US Geological Survey
USGS Staff – Published Research
University of Nebraska - Lincoln Year
Fate of Sulfamethoxazole, 4-Nonylphenol,
and 17?-Estradiol in Groundwater
Contaminated by Wastewater Treatment
∗U.S. Geological Survey
†U.S. Geological Survey
‡U.S. Geological Survey
∗∗U.S. Geological Survey
††U.S. Geological Survey
‡‡U.S. Geological Survey
§U.S. Geological Survey
¶U.S. Geological Survey
?U.S. Geological Survey
This paper is posted at DigitalCommons@University of Nebraska - Lincoln.
Fate of Sulfamethoxazole,
4-Nonylphenol, and 17?-Estradiol in
Groundwater Contaminated by
Wastewater Treatment Plant
L A R R Y B . B A R B E R , *, †
S T E F F A N I E H . K E E F E ,†
D E N I S R . L E B L A N C ,‡P A U L M . B R A D L E Y ,§
F R A N C I S H . C H A P E L L E ,§
M I C H A E L T . M E Y E R ,|K E I T H A . L O F T I N ,|
D A N A W . K O L P I N ,⊥A N D
F E R N A N D O R U B I O#
U.S. Geological Survey, 3215 Marine Street, Boulder,
Colorado 80303, U.S. Geological Survey, 10 Bearfoot Road,
Northborough, Massachusetts 01532, U.S. Geological Survey,
720 Gracern Road, Suite 129, Columbia,
South Carolina 29210, U.S. Geological Survey,
4821 Quail Crest Place, Lawrence, Kansas 66049, U.S.
Geological Survey, 400 South Clinton Street, Iowa City,
Iowa 52244, and Abraxis, 54 Steamwhistle,
Warminster, Pennsylvania 18974
Received November 20, 2008. Revised manuscript received
March 19, 2009. Accepted March 30, 2009.
Organic wastewater contaminants (OWCs) were measured
in samples collected from monitoring wells located along a 4.5-
km transect of a plume of groundwater contaminated by 60
years of continuous rapid infiltration disposal of wastewater
were detected, including the antibiotic sulfamethoxazole
(NP), the solvent tetrachloroethene (PCE), and the disinfectant
1,4-dichlorobenzene (DCB). Comparison of the 2005 sampling
results to data collected from the same wells in 1985 indicates
that PCE and DCB are transported more rapidly in the
hydrophobicity. Natural gradient in situ tracer experiments
were conducted to evaluate the subsurface behavior of SX,
NP, and the female sex hormone 17?-estradiol (E2) in two oxic
zones in the aquifer: (1) a downgradient transition zone at
uncontaminated groundwater and (2) a contaminated zone
for 10 years. In both zones, breakthrough curves for the
whereas NP and E2 were retarded relative to Br-and
zone than in the transition zone. Attenuation of NP and E2
in the aquifer was attributed to biotransformation, and oxic
laboratory microcosm experiments using sediments from the
transition and contaminated zones show that uniform-ring-
recovered as14CO2in 13 days) than 4-14C E2 (20-90%
recovered as14CO2in 54 days). There was little difference in
mineralization potential between sites.
Reliance on wastewater treatment plant (WWTP) effluent
fate of organic wastewater contaminants (OWCs) in ground-
1995 were estimated at 2.9 × 105m3d-1and provided public
water supply for over 40 million people (2). Groundwater is
a major source of water for irrigation and contributes flow
(4), and their occurrence in WWTP effluents and surface
waters is well documented (5-11). Less is known about the
occurrence and fate of OWCs in groundwater (12-16).
The chemistry of WWTP effluents is complex and has
been linked to reproductive system disruption of fish living
in effluent impacted streams (17-20). The most potent
nonsteroid chemicals such as the nonionic surfactant
activity (21). Other OWCs, such as the antibiotic sul-
and there is concern about their ecological effects (24) and
development of antibiotic resistance in microbes (25).
This study was conducted at the U.S. Geological Survey’s
Cape Cod Toxic Substances Hydrology research site, an
intensively investigated plume of WWTP effluent contami-
nated groundwater (26, 27). The plume contains OWCs
including NP, the solvent tetrachloroethene (PCE), the
disinfectant 1,4-dichlorobenzene (DCB), the anionic sur-
factant linear alkylbenzene sulfonate (LAS), and many other
compounds (12, 28-30). A 3-tiered approach was used to
assess OWCs in the Cape Cod groundwater contamination
plume: (1) resampling monitoring wells 20 years after their
situ tracer experiments using SX, NP, and E2, and (3)
conducting laboratory sediment microcosm experiments
using NP and E2.
(Figure 1A) at the Cape Cod research site (26) results from
60 years of continuous rapid infiltration disposal of WWTP
effluent (about 4.6 × 107m3of effluent was discharged
between 1936 and 1995). The unconfined glacial outwash
grained sand and gravel, with groundwater flow to the
southwest at a velocity of 0.42 m d-1, a hydraulic gradient
of 0.39, and a longitudinal dispersitivity (for a 300-m long
tracer experiment) of 0.96 m (31, 32). Hydraulic properties
are relatively unweathered (deposited about 15 000 years
ago in the most recent glacial advance) and composed
primarily of quartz and feldspars with a complex suite of
* Corresponding author e-mail: email@example.com; phone: 303-
†U.S. Geological Survey, Boulder, CO.
‡U.S. Geological Survey, Northborough, MA.
§U.S. Geological Survey, Columbia, SC.
|U.S. Geological Survey, Lawrence, KS.
⊥U.S. Geological Survey, Iowa City, IA.
#Abraxis, Warminster, PA.
Environ. Sci. Technol. 2009, 43, 4843–4850
10.1021/es803292v CCC: $40.75
Published on Web 05/20/2009
2009 American Chemical SocietyVOL. 43, NO. 13, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY94843
This article is a U.S. government work, and is not subject to copyright in the United States.
with varying degrees of metal-oxide grain coatings (34).
(35). Dissolved oxygen in the uncontaminated groundwater
the unconfined aquifer) is near saturation, whereas the
contamination plume has a steep vertical gradient with an
anoxic core surrounded by oxic uncontaminated ground-
Groundwater Sampling and Analytical Procedures.
wells located along a 4.5 km longitudinal transect of the
contaminant plume (Figure 1A) using procedures described
elsewhere (37). Water for OWC analysis was filtered through
0.45-µm glass fiber filters and collected in 1-L amber glass
bottles either unpreserved or preserved with 1% (v/v)
formalin. Unfiltered samples for volatile organic compound
analysis were collected without headspace in 40-mL amber
glass vials. Samples were stored at 4 °C until analysis.
Continuous liquid-liquid extraction with methylene
chloride and gas chromatography/mass spectrometry (GC/
MS) analysis was used to measure NP and other OWCs (7).
Steroid hormones, including E2, were isolated using octa-
methoxime/trimethylsilyl derivatives were formed (7), and
analysis was by gas chromatography/tandem mass spec-
trometry (GC/MS/MS). Antibiotics, including SX, were
extraction with liquid chromatography/tandem mass spec-
trometry (LC/MS/MS). Volatile organic compounds, includ-
ing PCE, were analyzed by purge and trap GC/MS (39).
Additional OWCs were measured using methods described
in the Supporting Information.
Tracer Tests. Natural gradient in situ tracer experiments
(31, 40) were conducted during September 2006 at two
diameter polyethylene tubing vertically spaced at approxi-
mately 0.3-m intervals (37). The arrays have an extraction/
injection MLS and downgradient observation MLSs with
natureofthecontamination plumewasused toconductthe
tracer experiments in two discrete biogeochemical zones.
The transition zone site (F347) was located 0.19 km down-
gradient from the infiltration beds at a depth representing
the interface between the overlying oxic uncontaminated
groundwater and the anoxic contamination plume (Figure
located in the infiltration beds at a depth corresponding to
the historical anoxic core of the plume, which has subse-
attenuation following removal of the WWTP source in 1995
M1) and an observation MLS (S522-M4) located 2.1 m
downgradient. The water table altitude at the transition site
was 14.5 m above mean sea level and at the contaminated
site was 15.0 m above mean sea level.
using a peristaltic pump, amending the groundwater with
the groundwater solution back into the aquifer over ap-
proximately 3 h at the port from which it was withdrawn
(Figure 1B). Samples were collected from the injection MLS
at 3-h intervals for 4 days, followed by daily sampling for 5
days. The low pumping rate (<100 mL min-1) from the
sampling port had little effect on local groundwater flow,
and the protocol minimized removal of tracer cloud volume
(<5% of the injection volume was withdrawn during each
experiment). The downgradient observation MLSs were
sampled daily for 10 days.
The NP used in the tracer experiments was a branched-
chain isomeric mixture (Schenectady International,
FIGURE 1. (A) Map showing location of wells sampled during
September 2005 and the concentrations (µg L-1) of tetra-
chloroethene (PCE), 1,4-dichlorobenzene (DCB), 4-nonylphenol (NP),
and sulfamethoxazole (SX). Also shown are the locations of the
multilevel sampler arrays used for the transition zone (F347) and
contaminated zone (S522) tracer experiments. (B) Schematic of
natural gradient in situ tracer experimental procedures. [moni-
toring well F343-0036 (value after hyphen is depth in feet below
land surface) is 0.19 km downgradient from the infiltration beds;
F411-0065 is 1.7 km downgradient; F350-0064 and F350-0110 are 3.1
km downgradient; F375-0071 is 4.5 km downgradient].
4844 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 13, 2009
(>99%) single-component standards (Sigma Aldrich, St.
Louis, MO). Nominal concentrations (Table 1) were based
on the sensitivity of the analytical methods, background
concentrations, and a 100-fold factor with respect to
detection limits or background concentrations. Bromide
was analyzed by ion chromatography (37), and SX, NP,
and E2 were analyzed by enzyme linked immunosorbent
assays (ELISA), following manufacturers procedures (Abrax-
is, Warminster, PA). Detection limits were 100 µg L-1for
Br-, 2.5 µg L-1for NP, 0.015 µg L-1for SX, and 0.0015 µg
L-1for E2. A subset of samples was analyzed for SX and
NP using LC/MS/MS and GC/MS (as described above) to
confirm the ELISA results.
Breakthrough-Curve Analysis. Groundwater velocity at
the injection MLS (vinj) was estimated from the radius of the
spherical tracer cloud (xr) immediately after injection (t0)
) 0.5 (t0.5), where Ctis concentration at time t and C0is
concentration at t0. Based on a 200-L injection volume, an
and a porosity of 0.39, the radius of a spherical tracer cloud
At a flow velocity of 0.42 m d-1, a parcel of unperturbed
groundwater would move ∼0.05 m during the injection
period. Groundwater velocity at the observation MLS (vobs)
time for the Br-center of mass to reach the sampling point
(τ), and the distance between the injection and observation
FIGURE 2. Groundwater geochemical conditions in the aquifer where natural gradient in situ tracer experiments were conducted.
(A) Dissolved oxygen and specific conductance profiles for the F347 transition zone in July 2002 and June 2007. (B) Dissolved oxygen
and specific conductance profiles for the F347 transition zone during the September 2006 tracer experiment. (C) Dissolved oxygen
and specific conductance profiles for the S522 contaminated zone in July 2002 and June 2007. (D) Dissolved oxygen and specific
conductance profiles for the S522 contaminated zone during the September 2006 tracer experiment. [In 2006 the water table altitude
at F347 was 14.5 m above mean sea level and at S522 was15.0 m above mean sea level].
TABLE 1. Summary of Compounds Used in the Natural Gradient in Situ Groundwater Tracer Experimentsd
aAbbreviations used in text given in parentheses: CASRN, Chemical Abstracts Services Registry Number; Kow,
octanol-water partition coefficient; Co, concentration in the injection port following addition of tracers.
cReference (53);dReference (54).
VOL. 43, NO. 13, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 4845
and tbis time at the beginning of breakthrough and tfis time
source, i.e., center of mass at t0is at xl) 0). Relative mass
recovery (Mrel) was determined by (43)
of Br-at time t. Longitudinal dispersitivity (RL) of the
conservative tracer Br-was determined by (44)
16 ln 2
was greater than 1/2 peak concentration, and tpeakis time to
peak concentration. Relative retardation factor (Rf) was
determined by (45)
τBr-) 1 +
where τBr-is time for the center of mass of the Br-cloud to
pass the observation MLS, Fbis sediment bulk density, Kdis
sediment water distribution coefficient (sediment concen-
tration/water concentration), and η is porosity. Calculation
of Rfusing eq 5 assumes that initially the conservative and
retarded compounds are uniformly distributed throughout
during the 3-h injection phase when groundwater velocities
are relatively fast compared to the ambient groundwater
velocity. Sorption of nonpolar organic contaminants is
controlled by sediment organic carbon (46) and can be
expressed as the organic carbon normalized Kd(Koc) Kd/foc
where foc) fraction sediment organic carbon), assuming a
Rf values (42). However, consideration of such effects is
beyond the scope of this investigation.
Microcosm Experiments. Laboratory microcosm experi-
ments using aquifer sediments (47) were conducted to
evaluate NP and E2 biodegradation potential. Sulfamethox-
azole was not investigated because a14C-labeled standard
the MLS arrays at the same depths as the tracer experiments
using hollow-stem auger drilling and a wireline piston core
barrel. Uniform-ring-labeled14C 4-normal-NP (U-ring14C
4-n-NP) had an activity of 52 µCi µmole-1and 4-14C E2 had
an activity of 54 µCi µmole-1(>99% purity, American
Radiolabeled Chemicals, St. Louis, MO). The methods were
calibrated with14C HCO3-(98% purity, Sigma Biochemicals,
St. Louis, MO). Microcosms were prepared in triplicate and
consisted of 10 mL serum vials with 5 mL of saturated
of air. Duplicate sediment controls were prepared by
autoclaving 3 times for 1 h, and sediment free controls were
prepared in the same manner. Microcosms were amended
to yield initial U-ring14C 4-n-NP concentrations of 24 µg L-1
and 4-14C E2 concentrations of 34 µg L-1. Microcosms were
incubated in the dark at 23 °C for up to 54 days, and
concentrations of14CO2in the headspace were determined
gas chromatography/radiometric detection. Headspace
sample volumes were replaced with pure oxygen, and
maintenance of oxic conditions was monitored. Dissolved
phase concentrations were estimated using Henry’s Law
Results and Discussion
Plume Survey. The 2005 groundwater sampling (Figure 1A)
and NP, that were reported in the original 1985 sampling
(12). Additional OWCs that were not analyzed in the earlier
study were detected in the groundwater (15% of the 212
Tables S1 and S2). The distribution of NP was restricted to
the area near the infiltration beds, SX was detected along
most of the plume, and E2 was not detected. Changes in the
distributions of PCE, DCB, and NP between 1985 and 2005
during the 20 years between samplings was ∼3100 m),
consistent with earlier (12) estimated Rfvalues (based on
plume distributions) of 1.0 for PCE and DCB and 2.4 for NP.
Despite changes in analytical methods, concentrations for
the 1985 and 2005 samplings were similar.
Tracer Tests. Concentrations of Br-measured in the
center of the tracer cloud at t0were 99% and 94% of the
nominal values in the transition and contaminated zones.
Concentrations of NP at t0were less than nominal values,
and loss by sorption to the mixing containers and injection
system. Concentrations of SX and E2 were greater than
nominal values indicating potential analytical bias at high
concentrations where large dilutions are required. The
0.7% for Br-(n ) 12), 10.2% for SX (n ) 40), 39.0% for NP
(n ) 335), and 18.4% for E2 (n ) 60). Results for the LC/
data, having r2values of 0.95 for SX (n ) 327) and 0.90 for
NP (n ) 33).
The Br-results (presented as Ct/C0, Figure 3), were used
to estimate vinj and vobs (Table 2). In the transition zone
experiment, vinjwas 0.64 m d-1and vobswas 0.58 m d-1, and
in the contaminated zone experiment, vinjwas 0.72 m d-1
and vobswas 0.65 m d-1. In addition to differences in the
calculation methods, variability between vinjand vobsalso
reflects local heterogeneities in hydraulic conductivity (33)
the transition zone, the SX breakthrough curves coincided
mass loss, which is consistent with the reported behavior of
SX in WWTP effluents and receiving waters (9, 48-51). The
Rfvalues were similar (1.4 and 1.3, respectively). The Mrel
E2 (Table 2), resulting in estimated in situ attenuation rates
of 15% d-1and 18% d-1for NP and E2, respectively. Values
of RLfor Br-reflect aquifer characteristic controlled by the
physical structure of the sediments (32, 33) and differed
between the transition and contaminated sites. The calcu-
lated RLvalues for SX, NP, and E2 increased with increasing
relative retardation and tailing of the breakthrough curve.
zone injection MLS (Figure 3, Table 2) had similar shapes
and Ct/C0values as for the transition zone. In contrast, NP
sorption was greater in the contaminated zone than in the
transition zone and there was substantial tailing of concen-
to the injection solution after 65% of the volume had been
injected, which is reflected in faster breakthrough (Figure
4846 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 13, 2009
3C) owing to the smaller radius of the tracer cloud. In
contrast to Br-and SX, NP and E2 were not detected at
the downgradient observation MLS during the course
of the experiment. The estimated Rffor NP determined
from the injection MLS predicts a travel time of >14 days
to the MLS, whereas the experiment was terminated after
10 days and the NP and E2 breakthrough curves were not
than in the transition zone indicates geochemical het-
beds having enhanced sorption characteristics.
Retardation factors can be estimated from the Kocvalues
of the various compounds. The most water-soluble com-
pound studied was SX (Table 1), which has a log Kocof 1.8
and E2 has a log Kocof 3.3 L kg-1(54). Using the above log
Kocvalues, a bulk density of 1.59 g cm-3, a porosity of 0.39,
and a SOC of 0.005% (foc) 0.00005) for the course to fine
FIGURE 3. Breakthrough curves for bromide (Br-), sulfamethoxazole (SX), 4-nonylphenol (NP), and 17?-estradiol (E2) during the
natural gradient in situ tracer experiments conducted in September 2006. (A) Transition zone (F347) injection multilevel sampler
(MLS). (B) Transition zone observation MLS located 1.3 m downgradient. (C) Contaminated zone (S522) injection MLS. (D)
Contaminated zone observation MLS located 2.1 m downgradient. [Ct/C0) concentration at time t/concentration at time t0].
FIGURE 4. Results for oxic microcosm experiments using sediments collected from the transition zone (F347) and contaminated zone
(S522). (A) Mineralization of uniformly ring-labeled
17?-estradiol (4-14C E2) to14CO2. [Transition zone core 1 was collected at an altitude of 11.7-11.1 m above mean sea level and core 2
was collected at an altitude of 11.1-10.5 m above mean sea level (tracer injected at 11.5 m); contaminated zone core 1 was
collected at an altitude of 6.6-6.0 m above mean sea level and core 2 was collected at an altitude of 6.0-5.4 m above mean sea
level (tracer injected at 6.1 m); data are means ( standard deviations for triplicate microcosms; no recovery of14CO2was observed
in autoclaved or sediment free control microcosms.]
14C 4-normal-NP (U-ring
14C 4-n-NP) to
14CO2. (B) Mineralization of 4-14C
VOL. 43, NO. 13, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 4847
grained sand (35) results in calculated Rfvalues of 1.0 for SX,
(Table 2). Similar natural gradient in situ tracer experiments
separation of isomers with varying alkyl-chain lengths with
Rfvalues ranging from 1.0-2.9 (55, 56). Fractionation of NP
isomers was not observed (based on GC/MS results),
consistent with positional isomers on a single 9 carbon alkyl
Microcosm Experiments. Attenuation of NP and E2 was
observed during the natural gradient tracer tests in the
sediment microcosm experiments showed that 30-60% of
the U-ring14C 4-n-NP was recovered as14CO2in 13 days,
with initial linear rates of mineralization ranging from
3.8-3.9% d-1(Figure 4; Table 2). The 4-n-NP isomer used in
thus an indicator of maximum biodegradation potential.
between the transition and contaminated zones. Mineraliza-
tion of 4-14C E2 also was observed, with 20-90% of the
between the transition and contaminated zones although
there was a difference with depth.
Attenuation rates observed in the transition zone obser-
vation MLS was greater than in the laboratory microcosms
suggesting that other factors are influencing the estimates.
Attenuation in the field does not necessarily indicate
such as conversion of E2 to estrone (57). Transformation
products were not measured as part of this study. Although
microcosm and field determined removals are not directly
rather than mimic in situ conditions), the data demonstrate
that the subsurface microbial community is capable of
mineralizing 4-n-NP and E2. Similar oxic microcosm experi-
ments conducted with stream sediments reported initial
linear rates of mineralization of 7-10% d-1for U-ring14C
4-n-NP (47) and 2-6% d-1for 4-14C E2 (58). Mineralization
of E2 under oxic condition by WWTP solids (57) had an
average removal of 74% d-1. Although microcosm experi-
ments were not conducted with SX, the literature on its oxic
biodegradation reports mixed results. At low biomass con-
was observed in laboratory experiments over 28 days (48).
Under the high biomass conditions of a WWTP, SX biodeg-
radation can exceed 90% (49-51).
Implications. This study documents the subsurface
behavior of WWTP effluent derived contaminants at a range
of spatial scales from meters to kilometers and temporal
scales from hours to decades. The plume distributions
indicate long-range transport and persistence of OWCs,
observation MLSs were used to maximize the information
obtained from the resource intensive tracer experiments.
be estimated. Because breakthrough was not observed in
the downgradient observation MLS, Rfcannot be estimated
from the field results. Although hydraulically different (one
is natural gradient and the other is induced gradient), single
MLS pulsed experiments (40) are similar to single well
push-pull experiments (59) in that they can be completed
in relatively short times, only require one well, and do not
require a priori knowledge of local flow conditions. Natural
(40, 55) provide more integrated spatial and temporal data
disturbances in the injection well. However, they require
detailed knowledge of local hydraulic conditions, an ap-
The co-transport of SX and Br-suggests that SX may be
a useful tracer of subsurface contamination by WWTP
effluent. The occurrence and transport of SX in the con-
it is resistant to natural attenuation in the subsurface
environment. Previous studies have shown cotransport of
in which SX was detected, indicating long-term exposure of
the microbial community to antimicrobial compounds.
Although the effect of antibiotics on the biotransformation
of co-occurring OWCs is not known, previous in situ
experiments at the Cape Cod site showed that high con-
a sensitive measure of subsurface microbial activity (61).
TABLE 2. Results from Breakthrough-Curve Analysis of the
Natural Gradient in Situ Tracer and Sediment Microcosm
Experiments Conducted in the Transition and Contaminated
Zones in the Cape Cod Aquifera
vinj; vobs(m d-1)
aAbbreviations used in text given in parentheses: [MLS,
multilevel sampler; m, meters of transport; vinj, average
groundwater velocity at the injection MLS; vobs, average
groundwater velocity at the observation MLS; τ, hydraulic
residence time; RL, longitudinal dispersitivity; Rf, relative
mineralization, average initial linear mineralization rate.
bTracer pulse was injected into the transition zone at an
altitude of 11.5 m above mean sea level over 3.1 h, and the
observation MLS was located 1.3 m downgradient from
the injection MLS;
contaminated zone at an altitude of 6.1 m above mean sea
level over 3.3 h, and the observation MLS was located
2.1 m downgradient from the injection MLS;
ization determined from the amount of14CO2recovered (in
% of theoretical) and duration of the experiment.
cTracer pulse was injected into the
4848 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 13, 2009
Retardation of NP transport by sorption to the aquifer
sediments and mass loss by biotransformation under oxic
As suggested by the plume scale distributions, aquifer
degradation rates must be slower than those determined in
indicates that additional factors such as hydrodynamics,
temperature, oxygenation conditions, nutrient limitations,
of NP to the aquifer sediments occurred over the WWTP
plume near the infiltration beds. Concentrations and dis-
tributions of NP determined in the 2005 sampling were
Concentrations of NP measured in the groundwater were
similar to values reported for WWTP effluents and impacted
surface waters (5, 7, 8, 11), and were near those shown to
induce biological effects (19, 20, 62, 63).
The occurrence of OWCs such as SX, NP, and E2 at other
impacted groundwater sites (14, 15, 64) and in septic tank
effluents (65) indicates their potential as groundwater
contaminants, and understanding their behavior under
different hydrological, geochemical, and management con-
ditions requires further study. Once introduced into the
and be transported over long distances.
Greg Brown, James Gray, Doug Kent, Ed Furlong, Deborah
Repert, Steve Zaugg, and Kathy Conn. This research was
supported by the U.S. Geological Survey National Research
Program and the U.S. Geological Survey Toxic Substances
Hydrology Program. Use of trade names is for identification
purposes and does not imply endorsement by the U.S.
Supporting Information Available
Descriptions of the chemical analysis methods and results
from the 2005 groundwater sampling. This material is
of Augmenting Drinking Water Supplies with Reclaimed Water;
National Academy Press: Washington, DC, 1998.
in the United States in 1995. U. S. Geological Survey Circular
1200; U. S. Geological Survey: 1998.
Water Resources; U. S. Geological Survey Circular 1186; U. S.
Geological Survey: 1999.
(4) Daughton, C. G.; Ternes, T. A. Pharmaceuticals and personal
care products in the environment: Agents of subtle change.
Environ. Health. Perspect. 1999, 107 (Suppl 6), 907–938.
(5) Ahel, M.; Giger, W.; Koch, M. Behavior of alkylphenol poly-
and transformation in sewage treatment. Water Res. 1994, 28,
(6) Ternes, T. A. Occurrence of drugs in German sewage treatment
plants and rivers. Water Res. 1998, 32, 3245–3260.
(7) Barber, L. B.; Brown, G. K.; Zaugg, S. D. Potential endocrine
disrupting organic chemicals in treated municipal wastewater
and river water. In Analysis of Environmental Endocrine
ACS Symposium Series 747; American Chemical Society:
Washington, DC, 2000.
(8) Kolpin, D. W.; Furlong, E. T.; Meyer, M. T.; Thurman, E. M.;
Zaugg, S. D.; Barber, L. B.; Buxton, H. T. Pharmaceuticals,
streams, 1999-2000: A national reconnaissance. Environ. Sci.
Technol. 2002, 36, 1202–1211.
(9) Glassmeyer, S. T.; Furlong, E. T.; Kolpin, D. W.; Cahill, J. D.;
Zaugg, S. D.; Werner, S. L.; Meyer, M. T.; Kryak, D. D. Transport
discharges: Potential for use as indicators of human fecal
contamination. Environ. Sci. Technol. 2005, 39, 5157–5169.
(10) Lindqvist, N.; Tuhkanen, T.; Kronberg, L. Occurrence of acidic
pharmaceuticals in raw and treated sewages and in receiving
waters. Water Res. 2005, 39, 2219–2228.
(11) Barber, L. B.; Murphy, S. F.; Verplanck, P. L.; Sandstrom, M. W.;
along a hydrological, biogeochemical, and land use gradient:
A holistic watershed approach. Environ. Sci. Technol. 2006, 40,
(12) Barber, L. B., II.; Thurman, E. M.; Schroeder, M. P.; LeBlanc,
D. R. Long-term fate of organic micropollutants in sewage-
contaminated ground water. Environ. Sci. Technol. 1988, 22,
(13) Leenheer, J. A.; Rostad, C. E.; Barber, L. B.; Schroeder, R. A.;
Anders, R.; Davisson, M. L. Determination of the nature and
chlorine disinfection by-products of organic constituents from
Environ. Sci. Technol. 2001, 35, 3869–3876.
(14) Swartz, C. H.; Reddy, S.; Benotti, M. J.; Yin, H.; Barber, L. B.;
Brownawell, B. J.; Rudel, R. Steroid estrogens, nonylphenol
ethoxylate metabolites, and other wastewater contaminants in
groundwater affected by a residential septic system on Cape
Cod, MA. Environ. Sci. Technol. 2006, 40, 4894–4902.
(15) Godfrey, E.; Woessner, W. W.; Benotti, M. J. Pharmaceuticals in
on-site sewage effluent and ground water, western Montana.
Ground Water 2007, 45, 263–271.
(16) Barnes, K. K.; Kolpin, D. W.; Furlong, E. T.; Zaugg, S. D.; Meyer,
and other organic wastewater contaminants in the United
Statess(I) Groundwater. Sci. Total Environ. 2008, 402, 192–
(17) Jobling, S.; Nolan, M.; Tyler, C. R.; Brighty, G.; Sumpter, J. P.
1998, 32, 2498–2506.
(18) Norris, D. O.; Carr, J. A., Eds. Endocrine Disruption: Biological
Bases for Health Effects in Wildlife and Humans; Oxford
University Press: Oxford, U.K., 2006.
(19) Barber, L. B.; Lee, K. E.; Swackhamer, D. L.; Schoenfuss, H. L.
Reproductive responses of male fathead minnows exposed to
resin, and an environmentally relevant mixture of alkylphenol
compounds. Aquat. Toxicol. 2007, 82, 36–46.
(20) Vajda, A. M.; Barber, L. B.; Gray, J. L.; Lopez, E. M.; Woodling,
J. D.; Norris, D. O. Reproductive disruption in fish downstream
(21) Johnson, A. C.; Sumpter, J. P. Removal of endocrine-disrupting
chemicals in activated sludge treatment works. Environ. Sci.
Technol. 2001, 35, 4697–4703.
(22) Miao, X. S.; Bishay, F.; Chen, M.; Metcalfe, C. D. Occurrence of
antimicrobials in the final effluents of wastewater treatment
plants in Canada. Environ. Sci. Technol. 2004, 38, 3533–3541.
(23) Karthikeyan, K. G.; Meyer, M. T. Occurrence of antibiotics in
wastewater treatment facilities in Wisconsin, USA. Sci. Total
Environ. 2006, 361, 196–207.
(24) Costanzo, S. D.; Murby, J.; Bates, J. Ecosystem response to
2005, 51, 218–223.
(25) Ku ¨mmerer, K. Resistance in the environment. J. Antimicrob.
Chemother. 2004, 54, 311–320.
(26) LeBlanc, D. R. Sewage Plume in a Sand and Gravel Aquifer,
Paper 2218; U.S. Geological Survey: 1984.
Research Site. http://ma.water.usgs.gov/capecodtoxics.
fate of detergents in ground watersA field study. J. Contam.
Hydrol. 1986, 1, 143–161.
(29) Field, J. A.; Barber, L. B., II.; Thurman, E. M.; Moore, B. L.;
Lawrence, D. L.; Peake, D. A. Fate of alkylbenzenesulfonates
and dialkyltetralinsulfonates in sewage contaminated ground
water. Environ. Sci. Technol. 1992, 26, 1140–1148.
C.; Macalady, D. L.; Daniel, S. R. Identification of persistent
anionic-surfactant derived chemicals in sewage effluent and
ground water. J. Contam. Hydrol. 1992, 9, 55–78.
VOL. 43, NO. 13, 2009 / ENVIRONMENTAL SCIENCE & TECHNOLOGY 9 4849
(31) LeBlanc, D. R.; Garabedian, S. P.; Hess, K. M.; Gelhar, L. W.; Download full-text
1. Experimental design and observed tracer movement. Water
Resour. Res. 1991, 27, 895–910.
(32) Garabedian, S. P.; LeBlanc, D. R.; Gelhar, L. W.; Celia, M. A.
Cod, Massachusetts. 2. Analysis of spatial moments for a
nonreactive tracer. Water Resour. Res. 1991, 27, 911–924.
tracer test in sand and gravel, Cape Cod, Massachusetts. 3.
Hydraulic conductivity variability and calculated macrodis-
persivities. Water Resour. Res. 1992, 28, 2011–2027.
(34) Barber, L. B., II.; Thurman, E. M.; Runnells, D. D. Geochemical
heterogeneity in a sand and gravel aquifer - Effect of sediment
J. Contam. Hydrol. 1992, 9, 35–54.
sediments. Environ. Sci. Technol. 1994, 28, 890–897.
(36) Smith, R. L.; Howes, B. L.; Duff, J. H. Denitrification in nitrate-
contaminated groundwater: Occurrence in steep vertical
geochemical gradients. Geochim. Cosmochim. Acta 1991, 55,
(37) Savoie, J.; LeBlanc, D. R. Water-Quality Data and Methods of
Analysis for Samples Collected near a Plume of Sewage-
Contaminated Ground Water, Ashumet Valley, Cape Cod,
Massachusetts, 1993-1994; U. S. Geological Survey Water
(38) Meyer, M. T.; Lee, E. A.; Ferrell, G. F.; Bumgarner, J. E.; Varnes,
J. Evaluation of Tandem off-Line and on-Line Solid-Phase
Mass Spectrometry for the Analysis of Antibiotics in Ambient
Water and Comparison to an Independent Method; U. S.
Geological Survey Scientific Investigation Report 2007-5021;
U. S. Geological Survey: 2007.
(39) U. S. Environmental Protection Agency. Method 8620B. In
Methods for the Determination of Organic Compounds in
Drinking Water; Supplement III, 500 Series, U. S. EPA EPA-
600/R-95-131; U.S. Environmental Protection Agency: Wash-
ington, DC, 1995.
(40) Smith, R. L.; Baumgartner, L. K.; Miller, D. N.; Repert, D. A.;
using short term, single-well injection experiments. Microbial
Ecol. 2006, 51, 22–35.
(41) Repert, D. A.; Barber, L. B.; Hess, K. M.; Keefe, S. H.; Kent, D. B.;
LeBlanc, D. R.; Smith, R. L. Long-term natural attenuation of
of the treated wastewater source. Environ. Sci. Technol. 2006,
(42) Pang, L.; Goltz, M.; Close, M. Application of the method of
temporal moments to interpret solute transport with sorption
and degradation. J. Contam. Hydrol. 2003, 60, 123–134.
of microspheres and indigenous bacteria through a sandy
aquifer: Results of natural- and forced-gradient tracer experi-
ments. Environ. Sci. Technol. 1989, 23, 51–56.
(44) Harvey, R. W.; Garabedian, S. P. Use of colloid filtration theory
in modeling movement of bacteria through a contaminated
sandy aquifer. Environ. Sci. Technol. 1991, 25, 178–185.
(45) Freeze, R. A.; Cherry, J. A. Groundwater; Prentice-Hall, Inc.:
Englewood Cliffs, NJ, 1979.
compounds from surface water to groundwater - Laboratory
sorption studies. Environ. Sci. Technol. 1981, 15, 1360–1367.
(47) Bradley, P. M.; Barber, L. B.; Kolpin, D. W.; McMahon, P. B.;
Chapelle, F. H. Potential for 4-n-nonylphenol biodegradation
in stream sediments. Environ. Toxicol. Chem. 2008, 27, 260–
(48) Alexy,R.;Ku ¨mpel,T.;Ku ¨mmerer,K.Assessmentofdegradation
of 18 antibiotics in the closed bottle test. Chemosphere 2004,
of sulfamethazine, sulfamethoxazole, sulfathiazole, and trime-
Chem. 2005, 24, 1361–1367.
(50) Batt, A. L.; Kim, S.; Aga, D. S. Comparison of the occurrence of
antibiotics in four full-scale wastewater treatment plants with
varying designs and operations. Chemosphere 2007, 68, 428–
(51) Go ¨bel, A.; McArdell, C. S.; Joss, A.; Siegrist, H.; Giger, W. Fate
of sulfonamides, macrolides, and trimethoprim in different
wastewater treatment technologies. Sci. Total Environ. 2007,
(52) Drillia, P.; Stamatelatou, K.; Lyberatos, G. Fate and mobility of
(53) Du ¨ring, R. A.; Krahe, S.; Ga ¨th, S. Sorption behavior of nonyl-
(54) Lee, L. S.; Strock, T. J.; Sarmah, A. K.; Rao, P. S. C. Sorption and
dissipation of testosterone, estrogens, and their primary trans-
2003, 37, 4098–4105.
(55) Krueger, C. J.; Barber, L. B.; Metge, D. W.; Field, J. A. Fate and
transport of linear alkylbenzene sulfonate in a sewage-
tracer tests. Environ. Sci. Technol. 1998, 32, 1134–1142.
(56) Krueger, C. J.; Radakovich, K. M.; Sawyer, T. E.; Barber, L. B.;
Smith, R. L.; Field, J. A. Biodegradation of the surfactant linear
alkylbenzenesulfonate in sewage-contaminated groundwater:
A comparison of column experiments and field tracer tests.
Environ. Sci. Technol. 1998, 32, 3954–3961.
G. S. Mineralization of steroidal hormones by biosolids in
Technol. 2000, 34, 3925–3931.
(58) Bradley, P. M.; Barber, L. B.; Chapelle, F. H.; Gray, J. L.; Kolpin,
and testosterone in stream sediments. Environ. Sci. Technol.
2009, 43, 1902-1910.
V. Single-well “push-pull” partitioning tracer test for NAPL
detection in the subsurface. Environ. Sci. Technol. 2002, 36,
and dissolved organic compounds in a plume of contaminated
ground water. J. Contam. Hydrol. 1992, 9, 91–103.
(61) Harris, S. H.; Smith, R. L.; Suflita, J. M. In situ hydrogen
consumption kinetics as an indicator of subsurface microbial
activity. FEMS Microbial. Ecol. 2007, 60, 220–228.
(62) U. S. Environmental Protection Agency. Aquatic life ambient
water quality criteria - Nonylphenol FINAL; U. S. EPA EPA 822-
(63) Schoenfuss, H. L.; Bartell, S. E.; Bistodeau, T. B.; Cediel, R. A.;
Grove, K. J.; Zintek, L.; Lee, K. E.; Barber, L. B. Impairment of
the reproductive potential of male fathead minnows by
environmentally relevant exposures to 4-nonylphenol. Aquat.
Toxicol. 2008, 86, 91–98.
(64) Rudel, R. A.; Geno, P.; Melly, S. J.; Sun, G.; Brody, J. G.
Identification of alkylphenols and other estrogenic phenolic
compounds in wastewater, septage, and groundwater on Cape
Cod, Massachusetts. Environ. Sci. Technol. 1998, 32, 861–869.
and fate of organic contaminants during onsite wastewater
treatment. Environ. Sci. Technol. 2006, 40, 7358–7366.
4850 9 ENVIRONMENTAL SCIENCE & TECHNOLOGY / VOL. 43, NO. 13, 2009