contaminant travel speed relative to the bulk groundwater. The retardation factor, R, is estimated using a compound’s soil–water partition coefficient (KD) and soil characteristics shown in Eq. [1] below:


where ρb is the soil bulk density and θ is the soil water content. The KD is calculated as the product of the fraction of organic carbon in the soil and the organic carbon partition coefficient, KOC (Table 2). Calculated R and KD values are summarized in Table 2.

Table 2.
Calculated soil–water partition coefficients (KD) and retardation factors (R) of studied pharmaceutical compounds.
Pharmaceutical compound KD log(KOC)† Retardation factor (R)
L kg−1
Acetaminophen 0.04 1.32 1.1
Ampicillin 0.2 2.00‡ 1.6
Caffeine 1.5 2.87 5.3
Naproxen 0.7 2.52 2.9
Ofloxacin 88.3 4.64 258.7
Sulfamethoxazole 0.1 1.86 1.4
Trimethoprim 0.2 1.88 1.4
  • KD = Koc × foc, where foc is the fraction of organic carbon in the soil and KOC is the organic carbon partition coefficient. Source: https://pubchem.ncbi.nlm.nih.gov.
  • Values for amoxicillin were used due to lack of data for ampicillin.

A profile of a septic seepage bed designed to regulatory standards (Pennsylvania Code, 1997) is shown in Fig. 2a. The absorption–aggregate layer is typically 0.3 to 0.9 m deep with a minimum of 0.3‐m cover of backfill soil. A soil depth of at least 1.2 m, referred to as the suitable soil layer, between the bottom of the gravel absorption layer and a limiting zone of either a seasonal high water table or bedrock is required. Applying these design characteristics, septic effluent containing a mixture of pharmaceuticals discharged through a lateral pipe in the absorption layer will travel a total of 1.2 m from the bottom of the absorption layer to an underlying water table that acts as the limiting zone. To predict vadose zone transport of the seven compounds of interest from a conventional septic absorption unit to a water table limiting zone, a solute transport modeling approach following the HEIST model by Harman et al. (2011) was used (Fig. 2b).


Front view of (a) a typical septic seepage bed per regulatory standards in Pennsylvania and (b) contaminant transport model schematic, where M0 is the pharmaceutical mass leaving the absorption layer, M is the pharmaceutical mass leaving the vadose zone, R is the retardation factor, k is the first‐order degradation rate, Z is the depth of the vadose zone, and DR is the mass delivery ratio.

In addition to the compound physicochemical characteristics, transport in the vadose zone is influenced by the distribution of infiltration events such as precipitation. The range of storage in the vadose zone as function of hydrologic forcing is calculated as follows (Harman et al., 2011):


where Z is the total depth to groundwater (1.2 m), θfc is field capacity, θwp is the wilting point, and α is the mean storm depth, assumed to be 8.5 mm based on 15 yr of public rainfall data within the Susquehanna River basin in central Pennsylvania (https://www.wcc.nrcs.usda.gov). The dominant soil texture at the sampled sites was silt loam, and the corresponding soil properties used in the model are summarized in Table 3.

Table 3.
Silt loam soil properties used in model calculations.
Parameter Value Units
Bulk density (ρb) 1.33 g cm−3
Fraction of organic carbon (foc)† 0.002 g organic C g−1 soil
Vadose zone depth (Z) 1200 mm
Field capacity (θfc)‡ 0.215 v/v
Wilting point (θwp)‡ 0.115 v/v
Residual water content (θr)‡ 0.067 v/v
Saturated water content (θs)‡ 0.454 v/v
Saturated hydraulic conductivity (Ksat)‡ 622 mm d−1
Assuming uniform flow through the soil, the mean and variance of the travel time that each compound front takes to reach underlying groundwater is given by (Harman et al., 2011)


where λp is the average rainfall frequency for humid climates of 0.3 d−1 based on 15 yr of public rainfall data within the Susquehanna River basin in central Pennsylvania (https://www.wcc.nrcs.usda.gov). F is calculated by (θfc − θr)/(θfc − θwp), and θr is the residual water content (Harman et al., 2011). Since the compounds transported are subject to biodegradation, the delivery ratio (DR), which is the fraction of the initial mass that reaches groundwater, is a function of a compound’s first‐order biodegradation rate, k (half‐lives given in Table 1), and the average travel time in the vadose zone. This approach neglects the fact that some of the compounds are ionized under prevailing pH conditions, which could potentially influence their retention behavior. The mean and variance of the DR are given by (Harman et al., 2011)


Once the compound reaches the water table limiting zone, no further concentration reduction due to biodegradation or sorption is expected because of low organic carbon content and reducing conditions. Thus, pharmaceuticals that persist through the vadose zone can travel with groundwater to impact a water supply well.

Risk Calculations

The occurrence of pharmaceuticals in the environment poses potential ecosystem and human health risks. For a septic system, pharmaceuticals dispersed in the vadose zone pose an underlying risk for soil organisms (Verlicchi and Zambello, 2015). The antimicrobial triclosan has been linked to decline in microbial biomass community in soil (Zaayman et al., 2017), and antibiotics in soil can increase the presence of antimicrobial‐resistant bacteria and genes in the soil (Marti et al., 2013). The primary human exposure route to low‐concentration pharmaceuticals and other emerging contaminants is through drinking water. Because the sampled groundwater in this study is used as the primary drinking water sources in participating households, a human health risk assessment was conducted to evaluate if detected compounds occurred at levels that posed human health concern.

Assuming exposure to pharmaceuticals at average levels measured in domestic groundwater samples, a human health risk assessment was performed for an adult population characterized by a 50th percentile body weight (BW) of 60 kg and daily drinking water intake (DWI) of 2 L d−1 for a frequency of exposure (FOE) of 0.96 (350 d/365) (de Jesus Gaffney et al., 2015). Acceptable daily intakes (ADIs) of pharmaceuticals, which are the recommended daily exposure levels that pose no adverse effects on a population (Schwab et al., 2005), were used as exposure thresholds. Drinking water equivalent levels (DWELs) were then estimated as shown below (Blanset et al., 2007):


A risk quotient (RQ), which is calculated as the ratio of average pharmaceutical concentration in domestic groundwater and the estimated DWEL, was then used to characterize risk. Risk quotients >1 suggest possible human health risk from drinking water, whereas quotients <1 indicate minimal risk (de Jesus Gaffney et al., 2015). These risk calculations are, however, limited, as the impacts from mixtures of pharmaceuticals compounds and probable chronic effects are not addressed.


Occurrence in Groundwater and Surface Water

In winter 2017, samples were collected from 26 private groundwater sources in the West Branch of the Susquehanna River and surface water at the watershed outlet. All groundwater samples contained at least one of the selected pharmaceutical compounds (Supplemental Table S3), with ofloxacin as the most frequently detected compound (Table 4). Sulfamethoxazole was detected in 58% of the groundwater samples, with all levels above the MQL. The remaining compounds were detected in less than half of the samples collected, with caffeine and ampicillin detected in 46% of samples and acetaminophen and trimethoprim detected in 12% of samples. Naproxen, an anti‐inflammatory drug, was not detected in any of the groundwater samples. The most frequently detected compounds above the MDL were also detected at the highest concentrations (Fig. 3), with ofloxacin, sulfamethoxazole, and caffeine quantified at concentrations as high as 122.7, 32, and 13.1 μg L−1, respectively (Table 4). The mean concentrations detected in the groundwater samples were generally higher than the concentrations in surface water samples collected during the same period (winter 2017) at the outlet of the West Branch of the Susquehanna River (Table 4).

Table 4.
Summary of pharmaceutical concentrations in groundwater and surface water samples.
Pharmaceutical compound Groundwater samples (n = 26)† Surface water samples (n = 2)
Mean Median Min. Max. SD n > MDL n > MQL January March Mean
μg L−1 % μg L−1
Acetaminophen 0.9 0.4 0.2 2.2 1.1 12 12 0.05 0.4 0.3
Ampicillin 0.4 0.4 0.2 0.7 0.2 46 46 <MDL <MDL
Caffeine 8.1 10.4 1.7 13.1 4.7 46 46 0.05 4.5 2.3
Naproxen <MDL <MDL <MDL <MDL <MDL 0 0 <MDL <MDL
Ofloxacin 8.5 1.5 1.5 122.7 24.5 100 23 <MDL 1.5 1.5
Sulfamethoxazole 17.1 19.2 0.1 32.0 12.1 58 58 0.05 0.5 0.3
Trimethoprim 1.3 0.7 0.1 3.2 1.7 12 12 <MDL 0.5 0.2
  • MDL, method detection limit; MQL, method quantification limit. Concentrations are reported as 0.05 or 1.5 μg L−1 (ofloxacin) for samples >MDL but <MQL.

Frequencies of detection and maximum concentrations of each compound of interest in groundwater well samples.

Predicted Contaminant Travel Time and Delivery Ratios to Groundwater

Vadose zone modeling was performed to explore the extent to which pharmaceutical physicochemical characteristics influence transport in the vadose zone and subsequent groundwater impact. Estimated travel times and delivery ratio of pharmaceuticals to a groundwater water table limiting zone 1.2 m below a septic absorption field are summarized in Table 5. Travel times were estimated as a function of retardation factors, range of water storage in the vadose zone, and hydrologic variability. Average travel times varied between pharmaceuticals and ranged from <100 d for acetaminophen and sulfamethoxazole to several decades for ofloxacin.

Table 5.
Calculated mean travel and delivery ratio (M/M0) of pharmaceuticals to groundwater.
Pharmaceutical compound Travel time to groundwater Delivery ratio (M/M0) to groundwater
Mean SD Mean SD
Acetaminophen 78.1 22.8 7.9 × 10−8 6.5 × 10−5
Ampicillin 110.3 27.1 9.4 × 10−11 1.2 × 10−6
Caffeine 370.9 49.7 2.5 × 10−30 4.4 × 10−19
Naproxen 203.8 36.9 5.9 × 10−15 5.4 × 10−10
Ofloxacin 18,015 346.6 2.0 × 10−4 3.2 × 10−5
Sulfamethoxazole 98.9 25.7 0.2 8.3 × 10−2
Trimethoprim 100.1 25.8 0.3 9.5 × 10−2

In addition to sorption processes, pharmaceuticals in the vadose zone undergo microbial degradation that can further reduce contaminant mass delivered to groundwater. The delivery ratio (DR), which is defined as the fraction of the original pharmaceutical mass (M0) expected to impact groundwater, was calculated using Eq. [4], which accounts for pharmaceutical mass loss from the aqueous phase due to sorption and degradation processes in the vadose zone. The antibiotics trimethoprim and sulfamethoxazole had the highest delivery ratios, followed by ofloxacin, which was two orders of magnitude lower. The delivery ratios for the antibiotics indicate their likelihood to persist in the vadose zone and impact groundwater in comparison with the other evaluated pharmaceutical compounds.

Risk Calculations

Human health risk assessment was conducted by comparing average groundwater concentrations and calculated DWELs. All calculated RQs were <1, indicating minimal risk to human health (Table 6). The risk assessment is, however, limited, as it does not address potential additive or synergistic effects from mixtures of pharmaceuticals in drinking water. Samples generally contained more than one target pharmaceutical compound, and likely other emerging contaminants that were not analyzed for in this study.

Table 6.
Acceptable daily intakes (ADIs), estimated drinking water equivalent levels (DWELs), and corresponding risk quotients for selected compounds of interest.
Pharmaceutical compound ADI DWEL Risk quotient
μg kg−1 d−1 μg L−1 d−1
Acetaminophen 340† 10,428 0.00009
Ampicillin 0.5‡ 15 0.03
Caffeine 1.2§ 37 0.2
Naproxen 7.1§ 218 0
Ofloxacin 5.7§ 175 0.05
Sulfamethoxazole 130† 3987 0.004
Trimethoprim 4.2† 129 0.01


Comparison with Other Groundwater Data

The trends for maximum concentrations and most frequently detected pharmaceuticals in groundwater samples are similar to those found in other studies that have examined these compounds. In a nationwide study in the United States evaluating pharmaceuticals and hormones in groundwater used as drinking water sources, the frequency of detection was higher in domestic wells than in public supply wells, and sulfamethoxazole, acetaminophen, and caffeine were the most frequently detected compounds (Bexfield et al., 2019). For 20 domestic wells in Cape Cod, MA, Schaider et al. (2016) reported the highest detection frequencies for sulfamethoxazole (45%) and carbamazepine (25%) with corresponding highest maximum concentrations of 60 and 62 ng L−1, respectively. Likewise, low detection frequencies were observed for trimethoprim (5%). At a spray‐irrigation water reuse site in central Pennsylvania where wastewater effluent has been irrigated in forested and agricultural fields for >40 yr, Ayers et al. (2017) detected pharmaceuticals in 14 groundwater wells (depths ∼ 50–100 m). Caffeine was the most frequently detected compound (55% of samples), whereas ofloxacin was present at the highest concentration (up to 116.4 μg L−1). All other compounds were present in <45% of the samples, and concentrations of individual pharmaceuticals were typically <10 μg L−1 (Ayers et al., 2017). Caffeine is one of the most frequently detected compounds in groundwater samples (Verstraeten et al., 2005; Barnes et al., 2008; Fram and Belitz, 2011; James et al., 2016).

Well depth is an important factor influencing the concentrations of pharmaceuticals in groundwater. Studies have reported higher concentrations and frequencies of detection in shallow wells with depths typically <23 m in comparison with deeper wells (Verstraeten et al., 2005; Barnes et al., 2008; Bexfield et al., 2019). However, in the present study, no correlation was observed between well depth and concentrations of pharmaceuticals. This could be associated with different well selection strategies across studies. In the present study, well depth was not considered as a selection factor. Rather, homeowners volunteered for the study.

Hydrophobicity of a compound can influence its ability to sorb to organic matter and clay in the soil. Hydrophilic compounds are highly mobile and have the tendency to contaminate groundwater because of weaker retention by soil materials (Del Rosario et al., 2014). Pharmaceuticals selected for this study were generally hydrophilic, with ofloxacin, the most frequently detected compound above MDL in all groundwater samples, being the most hydrophilic, with a low octanol–water partition coefficient (KOW) [log(KOW) = −0.39, Table 1], followed by caffeine [log(KOW) = −0.07], which was detected in 46% of groundwater samples. Naproxen was the most hydrophobic compound of the selected pharmaceuticals and was not detected in any groundwater sample, potentially due to its high tendency to sorb to organic matter in soil. Similarly, other studies have reported emerging contaminants with high solubility, low log(KOW), and corresponding low log(KOC) to occur at high concentrations and frequencies in groundwater (McEachran et al., 2017; Bexfield et al., 2019). Other factors such as compound ionization at ambient soil pH values can increase mobility of hydrophobic compounds in soil due to charge repulsion with the soil matrix. For instance, ibuprofen [log(KOW) = 3.97] is anionic and was mobile in negatively charged soils, resulting in its occurrence in groundwater (Del Rosario et al., 2014). The role of compound physicochemical characteristics on their fate and transport in the vadose zone is further explored in the section below.

The one‐time sampling of groundwater at the study sites through MWON volunteers did not allow for evaluation of seasonal variations. Although emerging contaminants in surface water sources typically vary seasonally in conjunction with consumer use patterns (Hedgespeth et al., 2012), seasonal variations in the groundwater are dampened through vadose zone attenuation processes and groundwater residence times (Teijon et al., 2010). Furthermore, in a 1‐yr monitoring study at a wastewater reuse site in central Pennsylvania, insignificant seasonal variations in impacted groundwater were observed (Kibuye et al., 2019). Biel‐Maeso et al. (2018) reported lower biodegradation rates of pharmaceuticals in soil during colder seasons of the year due to reduced microbial activity, which can result in high vadose zone concentrations. Higher groundwater concentrations during such seasons are especially likely if transport is facilitated by the presence of preferential flow paths.

Comparison of Groundwater Samples with Surface Water Samples

The concentrations in groundwater samples were higher than the concentrations at the watershed outlet (Table 4). These findings contrast with a nationwide study conducted in the United States that found concentrations of antibiotics and over‐the‐counter medications, among other emerging contaminants, to be present at higher concentrations in surface water than in groundwater samples (Focazio et al., 2008). Concentrations of emerging contaminants in surface water fluctuate as a function of stream flow condition (Reif et al., 2012). Therefore, flow conditions at the time of surface water sampling likely affected the concentrations observed, with groundwater concentrations potentially varying less than surface water concentrations. High discharge during the sampling period (January–March) in the river may have resulted in a dilution effect that can lead to lower concentrations of pharmaceuticals that are predominantly present in wastewater effluent. In Pennsylvania, pharmaceuticals have been observed at higher concentrations in surface water during low‐flow conditions due to a more dominant contribution of wastewater effluent to the total stream discharge compared with the dilution effect of that wastewater signal during higher flow conditions (Reif et al., 2012).

Predicted Travel Times and Delivery Ratios to Groundwater

The modeling approach investigated the role of contaminant physicochemical properties in hydrologic‐biogeochemical filtering in a septic absorption unit. Average travel times and delivery ratios to a water table limiting zone varied by compound depending on sorption potential and biodegradability. In a septic drainfield study, Yang et al. (2016) reported sorption and microbial degradation as the major mechanisms that limited the transport of pharmaceuticals. Therefore, physicochemical characteristics of pharmaceuticals that influence sorption and degradation extent are important factors in understanding occurrence in groundwater impacted by septic tanks.

Compounds with the highest retardation factors, such as ofloxacin (Table 2), were predicted to have the longest travel time, since their transport is slowed through sorption processes (Table 5) as opposed to pharmaceuticals with retardation factors close to unity. However, due to slow biodegradation rates in soil, trimethoprim, sulfamethoxazole, and ofloxacin had the highest mean delivery ratios to groundwater. Consistently, in the current study, the highest detection frequencies and concentrations were exhibited by the antibiotics sulfamethoxazole and ofloxacin. The transport of pharmaceuticals in the vadose zone can further be influenced by ionization processes at ambient soil pHs that can increase mobility of both hydrophobic and hydrophilic compounds in soil water (Lapworth et al., 2012). For instance, sorption studies by Srinivasan et al. (2013) highlighted pH‐dependent sorption of sulfamethoxazole such that cationic form of sulfamethoxazole sorbed more at experimental pH values near pKa1 and low sorption was observed when pH ≥ pKa2, since the anionic species were dominant. This modeling approach is thus limited, as predicted groundwater delivery ratios do not incorporate pharmaceutical transport effects due to ionization characteristics.

In the studied private groundwater sources, concentrations and detection frequencies varied between sites, with detections at each site ranging from six out of the tested seven pharmaceuticals to only one detected compound (Supplemental Table S3). Similarly, concentrations of pharmaceuticals at some sites were >10 times higher than the average and median concentrations shown in Table 4. This can be due to varying household and/or site‐specific characteristics that influence pharmaceutical transport to groundwater. Household‐specific characteristics can include consumer pharmaceutical use patterns, number of household occupants, and volume of wastewater produced (Conn et al., 2006), which vary temporally within a household and spatially between households. Some site‐specific factors include varying degrees of pharmaceutical removal in septic tanks (Godfrey et al., 2007) before dispersal to the absorption field and hydraulic loading rates to the absorption fields. High pharmaceutical concentration in septic effluents dispersed at low hydraulic loading rates may have lower mass loading to the soil absorption unit than low concentrations dispersed at higher hydraulic loading rates (Conn et al., 2006).

Soil physical, chemical, and biological characteristics at the sites can affect attenuation processes as the septic effluent moves through the vadose zone. Sorption processes for antibiotics are higher in soils with high organic carbon contents and ionic strengths (Srinivasan et al., 2013), and degradation processes are higher in oxic conditions (Swartz et al., 2006; Carrara et al., 2008). Estimations of average groundwater delivery ratios and travel times in the current study did not include varying the vadose zone organic carbon content, since site‐specific averages obtained from the Web Soil Survey (https://websoilsurvey.sc.egov.usda.gov/) were used as model inputs. Other factors such as the presence of macropore flow in the soil profile can lower travel time in the vadose zone and increase contaminant mass delivered to groundwater. A 1.2‐m distance between the bottom of the absorption unit and a water‐limiting zone is used in the model. Higher average delivery ratios and shorter travel time to underlying groundwater are expected when this distance is short, as opposed to a longer separation distance with a deep suitable soil layer.

Compounds such as acetaminophen, ampicillin, caffeine, and naproxen are predicted to travel fast in the vadose zone; however, due to their rapid biodegradation rates in soil, their average delivery ratios are lower (Table 5). Acetaminophen (0.9 μg L−1) and ampicillin (0.4 μg L−1) had the lowest average concentrations in the measured private groundwater, whereas caffeine was among the most frequently detected with mean concentrations of 8.1 μg L−1 (Table 4). These compounds may have high loading rates to the absorption unit due to a more frequent consumption relative to other prescription and over‐the‐counter pharmaceuticals. For instance, the average caffeine consumption rate per person in the United States is about 210 mg d−1 (Buerge et al., 2003). Naproxen was not detected in any groundwater samples. Since ∼99% of naproxen is absorbed in the body with <1% excreted (Oosterhuis et al., 2013), concentrations in domestic waste are expected to be low. Biodegradation processes in septic tanks and soil can also lead to the transformation of naproxen into metabolites that were not evaluated in the current study. Other studies have detected naproxen in groundwater impacted by wastewater irrigation at concentrations ranging between 1.3 to 12 ng L−1 (McEachran et al., 2016, 2017), which is below the MQL for this study.

Human Health Risk Assessment

Risk assessment calculations revealed that even the highest RQ values were several orders of magnitude lower than 1 (Table 6), which is the threshold used to differentiate between a low and high risk (de Jesus Gaffney et al., 2015). This suggests minimal risk through consumption of drinking water, as substantial margins exist between average concentrations in wells and the estimated DWEL values. The World Health Organization (2012) concluded that adverse human health effects were unlikely as a result of chronic exposure to pharmaceuticals in drinking water. Risk assessments conducted by others have also found low human health impacts (de Jesus Gaffney et al., 2015).

The risk calculation may potentially underestimate overall risk, since potential mixture and chronic effects are unaccounted for and because the risk assessment does not take into consideration exposure due to other potential pathways. Therefore, although there appears to be minimal risk to human health from the concentrations present in the wells sampled in this study, the results do suggest the potential for septic tanks to impact private well water quality. Samples were only collected once as part of this study, and therefore it is unknown how much concentrations may vary in groundwater over an extended period of time. Routine monitoring may be desired to better understand the range of pharmaceutical concentrations present in private well water, in addition to other water quality parameters that have drinking water standards and better understood risks.


In winter 2017, the presence and concentrations of seven selected compounds, spanning a wide range of physicochemical parameters, were evaluated in 26 private wells in the West Branch of the Susquehanna River basin. The most commonly detected compounds were ofloxacin, sulfamethoxazole, ampicillin, and caffeine. Average concentrations of each pharmaceutical were <20 μg L−1 and were generally higher than concentrations observed at the watershed outlet during the same sampling period. A simple modeling approach was used to evaluate the influence of compound physicochemical properties on their fate and transport in the vadose zone. Estimations of average travel time in the vadose zone and delivery ratios to groundwater highlight that the extent of groundwater contamination by pharmaceutical compounds is controlled by both compound sorption potential and biodegradability. Compounds with the highest mean delivery ratios including antibiotics such as ofloxacin, sulfamethoxazole, and trimethoprim were consistently among the most frequently detected compounds at high concentrations in the analyzed groundwater samples. Transport in the vadose zone and groundwater impact were majorly a strong function of retardation factors and biodegradation rates in soil; however, influences from compound ionization are not incorporated in the model. Risk assessment calculations using measured average groundwater concentrations suggested minimal human health risk from consumption of the private well water analyzed. Nevertheless, samples collected for extended periods of time during additional seasons to capture a range of environmental conditions would provide further insight into the occurrence and concentrations of these pharmaceuticals across the watershed in both surface and groundwater sources.

Supplemental Material

The supplemental material includes (i) additional information about each of the wells sampled as part of this study (well depth and whether or not a known septic tank exists on the site), (ii) pharmaceutical concentrations for the field blanks collected at each well location, and (iii) pharmaceutical concentrations for each of the wells sampled.

Conflict of Interest

The authors declare no conflict of interest.


This study was funded by the Pennsylvania Sea Grant. Faith A. Kibuye is supported by a fellowship from The Pennsylvania State University Department of Agricultural and Biological Engineering. Heather E. Gall is supported, in part, by The Pennsylvania State University Institutes of Energy and the Environment. Herschel A. Elliot, Heather E. Gall, and John E. Watson are supported, in part, by the USDA National Institute of Food and Agriculture Federal Appropriations under Project PEN04574 and Accession no. 1004448. The authors would like to thank The Pennsylvania State University Extension program and all study participants from the Pennsylvania MWON. Mention of trade names or commercial products in this publication is solely for the purpose of providing specific information and does not imply recommendation or endorsement by The Pennsylvania State University Department or the US Department of Agriculture. All entities involved are equal opportunity providers and employers.

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