ARTICLE IN PRESS
Water Research 39 (2005) 5033–5044
www.elsevier.com/locate/watres
Removal of neutral chloroacetamide herbicide degradates
during simulated unit processes for drinking water treatment
Michelle L. Hladik, A. Lynn Roberts, Edward J. Bouwer
Department of Geography and Environmental Engineering, Johns Hopkins University, 313 Ames Hall, 3400 N. Charles St.,
Baltimore, MD 21218-2686, USA
Received 22 February 2005; received in revised form 26 September 2005; accepted 5 October 2005
Abstract
Four chloroacetamide herbicides and 20 neutral chloroacetamide derivatives (known to occur as their environmental
degradates) were subjected to simulated drinking water treatment (coagulation, oxidation and adsorption). Coagulation
with alum and ferric chloride, at doses for optimum turbidity removal, provided little to no (o10%) removal of parent
herbicides or neutral degradates. Chlorination with 6 mg/L applied free chlorine for 6 h was able to achieve 100%
removal of those degradates lacking an acetanilide substituent; compounds possessing this functional group exhibited
low (0–16%) removal efficiencies. Products were generally not identified, except in the case of dimethenamid and its
deschloro degradate, both of which formed a single ring-chlorination product on their ready reaction (84% and 96%
removal, respectively) with aqueous chlorine species. Treatment with ozone at an applied dose of 3 mg/L for 30 min
proved effective (60–100%) at transforming all of the compounds under investigation to unidentified products. The
parent herbicides and neutral degradates underwent adsorption by powdered activated carbon (PAC). Adsorption
capacities (Freundlich K constants) correlated with Kow values.
r 2005 Elsevier Ltd. All rights reserved.
Keywords: Alachlor; Acetochlor; Metolachlor; Dimethenamid; Herbicide degradates; CCL contaminants
1. Introduction
Chloroacetamide herbicides (acetochlor, alachlor,
metolachlor and dimethenamid) are among the most
widely used agricultural herbicides in the United States,
with approximately 77–98 million lb of active ingredients
applied to crops in 2001 (Kiely et al., 2004). Peak
concentrations of these pre-emergent herbicides in
runoff tends to occur in May and June (Thurman et
al., 1996). These micropollutants are frequently detected
at micrograms per liter concentrations in surface water
Corresponding author. Tel.: +1 410 516 7437;
fax: +1 410 516 8996.
E-mail address: [email protected] (E.J. Bouwer).
(Battaglin et al., 2000) and ground water (Barbash et al.,
2001), including water sources used to supply drinking
water (Kolpin et al., 1998a). Chloroacetamide herbicides
have also been detected in finished drinking water
(Blomquist et al., 2001; Coupe and Blomquist, 2004).
The parent chloroacetamide herbicides are known to
be toxicants, and are listed as probable (alachlor and
acetochlor) or possible (metolachlor) human carcinogens (US EPA, 1999). Even at micrograms per liter
concentrations, these herbicides may adversely affect
human health (Dearfield et al., 1999). The US EPA has
currently only defined a drinking water maximum
contaminant level (MCL) for alachlor (2 mg/L; US
EPA, 2001). Other chloroacetamide herbicides, however, are on the US EPA’s Contaminant Candidate List
0043-1354/$ - see front matter r 2005 Elsevier Ltd. All rights reserved.
doi:10.1016/j.watres.2005.10.008
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M.L. Hladik et al. / Water Research 39 (2005) 5033–5044
(CCL) of compounds that could be subject to future
regulation (US EPA, 1998).
Because these herbicides have been frequently encountered in drinking water sources, prior research has
focused on their removal during drinking water treatment. Many typical water treatment processes, such as
coagulation and chlorination, have been found to be
ineffective at removing alachlor and metolachlor (Miltner et al., 1989). Ozonation has been shown to be
successful in removing the parent herbicides (Miltner et
al., 1987; Griffini et al., 1999; Acero et al., 2003), but
does not completely mineralize these compounds (Somich et al., 1988). Activated carbon has also been shown
to provide substantial removal of the parent compounds
(Miltner et al., 1989; Griffini et al., 1999).
While the parent chloroacetamide herbicides have
been the primary concern of previous research, many of
their environmental degradates are also known to occur
in drinking water sources (Gustafson et al., 2003;
Hladik, 2005). Although toxicological data are available
for the parent chloroacetamide herbicides, considerable
uncertainty surrounds the toxicity of their environmental degradates. Some degradates, such as alachlor ESA
(Heydens et al., 1996, 2000), are thought to be relatively
non-toxic. Other chloroacetamide herbicide degradates
may have a toxicity similar to that of the parent
compound. Hydroxyalachlor [2-hydroxy-20 ,60 -diethylN-(methoxymethyl)acetanilide] has been shown to possess a toxicity similar to alachlor (Kross et al., 1992).
Two additional alachlor degradates, 2-chloro-20 ,60 diethylacetanilide and 2-hydroxy-20 ,60 -diethylacetanilide, have been shown to be weakly mutagenic (Tessier
and Clark, 1995), and 2-chloro-20 ,60 -diethylacetanilide
has been shown to form DNA adducts (Nelson and
Ross, 1998; Nesnow et al., 1995). The primary aniline
degradates of metolachlor and alachlor, 2-ethyl-6methylaniline and 2,6-diethylaniline, have been reported
to be teratogenic (Osano et al., 2002) and are promutagens (Kimmel et al., 1986).
Even though environmental degradates of the chloroacetamide herbicides are not currently regulated in the
United States, they have also been included on the US
EPA’s CCL (US EPA, 1998). More stringent European
Union guidelines (EU, 1998) state that the summed
concentration of a pesticide and its relevant metabolites
in drinking water must not exceed 0.1 mg/L. Given the
human health risks that may be posed by chloroacetamide degradates and the possibility of their regulation in
the future, an understanding of their behavior during
drinking water treatment would be of value.
Two of the most common chloroacetamide degradates
encountered in natural waters are the ethane sulfonic
acids (ESAs) and the oxanilic acids (OAs) (Kolpin et al.,
1998b). Removal of the chloroacetamide ESAs and OAs
during water treatment at 175 community water systems
was studied by Gustafson et al. (2003). This study found
that activated carbon can remove both the parent
herbicides and their ionic ESA and OA degradates.
Removal efficiencies were, however, lower for the
degradates than for the parent compounds; in one case,
the parent herbicides exhibited removals of 90% using
powdered activated carbon (PAC), while the degradates
revealed removals of only 40–45%. The acidic degradates are more water soluble and therefore are less likely
to adsorb to activated carbon than the parent herbicide.
In another study (Verstraeten et al., 2002), the ESA and
OA degradates (and the parent herbicides) were found
to decrease in concentration (by 79%) after ozonation at
a water treatment plant.
Prior work in our laboratory has revealed the
occurrence of many neutral degradates of chloroacetamide herbicides in water samples obtained from the
Chesapeake Bay (Hladik et al., 2005) and in drinking
water samples obtained from 12 Midwestern US water
treatment facilities (Hladik, 2005). These neutral chloroacetamide degradates have not been routinely monitored in drinking water systems; the effect of drinking
water treatment processes on these compounds is,
therefore, unknown. In the present study, the four
parent chloroacetamide herbicides (alachlor, acetochlor,
metolachlor and dimethenamid) and 20 of their neutral
environmental degradates were subjected to coagulation
(alum and ferric chloride), oxidation (chlorination and
ozonation) and adsorption (activated carbon) in order
to compute removal efficiencies in simulated drinking
water treatment processes. Degradates chosen have been
shown to occur in natural or affected environmental
samples (Hladik, 2005; Hladik et al., 2005). Structures of
the compounds under investigation, along with the
numbering scheme used to designate each analyte, are
provided in Fig. 1. Atrazine is also included in our
sorption studies, as it is currently the second most widely
used agricultural herbicide in the United States (Kiely et
al., 2004). It has an MCL of 3 mg/L (US EPA, 2001), and
many drinking water plants that are affected by
herbicide runoff optimize their addition of PAC to
achieve a target level for atrazine.
2. Materials and methods
2.1. Chemicals
Alachlor (I) [2-chloro-20 ,60 -diethyl-N-(methoxymethyl)acetanilide], metolachlor (IX) [2-chloro-20 -ethyl-60 methyl-N-(2-methoxy-1-methylethyl)acetanilide], deschloroacetylmetolachlor propanol (XV) [2-[(2-ethyl-6methylphenyl)amino]-1-propanol], acetochlor (XVI) [2chloro-20 -ethyl-60 -methyl-N-(ethoxymethyl)acetanilide],
dimethenamid (XXIII) [2-chloro-N-(2,4-dimethyl-3thienyl)-N-(2-methoxy-1-methylethyl)acetamide], and atrazine (XXV) [2-chloro-4-ethylamino-6-isopropylamino-s-
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Alachlor and its degradates
Cl
O
I
OH
O
II
N
III
O
N
Cl
O
O
N
O
IV
O
HN
OH
V
O
HN
Metolachlor and its degradates
O
OH
VI
VII
N
HN
O
O
O
XII
N
IX
NH2
HO
O
XI
VIII
XIII
O
N
Cl
XVI
N
O
O
O
N
O
N
XV
NH
Metolachlor or acetochlor degradates
OH
O
XVII
Cl
O
XVIII
N
O
XIX
O
N
XXI
XXIII
Cl
O
XXII
NH2
XX
O
O
Cl
XXV
N
N
O
N
HN
S
O
HN
Atrazine
XXIV
N
OH
O
HN
Dimethenamid and its degradate
O
X
NH
Metolachlor or acetochlor
degradates (continued)
HN
OH
HO
XIV
O
Acetochlor and its degradates
O
O
N
Cl
O
Cl
N
NH
S
Fig. 1. Structures of parent herbicides and neutral degradates under study. Structures I–VIII represent alachlor and its degradates,
structures IX–XV represent metolachlor and its degradates, structures XVI–XVIII acetochlor and its degradates, structures XIX–XXII
are those degradates that can result from either metolachlor or acetochlor, structures XXIII–XXIV represent dimethenamid and its
degradate, and XXV is atrazine.
triazine] were obtained from Chem Service (West
Chester, PA, USA). 2,6-Diethylaniline (VIII) and 2ethyl-6-methylaniline (XXII) were obtained from Aldrich (Milwaukee, WI, USA).
Hydroxyalachlor (II) [2-hydroxy-20 ,60 -diethyl-N(methoxymethyl)acetanilide], deschloroalachlor (III)
[20 ,60 -diethyl-N-(methoxymethyl)acetanilide], 2-chloro20 -60 -diethylacetanilide (IV), 2-hydroxy-20 -60 -diethylacetanilide (V), 2-hydroxy-20 -60 -diethyl-N-methylacetanilide (VI), 20 -60 -diethylacetanilide (VII), hydroxymetolachlor (X) [2-hydroxy-20 -ethyl-60 -methyl-N-(2methoxy-1-methyl-ethyl)acetanilide], deschlorometola-
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M.L. Hladik et al. / Water Research 39 (2005) 5033–5044
chlor (XI) [20 -ethyl-60 -methyl-N-(2-methoxy-1-methylethyl)acetanilide], a morpholinone derivative of metolachlor (XII) [4-(2-ethyl-6-methylphenyl)-5-methyl-3morpholinone], metolachlor propanol (XIII) [2-chloro20 -ethyl-60 -methyl-N-(2-hydroxy-1-methyl-ethyl)acetanilide], deschloroacetylmetolachlor (XIV) [20 -ethyl-60 methyl-N-(2-methoxy-1-methylethyl)aniline], hydroxyacetochlor
(XVII)
[2-hydroxy-20 -ethyl-60 -methyl-N(ethoxymethyl)acetanilide], deschloroacetochlor (XVIII)
[20 -ethyl-60 -methyl-N-(ethoxymethyl)acetanilide],
2chloro-20 -ethyl-60 -methylacetanilide (XIX), 2-hydroxy20 -ethyl-60 -methylacetanilide (XX), 20 -ethyl-60 -methylacetanilide (XXI) and deschlorodimethenamid (XXIV)
[N-(2,4-dimethyl-3thienyl)-N-(2-methoxy-1-methylethyl)acetamide] were synthesized in our laboratory. Procedures are given elsewhere (Hladik et al., 2005).
compounds). Relatively high concentrations (compared
to values measured in untreated drinking water using
gas chromatograph/mass spectrometer (GC/MS) methods; Hladik, 2005) were necessitated by the higher
detection limits of the gas chromatograph-nitrogen
phosphorus detector (GC-NPD) and high-pressure
liquid chromatograph with diode array detector
(HPLC-DAD) analytical methods employed in the
current study. Each compound group was tested in
triplicate with three 1-L beakers for each coagulant. The
coagulant was added at the optimum dose as determined
from the previous step (alum, 30 mg/L; FeCl3, 20 mg/L).
Upon settling, aliquots were removed to determine
remaining herbicide concentrations as described in
Section 2.5 below.
2.2. Coagulation
2.3. Oxidation
2.2.1. Optimal coagulant dose
Jar tests were carried out following procedures given by
Hudson (1981) using water from Loch Raven Reservoir
(which supplies drinking water to the Baltimore, MD
metropolitan area) to determine the appropriate coagulant dose. Alum (aluminum sulfate, Al2(SO4)3 18H2O;
Aldrich) and ferric chloride (FeCl3; Aldrich) were dosed at
0, 10, 20, 30, 40 and 50 mg/L. Six 1-L beakers filled with
the Loch Raven Reservoir water (pH 7.1) were placed in a
laboratory stirrer (Phipps and Bird; Richmond, VA,
USA). Coagulants (as a 1.5% by weight dosing solution)
were added via pipette to yield the appropriate concentrations. Samples underwent 2 min of rapid mixing at
100 rpm, followed by 60 min of slow mixing at 20 rpm.
After the completion of the slow mix period, the samples
were allowed to settle for 1 h. Supernatant samples were
collected for measurement of turbidity (HF Scientific
DRT 100B; Ft. Meyers, FL, USA), UV-absorbance at
254 nm (Shimadzu UV-1601; Columbia, MD, USA) and
TOC concentration (Dohrmann Phoenix 8000 UVpersulfate TOC analyzer; Mason, OH, USA). The
optimum coagulant dose was chosen as the dose at which
turbidity after settling was minimized.
Reductions in TOC concentrations following coagulation were checked for compliance with the Enhanced
Coagulation Rule (White et al., 1997). The Enhanced
Coagulation Rule is a regulatory strategy to limit the
formation of disinfection by-products through the
removal of TOC during coagulation. For the Loch
Raven water under study, which has a TOC of 3 mg/L,
a 40% reduction in the TOC upon coagulation is
required. Addition of alum and ferric chloride met the
enhanced coagulation requirement.
2.3.1. Chlorination
The free chlorine dosing solution was prepared by
diluting a 4–6% aqueous solution of NaOCl (Fisher
Scientific; Fairlawn, NJ, USA) to a concentration of
approximately 200 mg/L free chlorine and adjusting the
pH to 7. This solution was standardized iodometrically
(Standard Test Method 4500 Cl-G; Greenberg et al.,
1992). Serum bottles (60 mL) were prepared containing
59 mL of Loch Raven Reservoir water (pH 7.1), 50 mg/L
of the test compound and 6 mg/L free chlorine. As
disinfection with free chlorine typically occurs at
2–3 mg/L, this concentration may seem to exceed values
commonly used during drinking water chlorination; it is,
however, comparable to concentrations employed at
plants treating seasonal pesticides (Miltner et al., 1989).
Test compounds were introduced by adding concentrated acetone stock solutions; the acetone volumes used
were p10 mL. Each compound was treated individually
in triplicate bottles. The bottles were closed with Teflon
lined rubber stoppers with crimp caps and were mixed
(after dosing) on a stir plate for 1 min. The bottles were
then incubated for 6 h at 2271 1C in the dark. At the
end of the reaction period, the samples were immediately
extracted for analysis as described in Section 2.5 below,
eliminating the need for a chlorine-quenching agent.
One of the three bottles was also tested for residual
chlorine; all samples tested had a residual free chlorine
concentration 41 mg/L (DPD colorimetric method;
Standard Test Method 4500 Cl-G; Greenberg et al.,
1992). A control experiment conducted with 10 mL of
acetone plus 6 mg/L free chlorine in Loch Raven water
revealed a residual chlorine concentration that was not
significantly different (at the 95% confidence level) from
that obtained upon chlorination of Loch Raven water in
the absence of acetone, suggesting negligible consumption of the chlorine by acetone under our experimental
conditions.
2.2.2. Compound removal
The compounds of interest were spiked into Loch
Raven water at 50 mg/L (in groups of three to four
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2.3.2. Ozonation
Stock ozone dosing solutions were prepared according
to Standard Method 2350 (Greenberg et al., 1992) by
bubbling an ozone/oxygen gas stream (OREC V5-O
ozone generator; Phoenix, AZ, USA) through Milli-Q
water (Millipore; Billerica, MA, USA). Milli-Q water
(800 mL) sitting in an ice bath with continuous stirring
was ozonated for 1 h. The final ozone concentration in
the water was measured via an indigo colorimetric
method with a HACH (Loveland, CO, USA) DR-100
colorimeter. With this protocol, typical ozone concentrations in the aqueous dosing solution were 24–26 mg/L.
Reactivity to ozone was studied in triplicate 60 mL
serum vials that contained Loch Raven Reservoir water
(pH 7.1) and a single test compound (50 mg/L). Test
compounds were introduced as in the chlorination
experiments. The applied ozone dosage (3 mg/L) was
approximately 1 mg of ozone per mg of TOC in the
sample, a value that falls within the dosage range typical
in water treatment practice (MWH, 2005). The bottles
were prepared without headspace and were sealed with
Teflon-lined rubber stoppers and crimp caps. The bottles
were stirred and incubated for 30 min in the dark at
2271 1C. At the end of the reaction period, the samples
were immediately extracted as described in Section 2.5
below for analysis of the test compound; ozonequenching agents were not used. Ozone concentrations
were not remeasured upon completion of the reaction.
2.4. Activated carbon
Calgon (Pittsburgh, PA, USA) water powdered high
activity (WPH) PAC was used for these experiments. The
PAC was dried in an oven for 3 h at 150 1C before use.
Adsorption isotherms were determined at room temperature (2271 1C) in a series of five 60-mL serum bottles with
Teflon-lined crimp caps. Bottles contained filtered
(0.7 mm; Millipore) Loch Raven Reservoir water, 50 mg/
L of each individual test compound, and a known mass of
PAC (prepared by diluting a PAC slurry in deionized
water). A control experiment, conducted in the absence of
PAC, did not reveal any loss of the test compounds. The
PAC concentrations ranged from 0.5 to 8 mg/L. Bottles
were covered in aluminum foil and were placed on a roller
table at 45 rpm for 5 days. In separate 10-day kinetic
studies, a contact period of 5 days was found sufficient to
attain 99% of the long-term value. After 5 days, the bottle
contents were filtered through a 0.45 mm nylon membrane
filter (Millipore) and were extracted as described in
Section 2.5 below for analysis of the test compounds.
2.5. Quantitative analysis
2.5.1. GC-NPD
Compounds analyzed via this method were: alachlor
(I), deschloroalachlor (III), 2-chloro-20 -60 -diethylaceta-
5037
nilide (IV), 20 -60 -diethylacetanilide (VII), 2,6-diethylaniline (VIII), metolachlor (IX), deschlorometolachlor (XI),
deschloroacetylmetolachlor (XIV), acetochlor (XVI),
deschloroacetochlor
(XVIII),
2-chloro-20 -ethyl-60 0
methylacetanilide (XIX), 2 -ethyl-60 -methylacetanilide
(XXI), 2-ethyl-6-methylaniline (XXII), dimethenamid
(XXIII), deschlorodimethenamid (XXIV) and atrazine
(XXV).
Three 10-mL aliquots of each sample were successively extracted with 1 mL of dichloromethane for 1 min;
the three 1-mL extracts were then combined. The
dichloromethane was dried under a gentle stream of
nitrogen at ambient temperature and the sample was
reconstituted in 0.50 mL of toluene containing 2-nitrom-xylene as the internal standard. Injections (1 mL) were
made on-column onto a Carlo Erba (San Jose, CA,
USA) Mega 2 GC with a flameless nitrogen phosphorus
detector (NPD). A DB-5 (Agilent; Palo Alto, CA, USA)
30 m length 0.25 mm ID 0.25 mm film thickness
column was used to effect separations. The GC
temperature program was 110 1C for 1 min, 10 1C/min
to 290 1C, with a subsequent 5 min hold at 290 1C.
Extraction efficiencies ranged from 96% to 100%, and
detection limits were 0.5 mg/L.
2.5.2. HPLC-DAD
Compounds analyzed via this method were: hydroxyalachlor (II), 2-hydroxy-20 -60 -diethylacetanilide (V), 2hydroxy-20 -60 -diethyl-N-methylacetanilide (VI), hydroxymetolachlor (X), metolachlor morpholinone (XII),
metolachlor propanol (XIII), deschloroacetylmetolachlor propanol (XV), hydroxyacetochlor (XVII) and 2hydroxy-20 -ethyl-60 -methylacetanilide (XX). These compounds were analyzed via HPLC-DAD instead of GCNPD; most possess a hydroxyl functional group that
results in tailing peaks (leading to increased detection
limits) when analyzed by GC.
Three 10-mL aliquots of each sample were successively extracted with 1 mL of dichloromethane for 1 min;
the three 1-mL extracts were then combined. The
dichloromethane was dried under a gentle stream of
nitrogen at ambient temperature and the sample was
reconstituted in 0.50 mL of 10 mM ammonium acetate
in acetonitrile containing 2,4-dichlorophenylacetic acid
as the internal standard. A 100-mL sample of each
extract was then injected onto a Waters (Milford, MA,
USA) HPLC-DAD (1525 pump and 2996 photodiode
array detector) set to an analytical wavelength of
210 nm. The mobile phase was 50% 10 mM aqueous
ammonium acetate and 50% acetonitrile, with a flow
rate of 1.0 mL/min. The analytical column was a
Phenomenex (Torrance, CA, USA) Luna C18 5 mm,
250 4.6 mm. The column temperature was set at 60 1C
using a Phenomenex TS-130 column heater. Extraction
efficiencies ranged from 93% to 98%, and detection
limits were 0.5 mg/L.
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2.6. Product identification (GC/MS)
Injections of 1 mL were made onto a ThermoQuest
(San Jose, CA, USA) Trace 2000 GC with a programmed temperature vaporization (PTV) injector
coupled to a quadrupole MS. A DB-5ms (Agilent)
30 m length 0.25 mm ID 0.25 mm film thickness
column was used to effect separations. The GC
temperature program was 90 1C for 1 min, 10 1C/min
to 290 1C and then held for 5 min. The PTV injector
was maintained at 200 1C. The mass spectrometer temperature was set to 250 1C, with an energy
of 70 eV, and spectra were obtained in electron
ionization (EI) mode. The transfer line was maintained
at 285 1C.
2.7. Kow estimation
Estimates of the log Kow value for each parent
compound and neutral degradate were determined using
ClogP version 2.0 (BioByte; Claremont, CA, USA).
ClogP is a program that estimates octanol/water
partition coefficients based on the fragment method of
Hansch and Leo (1979).
3. Results and discussion
3.1. Coagulation
Coagulation with either alum (30 mg/L) or ferric
chloride (20 mg/L) provided little removal (o10%) of
the compounds studied (Table 1). Removal of the parent
chloroacetamides ranged from 4% to 6% (fractional
removal was computed as the amount of the test analyte
remaining in solution after treatment divided by the
amount of test compound initially present). Some of the
degradates, such as deschloroacetylmetolachlor propanol (XV), exhibited removal efficiencies as high as 10%.
These results indicate that coagulation with alum and
ferric chloride is not an effective means of treating the
parent chloroacetamide herbicides or their neutral
degradates. Similar results for the parent herbicides
were obtained by other researchers; Miltner et al. (1989)
obtained a removal efficiency for alachlor of 4%
Table 1
Removal efficiencies of chloroacetamide herbicides and neutral degradates with alum and ferric chloride during coagulation
Analyte #
Identitya
Alumb % removal
FeCl3c % removal
I
II
III
IV
V
VI
VII
VIII
IX
X
XI
XII
XIII
XIV
XV
XVI
XVII
XVIII
XIX
XX
XXI
XXII
XXIII
XXIV
Alachlor
Hydroxyalachlor
Deschloroalachlor
2-chloro-20 -60 -diethylacetanilide
2-hydroxy-20 -60 -diethylacetanilide
2-hydroxy-20 -60 -diethyl-N-methylacetanilide
20 -60 -diethylacetanilide
2,6-diethylaniline
Metolachlor
Hydroxymetolachlor
Deschlorometolachlor
Metolachlor morpholinone
Metolachlor propanol
Deschloroacetylmetolachlor
Deschloroacetylmetolachlor propanol
Acetochlor
Hydroxyacetochlor
Deschloroacetochlor
2-chloro-20 -ethyl-60 -methylacetanilide
2-hydroxy-20 -ethyl-60 -methylacetanilide
20 -ethyl-60 -methylacetanilide
2-ethyl-6-methylaniline
Dimethenamid
Deschlorodimethenamid
4(71)
6(71)
4(71)
5(71)
5(72)
4(72)
4(71)
5(71)
5(72)
8(73)
4(71)
4(71)
3(71)
5(72)
10(73)
4(72)
6(72)
5(72)
3(71)
5(72)
4(72)
3(71)
4(71)
4(72)
4(71)
5(71)
5(72)
4(72)
7(71)
3(72)
3(71)
2(71)
6(71)
9(72)
5(71)
7(73)
6(73)
3(71)
10(72)
6(71)
8(72)
6(71)
5(71)
8(72)
5(71)
4(71)
6(71)
4(72)
Stated uncertainties represent one standard deviation based on triplicate analyses.
a
Initial concentration ¼ 50 mg/L.
b
Alum ¼ 30 mg/L.
c
FeCl3 ¼ 20 mg/L.
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(15 mg/L alum dose) and a removal efficiency for
metolachlor of 11% (30 mg/L alum dose).
3.2. Oxidation
3.2.1. Chlorination
Upon aqueous chlorination (6 mg/L applied free
chlorine for 6 h), 2,6-diethylaniline (VIII), deschloroacetylmetolachlor (XIV), deschloroacetylmetolachlor
propanol (XV), and 2-ethyl-6-methylaniline (XXII),
along with dimethenamid (XXIII) and its deschloro
degradate (XXIV), exhibited removal efficiencies of
84–100% (Table 2). Those degradates of acetochlor,
alachlor and metolachlor that lack the acetanilide
functional group displayed complete (100%) removal
under the conditions investigated. Those degradates
containing the acetanilide functional group, but lacking
the N-(alkoxy)alkyl side chain, such as XIX and XX,
yielded removals of 0–16%. Degradates that still
maintained some portion of both the acetamide and
N-(alkoxy)alkyl side chain, such as X and XI, generally
displayed removal efficiencies of 1–13%, and most
parent herbicides did not undergo detectable removal
in the presence of aqueous chlorine. These results
5039
suggest that the acetanilide functional group confers a
degree of protection against chlorination. Similar studies
by other researchers reveal that anilines react readily
with aqueous chlorine, while substituted amide compounds display little to no reaction with aqueous
chlorine (Katz, 1986; Hwang et al., 1990).
One parent chloroacetamide that reacted readily
(84% removal) with aqueous chlorine was dimethenamid (XXIII). Its neutral degradate, deschlorodimethenamid (XXIV), also underwent 96% removal in the
presence of aqueous chlorine. These compounds both
possess a substituted thienyl ring in place of the alkylsubstituted benzene ring of the other compounds
investigated (Fig. 1). Product studies undertaken for
these two compounds using GC/MS show that the
reaction with free chlorine leads to a single observed
product in each case, involving addition of a single
chlorine onto the thienyl ring (Fig. 2). GC-NPD
analyses of solvent extracts suggest this is a
major product (although we cannot preclude the
possibility of relatively polar products that were not
efficiently extracted by the solvent employed); if
we assume a similar response for the starting
materials and the chlorinated products, the molar
Table 2
Removal efficiencies of chloroacetamide herbicides and neutral degradates following application of free chlorine and ozone
Analyte #
Identitya
Chlorinationb % removal
Ozonationc % removal
I
II
III
IV
V
VI
VII
VIII
IX
X
XI
XII
XIII
XIV
XV
XVI
XVII
XVIII
XIX
XX
XXI
XXII
XXIII
XXIV
Alachlor
Hydroxyalachlor
Deschloroalachlor
2-chloro-20 -60 -diethylacetanilide
2-hydroxy-20 -60 -diethylacetanilide
2-hydroxy-20 -60 -diethyl-N-methylacetanilide
20 -60 -diethylacetanilide
2,6-diethylaniline
Metolachlor
Hydroxymetolachlor
Deschlorometolachlor
Metolachlor morpholinone
Metolachlor propanol
Deschloroacetylmetolachlor
Deschloroacetylmetolachlor propanol
Acetochlor
Hydroxyacetochlor
Deschloroacetochlor
2-chloro-20 -ethyl-60 -methylacetanilide
2-hydroxy-20 -ethyl-60 -methylacetanilide
20 -ethyl-60 -methylacetanilide
2-ethyl-6-methylaniline
Dimethenamid
Deschlorodimethenamid
1(72)
7(72)
2(71)
0(71)
13(74)
16(73)
11(72)
100
2(72)
8(74)
2(71)
13(74)
7(73)
100
100
1(72)
8(72)
1(71)
0(71)
13(72)
10(73)
100
84(72)
96(71)
63(71)
70(71)
66(72)
75(72)
77(71)
88(72)
80(71)
100
60(72)
67(71)
63(71)
85(72)
72(72)
100
100
61(72)
69(71)
64(71)
74(72)
78(72)
82(71)
100
100
100
Stated uncertainties represent one standard deviation based on triplicate analyses.
a
Initial concentration ¼ 50 mg/L.
b
Applied free chlorine (HOCl at pH 7); dose ¼ 6 mg/L; contact time ¼ 6 h.
c
Applied ozone; dose ¼ 3 mg/L; contact time ¼ 30 min.
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M.L. Hladik et al. / Water Research 39 (2005) 5033–5044
Relative Abundancee
5040
154
100
230
80
230
40
275
50
100
150
200
250
300
m/z
(a)
188
100
Relative Abundance
S
MW = 275
203
20
154
O
N
203
60
0
Cl
O
Cl
O
264
80
188
O
N
237
60
40
S
264
Cl
237
20
MW = 309
309
0
50
100
150
200
Relative Abundance
O
196
80
40
20
S
MW = 241
196
169
154
O
N
169
60
241
0
50
100
150
(c)
200
250
300
m/z
O
188
100
Relative Abundance
300
154
100
230
80
188
O
N
203
60
S
40
Cl
203 230
20
0
(d)
250
m/z
(b)
MW = 275
275
50
100
150
200
250
300
m/z
Fig. 2. Electron ionization mass spectra of: (a) dimethenamid; (b) observed dimethenamid product after chlorination; (c)
deschlorodimethenamid; and (d) observed deschlorodimethenamid product after chlorination. The products achieved upon
chlorination have an additional chlorine on the thienyl ring. Retention times were 14.8 min for dimethenamid, 16.0 min for the
dimethenamid chlorination product, 12.4 min for deschlorodimethenamid and 13.9 min for the deschloro-dimethenamid chlorination
product on a DB-5 ms column (see Section 2.6).
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M.L. Hladik et al. / Water Research 39 (2005) 5033–5044
3.2.2. Ozonation
All of the compounds reacted with ozone (3 mg/L
applied dose for 30 min), exhibiting removal efficiencies
in the range of 60–100% (Table 2). Those compounds
that were more susceptible to reaction with free chlorine
showed complete removal upon ozonation (100%).
Those compounds that did not react with the free
chlorine were partially transformed on contact with
ozone (removals of 60–88%).
Although all the test compounds exhibited substantial removal in the presence of ozone, no attempt
was made to determine whether removal resulted
from reaction with ozone or alternatively with OH
radicals, which are always present during ozonation.
These studies took place at near neutral pH, which
typically favors direct reactions with ozone (MWH,
2005); however, prior work with acetochlor and
metolachlor (Acero et al., 2003) has demonstrated these
compounds react slowly with ozone (1.1 and
2.4 M1 s1) relative to OH radicals (6.9–6.3 109 M1 s1).
The products formed following ozonation of the test
compounds were not identified. Prior research (Somich
et al., 1988) into ozonation of alachlor indicates the
benzene ring is cleaved, although complete mineralization does not occur. When anilines were reacted with
ozone, many degradates have been shown to form,
including azobenzenes, azoxybenzenes, and benzidines
(Chan and Larson, 1991). These aniline products are Nsubstituted polyaromatic compounds that are likely to
be more mutagenic than their parent compounds (Chan
and Larson, 1991).
The ozonation conditions tested resulted in higher
removal efficiencies for most of the compounds studied
relative to removal during chlorination (Table 2).
Complete mineralization is unlikely to occur under
ozonation and chlorination conditions encountered
at most water treatment plants. Additional studies
will be required to assess the extent to which oxidative treatment eliminates the human health risk
associated with the parent herbicides or their neutral
degradates.
3.3. Activated carbon
Adsorption data for each compound onto PAC were
fit to a Freundlich isotherm. The Freundlich isotherm is
an empirical correlation
qe ¼ KC 1=n
e ,
(1)
where qe is the sorption capacity (milligram adsorbate
per gram carbon); Ce the equilibrium solution concentration (mg/L); and K and n are constants. K is primarily
related to the capacity of the adsorbent and n is related
to the strength of adsorption (Snoeyink, 1990). Example
plots of qe versus Ce are shown in Fig. 3. Such data
(when transformed to log–log plots) were used to
determine 1/n (slope) and K (determined from qe value
where Ce ¼ 1).
Values for K and 1/n obtained for each compound are
listed in Table 3. The environmental degradates and
parent herbicides vary considerably in their affinity for
PAC; the parents have a higher adsorption affinity than
all of their degradates, with the exception of the
morpholinone derivative of metolachlor (XII). As the
data in Fig. 3 show, loss of an alkoxyalkyl side chain, or
replacement of –Cl by –OH, results in degradates with a
lesser affinity for PAC. When the log K values are
compared with the estimated log Kow values, a reasonable correlation ðR2 ¼ 0:87Þ results (Fig. 4). Those
compounds with lower estimated Kow values tend to
have lower adsorption affinities for the PAC under test
conditions, implying that those compounds that are
more water-soluble sorb less to the activated carbon.
These lower adsorption affinities for the neutral
degradates imply less removal of these compounds is
likely than for the parent herbicide at a given PAC
loading. Similar results were observed by Gustafson
O
70
Cl
N
60
O
50
qe (mg/g)
concentrations of the respective chlorination products
are approximately equal to the concentration of the
parent compounds lost during reaction with aqueous
chlorine.
Product studies for the other compounds that reacted
with aqueous chlorine were attempted, but we were
unable to identify any products using GC/MS. Relatively polar products may well be formed but would not
be amenable to GC or GC/MS analysis using the
techniques employed. Previous studies of the chlorination of substituted anilines have found the products
involved chlorine addition to the aniline ring (Hwang
et al., 1990).
5041
Cl
40
HN
O
30
OH
20
HN
O
10
0
0.00
0.01
0.02
0.03
0.04
0.05
Ce (mg/L)
Fig. 3. Equilibrium data showing adsorption onto PAC of: (K)
metolachlor; (.) 2-chloro-20 -ethyl-60 -methyl acetanilide; and
(’) 2-hydroxy-20 -ethyl-60 -methyl acetanilide. Solid lines represent Freundlich isotherm fit.
ARTICLE IN PRESS
M.L. Hladik et al. / Water Research 39 (2005) 5033–5044
5042
Table 3
Freundlich parameters for adsorption of chloroacetamide herbicides and their neutral degradates onto PAC
Analyte #
Identitya
Kb
Kb range
1/n
1/n range
Kowc
I
II
III
IV
V
VI
VII
VIII
IX
X
XI
XII
XIII
XIV
XV
XVI
XVII
XVIII
XIX
XX
XXI
XXII
XXIII
XXIV
XXV
Alachlor
Hydroxyalachlor
Deschloroalachlor
2-chloro-20 -60 -diethylacetanilide
2-hydroxy-20 -60 -diethylacetanilide
2-hydroxy-20 -60 -diethyl-N-methylacetanilide
20 -60 -diethylacetanilide
2,6-diethylaniline
Metolachlor
Hydroxymetolachlor
Deschlorometolachlor
Metolachlor morpholinone
Metolachlor propanol
Deschloroacetylmetolachlor
Deschloroacetylmetolachlor propanol
Acetochlor
Hydroxyacetochlor
Deschloroacetochlor
2-chloro-20 -ethyl-60 -methylacetanilide
2-hydroxy-20 -ethyl-60 -methylacetanilide
20 -ethyl-60 -methylacetanilide
2-ethyl-6-methylaniline
Dimethenamid
Deschlorodimethenamid
Atrazine
266
127
131
117
49
105
63
121
304
139
156
341
121
106
53
282
134
123
85
30
48
89
129
89
131
169–419
92–176
118–145
117–129
39–60
81–135
48–83
104–141
213–433
104–186
133–182
235–493
97–151
80–140
41–68
181–439
99–181
96–159
70–103
25–36
38–61
64–124
88–189
80–98
96–179
0.48
0.43
0.46
0.44
0.38
0.43
0.41
0.42
0.50
0.44
0.50
0.52
0.44
0.42
0.38
0.48
0.44
0.46
0.43
0.35
0.39
0.43
0.43
0.42
0.48
0.41–0.55
0.37–0.49
0.43–0.49
0.42–0.46
0.33–0.43
0.38–0.48
0.35–0.47
0.39–0.45
0.44–0.56
0.39–0.49
0.47–0.53
0.46–0.58
0.39–0.49
0.36–0.48
0.32–0.44
0.41–0.55
0.38–0.50
0.41–0.51
0.39–0.47
0.30–0.40
0.34–0.44
0.36–0.50
0.36–0.50
0.40–0.44
0.42–0.54
1500
560
1200
680
35
690
170
740
1800
650
1400
3800
430
170
39
1500
560
1200
200
10
49
220
300
220
250
Stated uncertainties in K and 1/n indicate 95% confidence intervals based on linearized data.
a
Initial concentration
¼.
50 mg/L.
b
Units of K are mg=g ðmg=LÞ1=n :
c
Kow estimated using ClogP.
log K (Freundlich Parameter)
3.0
log K = 0.389(±0.032)(log Kow) + 1.050(±0.084)
R2 = 0.872
2.5
2.0
1.5
1.0
1.0
1.5
2.0
2.5
3.0
3.5
4.0
log Kow (estimated using ClogP)
Fig. 4. Comparison of log K (Freundlich parameter), determined from PAC adsorption studies versus log Kow estimated
using ClogP. Error bars represent 95% confidence intervals.
et al. (2003) with the ESA and OA degradates, which are
comparatively more water soluble than the parent
herbicides, and are less readily removed by sorption
onto PAC.
Adsorption by PAC was also explored with atrazine
(XXV) as a sorbate (Table 3), as many drinking water
systems optimize their PAC addition to achieve a target
level for atrazine. For example, if atrazine is assumed to
have an influent concentration of 5 mg/L and an effluent
concentration of 2.5 mg/L (slightly below the MCL of
3 mg/L) is desired, our results would yield a recommended PAC dosage of 0.3 mg/L. The parent chloroacetamides have higher K values than atrazine, resulting
in greater predicted removal efficiencies under these
conditions. Many of the neutral chloroacetamide
degradates have lower K values than atrazine. Given
the same influent and desired effluent concentrations, 2chloro-20 -ethyl-60 -methylacetanilide (XIX) would require
0.4 mg/L of PAC and 2-hydroxy-20 -ethyl-60 -methylacetanilide (XX) would require 0.7 mg/L of PAC. These
values are 1.1 times and 1.9 times the PAC dose needed
for comparable control of atrazine.
4. Conclusions
Coagulation with alum and ferric chloride is ineffective at removing chloroacetamide herbicides and their
ARTICLE IN PRESS
M.L. Hladik et al. / Water Research 39 (2005) 5033–5044
degradates. Chlorination with an applied dose of 6 mg/L
for 6 h removes those degradates of acetochlor, alachlor
and metolachlor that lack the acetanilide functional
group, but resulting products are unknown. Dimethenamid (XXIII) and its deschloro degradate (XXIV)
reacted with chlorine to give a single product in each
case, involving addition of chlorine to the thienyl ring.
Ozonation with an applied dose of 3 mg/L for 30 min
proved more successful than chlorination in terms of
removal of all of the compounds, including the parent
herbicides. All of the test compounds were adsorbed by
PAC. Neutral chloroacetamide degradates, which generally have lower Freundlich K values than the parent
compounds or atrazine are likely to require higher PAC
doses to achieve removals comparable to those desired
for atrazine. We note that concentrations of the oxidant
used for chlorination and ozonation experiments are
representative of the upper limits used in drinking water
treatment facilities and that the test compound concentrations are higher than typically reported in source
waters; observed removal percentages may not, therefore, apply directly to the performance encountered
during actual drinking water treatment.
Acknowledgements
We are grateful for the comments made by two
anonymous reviewers. Funding was provided by the
American Water Works Association Research Foundation/US Environmental Protection Agency (contract
02903). The Johns Hopkins University gratefully
acknowledges that the AWWA Research Foundation
is the joint owner of the technical information upon
which this publication is based. The Johns Hopkins
University thanks the Foundation and the US government, through the Environmental Protection Agency,
for its financial, technical, and administrative assistance
in funding and managing the project through which the
information was discovered. The comments and views
detailed herein may not necessarily reflect the views of
the AWWA Research Foundation, its officers, directors,
affiliates or agents, or the views of the US Federal
Government.
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