Full text of DOI:10.1089/109287503768335940

ENVIRONMENTAL ENGINEERING SCIENCE
Volume 20, Number 5, 2003
© Mary Ann Liebert, Inc.
Occurrence and Behavior of Alkylphenol Polyethoxylates in
the Environment
John Montgomery-Brown and Martin Reinhard*
Department of Civil and Environmental Engineering
Stanford University
Stanford, CA 94305-4020
ABSTRACT
Alkylphenol polyethoxylates (APEOs) are nonionic surfactants whose degradation metabolites have be-
come ubiquitous in the aquatic environment. The environmental significance of APEO metabolites and
the threat they pose to wildlife is still a matter of debate. Even though advanced analytical procedures are
available, researchers have been unable to obtain a complete mass balance during biodegradation studies.
The ultimate fate of APEOs and their metabolites is not adequately understood. Biodegradation is be-
lieved to be the dominant attenuation process, but photodegradation may also play an important part. The
synthesis of single isomer standards for APEOs and their metabolites is necessary for accurate quantifi-
cation, estrogenicity experiments, and biological and chemical fate studies.
Key words: alkylphenol polyethoxylates; environment
alkyl chain connected to a benzene ring (Fig. 1). The
alkylphenol (AP) is formed through the acid-catalyzed
INTRODUCTION
LKYLPHENOL POLYETHOXYLATES (APEOs) are non- Friedel-Crafts alkylation of phenol with an alkene. The
para
or-
ionic surfactants that have been used for more than phenol is preferentially alkylated at the
position;
A
tho
40 years in household and industrial detergents, petro-
isomers make up less than 10% of a technical APEO
leum refining, pulp and paper production, crop protec- mixture (Talmage, 1994). Octyl and nonyl groups are
tion chemicals, and plastics and textiles manufacturing made from diisobutylene and tripropylene, respectively;
(Talmage, 1994). The worldwide demand for APEOs is octyl groups are predominantly 1,1,3,3-tetramethylbutyl,
greater than 500 million kg/year (Naylor, 1995); of this while nonyl groups are mixtures of branched isomers.
total, approximately 80% are nonylphenol polyethoxy- The AP is then reacted with ethylene oxide; the result-
lates (NPEOs) and 15–20% are octylphenol polyethoxy- ing product is a mixture of APEOs having various poly-
et al
lates (OPEOs) (Staples
, 1999).
oxyethylene chain lengths (Talmage, 1994). In this arti-
APEOs are composed of a polyethoxy chain and an cle, APnEO will be used to represent an alkylphenol
Phone
: 650-723-
*Corresponding author: Civil & Environmental Engineering, Stanford University, Stanford, CA 94305-4020;
Fax E-mail
0308;
: 650-723-7058; : reinhard@ce.stanford.edu
471

472
MONTGOMERY-BROWN AND REINHARD
view will focus on the advances and current trends in
these areas.
ENVIRONMENTAL SIGNIFICANCE
Figure 1. Structure of an alkylphenol polyethoxylate.
Indigenous fish exhibiting symptoms of endocrine dis-
ruption can be found in a number of aquatic environments
et al
., 1998). For example, a report by Harries
. (1997) found that fish caught at a wastewater treat-
(Jobling
et al
n
polyethoxylate with
ethoxy units; most commonly,
APEOs have between 1 and 20 ethoxy (EO) units. The
analytical detection, identification, and quantification of
APEOs and their metabolites has been hindered by a lack
of standards; even the available standards are mixtures
of various structural isomers. For example, using high-
resolution gas chromatography-mass spectrometry (GC-
ment plant (WWTP) outfall had abnormal ratios of sex
steroid hormones and exhibited histological changes in
their reproductive organs. Another study was able to at-
tribute the feminization of male reproductive organs in
et al
.,
fish to WWTP effluent exposure (Rodgers-Gray
2001). Natural estrogens play important roles in the de-
velopment of male and female reproductive organs and
in metabolic regulation, even at very low concentrations
(Servos, 1999). Therefore, they, and their synthetic ana-
logues, might be responsible for the observed endocrine
disruption in fish. Another possibility may be estrogen
mimics (i.e., compounds able to elicit estrogenic re-
sponses in test organisms). In 1991, while culturing hu-
et al
. (1997) determined that there were
MS), Wheeler
22 distinct
para
-nonylphenol (NP) isomers in a techni-
cal NP mixture.
During biological treatment, APEO metabolites like
n 5
short-chained APnEOs (
1–3), alkylphenol poly-
ethoxycarboxylates (APECs), carboxyalkylphenol poly-
ethoxycarboxylates (CAPECs), and APs are produced.
Halogenated APEOs and APEO metabolites can also be
et al.
formed during wastewater chlorination (Reinhard
et al
,
1982; Ball, 1985; Ding
., 1996). Structures for sev-
eral of the aforementioned metabolites are shown in Fig.
2. APEO metabolites were among the most commonly
detected wastewater-derived contaminants detected by
et al
., 2002)
the United States Geological Survey (Kolpin
during a national survey of lakes, rivers, and streams in
the United States. Because these metabolites are rela-
et al
tively recalcitrant in the environment (Fujita
et al. et al.
, 1999), can bind to
., 1996;
DiCorcia
, 1998; Shang
et al
and accumulate in soils and sediments (Khim
., 1999;
et al.
, 2001), can bioaccumulate in plants and
Ferguson
animals (Ahel
et al. et al
, 1993; Staples
be estrogenic to wildlife at low concentrations (Jobling
et al. et al.
., 1998), and can
, 1996; Routledge , 1998), their ubiquitouspres-
ence in the environment may be cause for concern.
In 1999, the Water Quality Research Journal of Canada
featured several comprehensive reviews on the occur-
rence, persistence, and aquatic toxicity of APEOs and
their metabolites in the environment (Bennie, 1999;
Maguire, 1999; Servos, 1999); the analytical techniques
used in the detection of these compounds were also ex-
tensively reviewed (Lee, 1999). A more recent review
covered the APEO removal in activated sludges (John-
son and Sumpter, 2001). Because recent research has
focused on (1) determining the estrogenic potency of
in vivo
, (2) developing analytical techniques
APEOs
capable of detecting both neutral and acidic APEO
metabolites, and (3) obtaining a better understanding of
degradation pathways in environmental matrices, this re-
Figure 2. Potential structures of several APEO metabolites
(NPEC, CA8PEC, NP, and a halogenated NP2EC, respec-
tively).

ALKYLPHENOL POLYETHOXYLATES IN THE ENVIRONMENT
473
et al
man breast cancer cells, Soto
para
. (1991) discovered that
b
Because all vertebrates produce 17 -estradiol, and it
is very potent at low concentrations(ng/L), this hormone
was selected as the basis for determining the estrogenic-
-nonylphenol was a weak estrogen. Because of the
structural similarities between some nonylphenolisomers
and natural and synthetic estrogens (Fig. 3), it is likely
In vitro
ity of xeno- and synthetic estrogens.
tests indi-
b
that nonylphenol can compete with 17 -estradiol for the
b
cate that, compared to 17 -estradiol, APEOs and their
metabolites are weak xenoestrogens; for example,
estrogen binding receptor in vertebrates. Thus, it seems
reasonable that APEOs and their metabolites could also
be weak estrogens. Furthermore, the high concentrations
et al
Jobling
metabolites were 1,000 to 10,000 times less than that of
in vivo
. (1996) found that the potencies of APEO
m
of APEOs and their metabolites ( g/L) in WWTP efflu-
b
17 -estradiol. However,
experiments have found
ents relative to other estrogens and estrogen mimics
(ng/L), suggests that these compounds could play an im-
portant role in the observed endocrine disruption in fish.
b
that OP is only 100 to 1,000 less effective than 17 -estra-
diol; the increased potency of OP was postulated to re-
et al
sult from bioaccumulation (Routledge
et al
., 1998;
in vivo
in vitro
in vivo
tests, it was shown
Through various
that APEO metabolites like APs and short-chained
n 5
and
Thorpe
., 2000). Conclusions based on
ex-
in vitro
periments may be more relevant than those from
APnEOs (
low microgram per liter concentrations(White
et al.
1–3) are active environmental estrogens at
experiments because living organisms may be better able
to inactive or excrete natural plant and animal estrogens
than xenoestrogens like APEO metabolites (Johnson and
et al
., 1994;
Jobling
, 1996). To compound the problem, the es-
trogenic responses to these xenoestrogens may be additive
et al
in vivo
Sumpter, 2001). However,
estrogenic responses
(Sumpter and Jobling, 1995; Routledge
trogenicitytests are commonly used to determine the over-
et al.
, 1998). Es-
are affected by the organisms being used, their sex (i.e.,
males are more sensitive than females), and their devel-
opmental stage (i.e., developing organisms are more sen-
all estrogenic potency of water samples (Khim
et al.
,
2001b; Korner
, 2001). Of all APEO metabolites,
et al
et al
, 1996; Routledge .,
sitive than adults) (Jobling
1998).
octylphenol(OP) is the most potent xenoestrogen,and can
induce significant effects in fish at concentrations of 3
In general, the APEO metabolite concentrations ob-
served in the environment (Table 1) are below the levels
known to cause estrogenic effects in fish and other
wildlife. However, OP and NP formation from the degra-
dation of other APEO metabolites (e.g., APECs) could
result in OP and NP concentrations capable of inducing
estrogenic effects. Furthermore, because several APEO
metabolites can accumulate and persist in the environ-
ment, and bioaccumulate in wildlife, it is difficult to as-
sess what concentrations pose a significant risk for en-
docrine disruption.
m
et al
g/L (Jobling
., 1996); technical-grade nonylphenol
et al.
m
(NP) is active at concentrations of 8.3 g/L (Harris
,
m
2001). At concentrationsbelow 100 g/L, APECs and AP-
n .
nEOs (
2) do not appear to elicit estrogenic responses
et al et al
in fish (Metcalfe
., 2001; Nichols ., 2001); estro-
genic studies on CAPECs have yet to be performed. The
lack of standards for APEO metabolites (especially indi-
vidual isomers) has prevented researchers from accurately
assessing structure–activity relations and the impact of
these chemicals in the environment.
Figure 3. Structural similarities between two NPEO metabolites and 17-b-Estradiol (a human estrogen) and Diethylstilbestrol.
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003

474
MONTGOMERY-BROWN AND REINHARD
Table 1. Occurence of APEOs and APEO metabolites in the environment.
Location
Compound
Concentration
Reference
1° WWTP effluenta
2° WWTP effluenta
WWTP effluentb
NP(1–18)EO
NP(1–18)EO
NP(1–2)EO
NP
OP
NP1EO
NP
CAPEC
NP(1-2)EO
NPEC
CAPEC
NP2EO
NP1EC
NP
96–430 mg/L
2–8 mg/L
Ahel
. (2000)
et al
et al.
(2001)
0.4–8.9 mg/L
1.2–2.7 mg/L
30–190 ng/L
45 6 16 mg/L
3.00 6 0.85 mg/L
58 mg/L
1.7–6.6 mg/L
0.6–15 mg/L
6.31–40.56 mg/L
bd–5.5 mg/L
0.17–5.8 mg/L
0.33–2.3 mg/L
111 ng/L
Rodgers-Gray
WWTP effluentb
Lye
(1999)
et al.
WWTP effluentc
WWTP effluentc
DiCorcia
DiCorcia
. (1998)
. (2000)
et al
et al
WWTP effluentd
Spengler
(2001)
et al.
WWTP effluentd
NP
OP
NP
NP
Kuch and Ballschmiter (2001)
et al
14 ng/L
WWTP effluente
WWTP Sedimentse
Septagee
, 1.0–33 mg/L
, 5–12,400 mg/L dw
. 1,000 mg/L
Hale
. (2000)
et al.
(1998)
NP
Rudel
Land-applied sewage sludgee
NP(1–2)EO
NP
OP
, 1.5–408 mg/kg dw
5.4–887 mg/kg dw
, 0.5–12.6 mg/kg dw
27–113 mg/g dw
La Guardia
(2001)
et al.
Sewage sludgee
NP(1–4)EC
Field and Reed (1999)
Rivers, streams, and lakese
NP(1–2)EO
NP
OP(1–2)EO
NP(2–8)EO
NP
NP1EO
NP
OP
, 1–20 mg/L
, 0.5–40 mg/L
, 0.2–2 mg/L
2–35 nM
Kolpin
et al
. (2002)
River waterf
Maruyama
(2000)
et al.
River waterd
6–135 ng/L
Kuch and Ballschmiter (2001)
et al
River sedimentsb
160–3970 ng/g dw
30–9050 ng/g dw
2–340 ng/g dw
Lye
. (1999)
Estuary watere
NP(1–3)EO
NP
OP (1–3)EO
OP
NP(1–3)EO
OP(1–3)EO
697 ng/L
201 ng/L
31.9 ng/L
3.27 ng/L
0.22–1.05 mg/L
7–40 ng/L
Ferguson
. (2000)
. (2001)
et al
et al
Estuary watere
Ferguson
Ferguson
Estuary sedimentse
NP(1–3)EO
NP
Cl- and Br-NP
OP(1–3)EO
OP
NP(1–3)EO
Cl- and Br-NP
OP(1–3)EO
NP
3.15 mg/L
846 ng/g dw
13.8 ng/g dw
79.7 ng/L
8.11 ng/L
0.05–30 mg/g dw
nd–45.6 ng/g dw
5–90 ng/g dw
113–3890 ng/g dw
4–179 ng/g dw
, 1–19.9 ng/g dw
, 1–2.91 ng/g dw
1.5 mg/g dw
(2000)
(2001)
et al.
et al.
Estuary sedimentse
Ferguson
Estuary sedimentsg
Estuary sedimentsg
Marine sedimentsh
Khim
Khim
Shang
(1999)
et al.
OP
NP
OP
et al
. (2001a)
et al.
NP(1–19)EO
(1999)

ALKYLPHENOL POLYETHOXYLATES IN THE ENVIRONMENT
475
Table 1. (Cont’d)
Location
Compound
NP
OPEO
OP
NP1EO
NP
Concentration
Reference
et al.
Edible mollusksi
67–696 ng/g fw
nd–0.43 ng/g fw
2.7–18.6 ng/g fw
190–940 ng/g fw
5–55 ng/g fw
Ferrara
(2001)
Fish tissueb;
Lye
. (1999)
et al
Platichthys flesus
Mature
Drinking waterd
Foodd
NP
OP
NP
2–15 ng/L
0.15–5 ng/L
0.1–19.4 mg/kg fw
Kuch and Ballschmiter (2001)
et al
Guenther . (2002)
dw, dry weight; fw, fresh weight; bd, below detection;nd, not detected.
aSwitzerland; bUnited Kingdom; cItaly; dGermany; eUnited States; fJapan; gKorea; hCanada; iAdriatic Sea.
ration of APEOs and their metabolites. The advantages
of liquid chromatography include its versatility and its
ability to analyze nonvolatile, high molecular weight
samples without derivatization; its main drawback is its
low resolution. For example, normal phase (silica and
aminopropyl-coated silica solid phases) LC techniques
can only separate APEOs based on the length of their
ethoxylate chain. Reverse phase (C8- and C18-coated sil-
ica solid phases) can only separate APEOs based on the
length and structure of their alkyl chain. Therefore, nor-
mal-phase LC cannot distinguish between OP1EO and
NP1EO, and reverse-phase LC cannot distinguish be-
tween NP1EO and NP10EO. Recently, however, Gun-
derson (2001) developed a reverse-phase technique, us-
ing graphitized carbon black, which could resolve the 22
isomers found in technical grade NP into 12 groups. This
method could also differentiate between the different iso-
mers found in NPEO mixtures, but more work is neces-
sary to achieve better separation of different ethoxylate
chain lengths.
ANALYTICAL METHODS
Sample preparation
Water
. For water-based samples, the concentration
technique most often used is reverse-phase solid-phase
et al
extraction (SPE) (Ding and Tzing, 1998; DiCorcia
.,
., 2001). For reverse-phase SPE,
columns are most often filled with C18-coated silica par-
et al
2000; Ferguson
et al
et
ticles (Jonkers
., 2001; Schroder, 2001; Spengler
al
et al
, 2000;
., 2001), graphitized carbon black (Ahel
et al
DiCorcia
., 2000), or hydrophobic polymers (e.g.,
et al
XAD-16) (Maruyama
., 2000). Differential elution
from the SPE cartridge is often used for the separation
and enrichment of acidic and neutral compounds (Di-
et al
et al
., 2000; Castillo , 2001). Acidic metabo-
Corcia
lites can be separated from the reverse-phase SPE eluate
et al
by silica gel chromatography (Spengler
., 2001;
2001). Typical SPE recoveries range from
et al.,
Korner
78 to 95%. Other techniques used to concentrate APEOs
and their metabolites from water are evaporation (Ding
Fluorescence detectors are increasingly being used for
the detection of APEOs and their metabolites because of
their higher selectivity and sensitivity than UV diode ar-
ray detectors. However, as with UV adsorption, the rel-
ative response factors for fluorescence change consider-
ably with molecular weight; therefore, it is necessary to
know the oligomeric composition of the mixture when
determining total NPEO concentrations using fluores-
et al
1999; Ahel
et al
.,
., 1996) and liquid–liquid extraction (Potter
et al
., 2000).
Soils and sediments
. Extraction of APEOs and their
metabolites from soils, sediments, and sludges can be ac-
et al.
complished using Soxlet extraction (Khim
, 1999),
enhanced solvent extraction (Field and Reed, 1999;
et al. et al.
Shang
, 1999; Ferguson
, 2000), and steam dis-
et al.
, 1999). Recoveries range from 60
et al
cence spectroscopy (Ahel
and emission wavelengths most commonly used are
et al
., 2000). The excitation
tillation (Tanghe
.
to 95%. Aminopropyl and cyanopropyl normal phase
,225 and ,305 nm, respectively (Tanghe
., 1999;
., 2000; Gundersen, 2001). The UV absorption
wavelengths most commonly used are ,230 and ,277
et al et al.
SPE columns have been used to clean up sediment ex-
tracts (Shang
et al
Ahel
et al
et al
., 2000).
., 1999; Ferguson
nm (Brand
., 1998; Potter
, 1999).
Analytical separation and detection
For LC mass spectrometry (MS), reverse-phase C8 and
C18 analytical columns are most frequently used (Fer-
Liquid chromatography
. Several liquid chromatogra-
phy (LC) techniques have been developed for the sepa-
et al
., 2000; Jonkers
et al
., 2001; Schroder, 2001).
guson
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003

476
MONTGOMERY-BROWN AND REINHARD
Mass spectrometry
The two most common ionization techniques used for
. Triple-stage quadrupole (Di-
et al
et al
APEOs and their metabolites are atmospheric pressure Corcia
., 2000; Jonkers
., 2001) and ion-trap
et al
et al.
, 2001; (Ding and Tzing, 1998; Tanghe
chemical ionization (APCI) (Castillo
., 1999) mass spec-
Schroder, 2001) and electrospray ionization (ESI) (Di- trometers are becoming more common because they
et al et al. et al
Corcia
., 2000; Ferguson
, 2001; Jonkers
., provide more detailed information than single-stage
2001). ESI is more popular because of its wide applica- quadrupole mass spectrometers. APEO metabolites are
bility to large and small, polar and nonpolar compounds, identified using full-scan spectra and quantified using
as well as singly and multiply charged species. For ESI single or multiple ion monitoring. Quantification of
LC-MS, parent APEOs are generally detected by moni- APEOs and their metabolites by MS is difficult be-
toring the positive ions produced during ionization while cause of the lack of standards. Even when standards
APEO metabolites are detected by monitoring the nega- are available (as for NP), quantification is still diffi-
et al
tive ions produced during ionization (DiCorcia
et al. et al.
., cult because each NP isomer may have a different re-
sponse factor. Because the response factors for AP1EC
and AP2EC are similar, most researchers assume that
2000; Ferguson
, 2000; Jonkers , 2001).
Gas chromatography
. Capillary column GC has been the response factors for other metabolites (e.g.,
et al
used for the analysis of underivatized and derivatized CAPECs) will not differ significantly (DiCorcia
.,
et al
APEOs and their metabolites (Ding
., 1996; Wheeler 2000; Schroder, 2001).
., 2001). Despite its higher res-
et al et al
., 1997; Spengler
olution, and its ability to distinguish between alkyl chain
isomers and ethoxylatechain length, GC is becoming less
n 5
common as it is limited to short-chained(
1–4) APEO
SOURCES OF APEOS AND
APEO METABOLITES
metabolitesbecause of their low volatility. Derivatization
is frequently used to increase the volatilityof APECs and
CAPECs; it can also reduce matrix effects, enhance chro-
Approximately 55% of the annual APEO production
matographicseparation, and facilitate mass spectral iden- is used in industry (e.g., plastics, textiles, and paper man-
tification. APECs and CAPECs are most commonly de- ufacturing), 30% for industrial cleaning products, and the
et al
rivatized to their methyl (Ball
., 1989; Field and remaining 15% for consumer products (Talmage, 1994).
Reed, 1999), propyl (Ding and Tzing, 1998; Wild and Thus, approximately 200 million kilograms of APEOs
Reinhard, 1999), and pentafluorobenzoyl (Kuch and enter WWTPs per year. APEO removal in WWTPs varies
. et al.
widely (60 to 98%; Thiele , 1997) because of the
Ballschmiter, 2001) esters.
GCs may be used in conjunction with flame ionization different source waters, operating conditions, and treat-
et al. et al.
detectors (Wheeler
, 1997; Guenther
, 2002), ment technologies used. In a study using Swiss WWTPs,
et al
electron capture detectors (ECD) (Kuch and Balls- Ahel
. (1994a) determined that approximately 40%
chmiter, 2001), and mass spectrometers (Ding and Tz- of the incoming APEO mass is mineralized; the remain-
et al.
ing, 1998; Ejlertsson
, 1999; Field and Reed, 1999). der was biotransformed to a mixture of metabolites (de-
For GC-MS, both electron ionization (EI) (Ding and Tz- spite the recent discovery of CAPECs, these percentages
et al
ing, 1998; Spengler
., 2001) and chemical ionization are likely still reasonable because any CAPECs present
(CI) (Field and Reed, 1999; Kuch and Ballschmiter, in the effluent would have been quantified along with the
2001) are used. Because APEO molecules and their APEC metabolites).
metabolites fragment readily, it is often difficult to iden-
tify the molecular ion using EI. CI results in less frag- discharged into the environment via sewage sludge and
et al.
These hydrophobic and hydrophilic metabolites are
mentation and greater molecular ion abundances, and is wastewater effluents (Ahel
, 1994a). Although mu-
hence, very useful when used in conjunction with EI for nicipal and industrial WWTP effluents have traditionally
the identification of APEOs and their metabolites. CI been considered the main sources of APEO metabolites
techniquesmost commonly use methane as a reagent gas; in the environment, recent research suggests that other
et al
however, ammonia is very useful for APEOs and their sources may also be important. For example, Hale
.
1
metabolites because it gives very intense [M1-NH4
]
. m
(2000), found high concentrations ( 100 g/kg) of sed-
peaks (Field and Reed, 1999). Positiveions are most com- iment-bound APEO metabolites near storm water dis-
monly detected in both EI and CI GC-MS (Ding and Tz- charges, civilian and military shipyards, and other gov-
ing, 1998; Field and Reed, 1999). In CI GC-MS, nega- ernment and industrial facilities. Septic systems may also
tive ion detection is occasionally used to obtain very low be significant sources of APEO metabolites; all septage
et al
et al
., 1999; Kuch and sampled by Rudel . (1998) had NP concentrations
detection limits (Ejlertsson
Ballschmiter, 2001).
m
greater than 1,000 g/L.

ALKYLPHENOL POLYETHOXYLATES IN THE ENVIRONMENT
477
and riverbed sediments may have accumulated large
amounts of APEO metabolites, and could serve as reser-
voirs for these compounds.
OCCURRENCE DATA
Table 1 summarizes the observed concentrations of
APEOs and their metabolites in the environment since
1999; for a summary of occurrence data prior to 1999,
see Bennie (1999).
In sediments and sludges, it is important to analyze for
n .
long-chained APnEOs (
3) in addition to short-
1–3). For example, more than half
et al
n 5
chained APnEOs (
of the sediments analyzed by Shang
. (1999) had
NPEOs with more than two ethoxy units; this lead these
researchers to conclude that analyses, which only focus
on short ethoxy chain oligomers in sediments or sludges,
may underreport total NPEO concentrations by a factor
of 2.
Water
APEO metabolites have become almost ubiquitous in
the aquatic environment due to their extensive industrial
and domestic use. For example, during a national recog-
nizance of streams, lakes and rivers in the United States,
the United States Geological Survey found that APEO
metabolites (NP, NP1EO, NP2EO, OP1EO, and OP2EO)
were detected approximately 70% of the time and ac-
counted for approximately 36% of the total concentration
of 95 different antibiotics, pharmaceuticals, nonprescrip-
tion drugs, steroids and hormones, and wastewater-re-
DEGRADATION IN THE ENVIRONMENT
Research on the physiochemical properties of APEOs
and their metabolitesand the impact these properties have
on the environmental fate of these compounds is limited.
Due to the lack of standards, even less is known about
APECs and CAPECs. Table 2 summarizes the observed
and estimated physical properties for several APEO
metabolites. Based on these properties, volatilization
et al
lated compounds (Kolpin
., 2002); if acidic APEO
metaboliteswere included,the frequency of detectionand
total concentration would likely have been significantly
higher. APEOs and their metabolites have also been de-
et al
tected in estuaries (Ferguson
., 2001), groundwater
, 1996), and even drinking water (Kuch and
et al
(Dachs
., 1999) and sorption to suspended solids and
et al.
(Ding
sediments may be important physicochemical removal
methods for some metabolites (APs and short-chained
APnEOs). Other removal mechanisms include chemical
and biologicalprocesses; chemical removal is dominated
by photochemical degradation and biological removal is
accomplished by APEO uptake and transformation by
micro-organisms (Maguire, 1999).
Ballschmiter, 2001).
APECs and CAPECs are frequently the most abundant
APEO metabolites present in WWTP effluents. For ex-
et al
ample, DiCorcia
. (2000) determined that CAPECs
accounted for approximately 65% of the APEO metabo-
lites in Italian WWTP effluents; the concentrations of
these metabolites ranged between 6.31 and 40.56 g/L.
m
Few researchers have examined the occurrence and dis-
et al
.,
tribution of halogenated APEO metabolites (Ding
et al et al.
, 2001).
Biodegradation
1996; Ferguson
., 2001; Petrovic
APEO biodegradation is believed to be the dominant
removal process for these compounds in the environ-
ment. Despite the fact that the degradation of these com-
Sewage sludges and sediments
In WWTPs, hydrophobic APEO metabolites tend to pounds has been studied for 40 years, the degradation
et al
sorb to sewage sludges. Tanghe
. (1998) found that mechanisms for certain APEO metabolites remain un-
between 60 and 90% of spiked NP, a hydrophobic (log known. In fact, one group of metabolites (CAPECs) was
KOW et al
, 4.2) APEO metabolite, sorbs to 3–4 g/L of not even discovered until 1996 (Ding
sewage sludge. Even relatively hydrophilic metabolites thermore, although there is no disagreement about the
et
., 1996). Fur-
like APECs have been found in sewage sludges (Ahel
degradability of long-chained APEOs, experimental ev-
., 1994a; Field and Reed, 1999). Interestingly, approx- idence for the degradation of their metabolites is in-
imately 45% of the NPECs detected by Field and Reed consistent—some experiments indicate that APEO
ortho ortho et al
al
(1999) were
isomers are more recalcitrant than
Sorbed APEO metabolites have also been found in es- that they are very recalcitrant (DiCorcia
et al
isomers. This may indicate that
metabolites are readily biodegradable (Staples
et al
1999; Montgomery-Brown
.,
., 2002), others indicate
et al
para
isomers.
., 1998;
tuarine, lacustrine, and riverbed sediments (Table 1). In Shang
., 1999).
anoxic marine sediments from the Strait of Georgia,
It is generally accepted that APEO biodegradation be-
et al
Canada, Shang
. (1999) found very little biodegra- gins by the rapid stepwise shortening of the ethoxy chain
dation of NPEOs; half-lives were estimated to be greater (Fig. 4). Degradation of the branched alkyl chain may
than 60 years. Therefore, estuarine, lacustrine, marine, proceed as outlined in Fig. 5. Little is known about the
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003

ALKYLPHENOL POLYETHOXYLATES IN THE ENVIRONMENT
479
Figure 4. Ethoxy side-chain transformation. The exact structure of the alkyl chain is not specified.
biodegradation pathways that lead to the mineralization wise ether hydrolysis accompanied by the loss of ethyl-
of APs and carboyxlated APEO metabolites (Fig. 6). ene glycol (Fig. 4, pathway 2). This process proceeds
n 5
rapidly until short-chained APnEOs (
1–3) are
Ethoxylate chain degradation
. Degradation of the formed. As the ethoxy side chain is shortened, APEO me-
ethoxylate chain appears to be the first APEO biotrans- tabolite solubility decreases and their persistence in-
formation process. As seen in Fig. 4, transformation of creases. Under anaerobic conditions short-chained AP-
n 5
the parent APEO can occur by two pathways: terminal nEOs (
alcohol oxidation, or ether hydrolysis (Fig. 4, pathways 8); under aerobic conditions, they are oxidized (Fig. 4,
et al
1–3) are converted to APs (Fig. 6, pathway
1 and 2, respectively). Because long-chained carboxyl- pathway 1) (Thiele
ated ethoxylates had not been detected, the primary boxylation increases both the solubility and persistence
et al
., 1997; Maguire, 1999). Car-
ethoxylate chain shortening pathway for under both aer- of these metabolites (Ahel
., 1994b).
obic and anaerobic conditions, was believed to be step- As with most biological processes, biodegradation
Figure 5. Alkyl side-chain transformation. The exact structure of the alkyl chain is not specified.
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003

480
MONTGOMERY-BROWN AND REINHARD
Figure 6. Mineralization. The exact structure of the alkyl chain is not specified.
rates are significantlyaffected by temperature. Generally, APEOs and APs appear to be the ultimate products un-
higher rates and lower concentrationsare observed in the der anaerobic conditions, it is plausible that ethoxycar-
summer, while lower rates and higher concentrations are boxylates formed, or deposited, under anaerobic condi-
et al.
et
observed in the winter (Ahel
al
, 1996; Maruyama
tions, are later transformed by hydrolysis of the terminal
., 2000). Recent research indicates that the length of the ether bond (Fig. 4, pathway 3).
ethoxy chain also depends on temperature; in rivers
around Tokyo, Japan, APEO metabolites generally had
Alkyl chain degradation
. Although alkyl chain degra-
between two and five ethoxy units in the summer and be- dation (Fig. 5) can occur concurrently with ethoxy chain
et
tween five and eight units in the winter (Maruyama
shortening, the highly branched nature of this chain of-
ten limits the rate and extent of its biodegradation. The
al
., 2000).
et al
v
. (2001) concludedthat the dom- initial step in alkyl chain degradation is -oxidation (Fig.
Recently,Jonkers
inant pathway in aerobic APEO biodegradationis the ox- 5, pathway 6), a process by which a terminal methyl
idation of the terminal alcohol on the ethoxy side chain group is converted into a carboxylic acid. After oxida-
a
b
(Fig. 4, pathway 1), not, ether hydrolysis. This conclu- tion, the alkyl chain can be shortened by - or -oxida-
n 5
b
15) tion. -Oxidation occurs unless there is a methyl group
sion was made because long-chained (up to
b
APnECs were observed almost immediately upon begin- on the -carbon (i.e., a tertiary carbon); in these cases,
a
ning their experiment. Stepwise shortening of the car-
-oxidation takes over (Swisher, 1987). Conceivably,
boxylated ethoxy chain might proceed by terminal ether this process could continue until the entire alkyl chain
hydrolysis (Fig. 4, pathway 3). Because only small has been degraded. Alkyl chains with quaternary carbon
amounts of AP2EO were detected, it is plausible that the atoms (e.g., OP) are generally very difficult to degrade,
APEOs observed in the environment are produced in as only a few pathways have been documented for the
anaerobic microenvironments. The researchers were only biotransformation of quaternary carbon atoms (Swisher,
able to account for 19% of the initial APEO—this sug- 1987).
gests either complete APEO mineralization or the for-
mation of undetected metabolites.
Using a GC/TSQ-MS with CI and EI capabilities,Ding
et al
. (1996) detected 10 dicarboxylated APEO metabo-
Under anaerobic conditions, the dominant transforma- lites (i.e., CAPECs) in a WWTP effluent. Another study
tion process appears to be stepwise ether cleavage to in Italy found that the dominant APEO metabolites in
n 5
et al.
1–3) APEOs and APs (Fig. 4, path- WWTP effluents are CAPECs (DiCorcia , 2000). In
short-chained (
way 2). Interestingly, even under strict anaerobic condi- these effluents, CA6PECs and CA8PECs (i.e., those hav-
tions, some researchers have observed substantial oxida- ing a total of seven and nine carbon atoms in the alkyl
tion of the ethoxylate chain (Fig. 4, pathway 1) (Ball chain) accounted for approximately 87% of the CAPEC
et al
., 1989; Schroder, 2001). Because short-chained class. The researchers postulated that these metabolites

ALKYLPHENOL POLYETHOXYLATES IN THE ENVIRONMENT
481
tify APEO metabolites, it is likely that many of the
APECs previously identified were, in reality, CAPECs.
v b
v
were formed by / -oxidation and -oxidation, respec-
tively, of the alkyl side chain. CA5PECs and CA7PECs
v b
v
were postulated to result from the / - and -oxidation,
Mineralization
. Evidence for aromatic ring degrada-
respectively, of OPEOs. CA5PECs may also result from
tion is mounting. For example, during the aerobic degra-
dation of NPEO, Schroder (2001) observed the disap-
pearance of NPEC without the formation of NP or
carboxylated alkyl chain metabolites (CAPECs). Potter
v b
the / -oxidation of certain NPECs (Fig. 7). CAP2ECs
were approximately twice as abundant as CAP1ECs; in-
et al
. (1996) observed that CAP1ECs
terestingly, Ding
were more abundant.
Similar results were found in batch aerobic biodegra-
et al
et al
. (1999) studiedthe aerobic static-die away of NPEOs
and their metabolites in estuarine water samples. NP2EC
was the dominant degradation product and was relatively
persistent, but it was eventually transformed; the mech-
anism whereby the NP2EC was transformed is unknown,
but no NP was detected. Other experiments monitored
the amount of carbon dioxide produced during the aero-
bic biodegradation of OP1EC, OP2EC, NP1EC, and
NP2EC and determined that only 15 to 35% of the ini-
dation studies (DiCorcia
., 1998). Interestingly, al-
though there was no significant accumulation of
NP1EC, a substantial amount of CAP1ECs were formed
or during the experiment. Over time, there was a slow
disappearance of CAP2ECs and a concurrent increase
in CAP1ECs. However, no decreases were noted in to-
tal CAPEC concentrations in more than 5 months; this
suggests that CAPECs may be very recalcitrant in the
environment. However, results from a Soil Aquifer
Treatment site in Mesa, AZ, indicate that CAPECs are
rapidly transformed during infiltration into the subsur-
et al
tial carbon was not mineralized (Staples
., 1999,
2001); these results suggest some biodegradation of the
aromatic moiety.
Despite these findings, little effort has been made to
detect transformation metabolites or determine the
biodegradation pathways. Because of this, our under-
standing of APEO mineralization in the environment is
very limited (Fig. 6). Under aerobic conditions, biodegra-
dation of the aromatic moiety is likely the dominant
APEO biotransformation mechanism. For example, Fu-
jita and Reinhard (1997) determined that OP1EC was
rapidly transformed in stoichiometric quantities, through
ring opening, to 2,4,4-trimethyl-2-pentanol. This alcohol
was resistant to further degradation. An aerobic column
study by Wild and Reinhard (1999) also observed very
rapid OP1EC transformation; OP1EC concentrations de-
creased by more than 90% in less than 30 min. However,
et al
et
face (Montgomery-Brown
al
., 2003). As DiCorcia
. (1998) were unable to account for all the NPEO they
added, they postulated the presence of undetected
metabolites.
Based on recent results, and due to the lack of selec-
tivity of the early GC and HPLC techniques used to iden-
m
residual OP1EC concentrations (0.3–3 g/L) persisted
for more than 2 months; the authors suggest that perhaps
OP1EC concentrations were too low to support micro-
bial growth. Data indicated that approximately 8.3 mol
of dissolved oxygen were required per mol of OP1EC.
Because the complete mineralization of 1 mol of OP1EC
would require 16.5 mol of dissolved oxygen and no in-
termediates were detected, it was proposed that the re-
maining material was incorporated into bacterial biomass
(Wild and Reinhard, 1999).
Further evidence for ring-opening reactions came from
branched NP degradation studies using a liquid culture
Sphingomonas
et al.
(1999) found that one
of
isolate began degrading NP by phenyl ring scission and
para ortho
sp. Tanghe
preferentially metabolized
-NP over -NP iso-
mers. This preferential degradation might explain the
ortho para
prevalence of
- vs. -isomers observed in sewage
sludges by Field and Reed (1999). Between 6.2 and 7.4
mol of oxygen were required per mol of NP oxidized;
Figure 7. One possible v/b-oxidationroute for a NP2EC iso-
mer.
ENVIRON ENG SCI, VOL. 20, NO. 5, 2003

482
MONTGOMERY-BROWN AND REINHARD
these values are similar to those obtainedfor OP1EC (Fu- their metabolites, are degraded in the environment. These
jita and Reinhard, 1997; Wild and Reinhard, 1999). There pathways are dependent on the photosensitizer and
was no distinct accumulation of products, but the re- whether the resulting solution is heterogeneous or ho-
searchers detected various branched 5 to 10 carbon al- mogeneous. Early research focused on using titanium
cohols, ethers, and esters; other than NP, no aromatic dioxide particles to enhance APEO photodegradation
et al et al.
., 1988, 1990; Pelizzetti , 1989).
compounds were detected.
rates (Hidaka
In contrast, while studying OP biodegradation using The first step in the photodegradationof APEOs appeared
inoculums from both rural and urban locations along a to be sorption of the aromatic ring or the ethoxylate chain
et al
U.K. river, Johnson
. (2000) concluded that the alkyl onto the surface of the titanium oxide. In this system, the
side chain was degraded prior to phenyl ring scission be- surface of the titanium dioxide both facilitated the photo-
cause the rate of OP transformation was greater than the formation of hydroxyl radicals and concentrated these
et al
.,
rate of phenyl ring mineralization. In fact, these re- radicals relative to the bulk solution (Pelizzetti
searchers only observed ring opening once during their 1989). The researchers determined that the next step in-
experiments. The half-life for OP in aerobic microcosms volved the competitive attack of a hydroxyl radical on
?
was between 7 and 50 days; lower degradation rates were the aromatic ring (OH addition to the ring) and the
et al.
observed when using inoculums from rural areas. Their ethoxylate chain (H abstraction) (Hidaka
, 1988;
., 1989). Several OH attacks on the aro-
et al
?
data indicated that OP is not more persistent at low con- Pelizzetti
m
centrations (,0.5 g/L) than at higher concentrations. matic ring were required to effect ring cleavage; ring
No degradationwas observedin anaerobic bed sediments. opening may also occur through the addition of mole-
Under anaerobic conditions, many researchers have ob- cular oxygen to a hydroxy cyclohexadienyl radical
et al.
served the formation of APs from short-chained APnEOs (Pelizzetti
, 1989). Complete APEO mineralization
n 5
(
et al.
, 1899; Hidaka
et al.
, 1990)
1–3) (Fig. 6, pathway 8); it is plausible that short- was observed (Pelizzetti
n 5
chained APnECs or CAPnECs (
et al
., 1989).
1–3) could also be without the formation of APs (Pelizzetti
transformed into APs, but thus far, this has not been ob-
More recently, researchers have focused on the ho-
served. Because neitherthe breakdown of the aromatic moi- mogeneous photodegradation of APEOs using trivalent
et al et al
ety nor transformationof the alkylchain have been observed iron (Brand
et al.
et al.
, 1998; Castillo
, 2001). Trivalent
under anaerobic conditions(Ejlertsson
, 1999; Johnson iron was chosen as a photosensitizer because Fe(III)–
, 2000; Schroder, 2001), it is believed that APs are the aquo complexes are very efficient at forming hydroxyl
ultimate products anaerobic APEO biodegradation.
radicals through the following reaction:
21
Fe(OH)21
Fe
1 h n R
1
OH·(aq)
(aq)
(aq)
Photodegradation
The hydroxyl radical produced in this reaction can re-
act with the alkyl chain, the ethoxy chain, or the aromatic
ring of APEOs or their metabolites. In contrast with ear-
lier results using titanium dioxide, neither research group
Very little research has focused on the importance of
photodegradationin the fate of APEOs, and their metabo-
lites, in the environment. The main factors that influence
APEO photodegradation rates are light intensity and the
?
observed OH addition to the aromatic ring. Rather, they
et
presence of photosensitizers. Early research by (Ahel
observed that the hydroxyl radical preferentially attacked
the ethoxy chain, and resulted in shorter chained APEOs.
Other intermediates and metabolites observed were:
APECs, alkylphenol formate and aldehyde ethoxylates,
alkyl chain carboxylated APEOs (CAPEOs), and
al.
, 1994c) determined that APEO degradation rates de-
crease substantiallywith depth. At the surface of the lake,
the half-life for NP was between 10 and 15 h; at depths
of 20–25 cm [dissolved organic carbon (DOC) ,4 mg-
C/L], photodegradation rates were approximately 1.5
et al.
, 1998; Castillo
et al
., 2001). Alkyl
CAPECs (Brand
et al.
times slower. Ahel
(1994c) also determined that
chain degradation and ring opening were only observed
et al.
naturally occurring photosensitizers like nitrate, Fe(III)-
aquo complexes and humic acids play a major role in the
photodegradation of APEOs and their metabolites. For
example, the half-life for NP in lake water (DOC ,4 mg-
in the presence of Fe(III) (Castillo
, 2001); no mech-
anisms for these transformations were proposed. During
prolonged irradiation (,24 h) more than 90% of the ini-
tial NPEO was transformed; no degradation products
l .
C/L) exposed to a mercury vapor lamp (
280 nm; in-
et al
., 1998; Castillo
et al.
, 2001).
were detected (Brand
3
tensity ,10 greater than natural sunlight) was approx-
imately 45 min; in distilled water, its half-life was
approximately five times longer.
In the absence of Fe(III), degradation rates were much
et al.
slower (Castillo
, 2001).
Because both of the photodegradationmechanisms de-
scribed above rely on hydroxyl radicals, it should be pos-
sible to estimate the importance of these radicals on the
Photodegradationpathways
. Relatively little is known
about the photochemical pathways whereby APEOs, and

ALKYLPHENOL POLYETHOXYLATES IN THE ENVIRONMENT
483
degradation of APEOs and their metabolites in surface
waters. However, estimating the half-lives of APEOs and
pounds and whether they are responsible for the symp-
toms of endocrine disruption observed in wild fish.
their metabolitesusing steady-state hydroxyl radical con- 6. Environmental occurrence of APEOs and their meta-
centrations in surface waters does not support experi-
mental results. This suggests that other factors and pro-
cesses likely play an important role in the photo-oxidation
of APEOs and their metabolites.
bolites. Only a few studies have looked for halo-
genated APEO metabolites. Because these metabo-
lites can be formed during the chlorination of WWTP
effluents, their environmental occurrence could be
high.
FUTURE RESEARCH NEEDS
ACKNOWLEDGMENTS
Over the past couple of years, our analytical detection
capabilitieshave increased dramatically, allowing the de-
tection and identification of APEO metabolites previ-
ously missed. The detection of these metabolites has per-
mitted a greater understanding of APEO degradation in
the environment. Future research should concentrate on:
Funding was provided by the Environmental Protec-
tion Agency (EPA) and American Water Works Associ-
ation Research Foundation (AWWARF) through a grant
to the National Center for Sustainable Water Supply at
Arizona State University in Tempe, AZ.
1. Biodegradation of APEOs and their metabolites. Cur-
rent research is still divided about the ultimate fate of
APEOs metabolites in the environment. A better un-
derstanding of the biodegradation pathways and the
enzymatic regulation of these pathways is necessary.
For example, how common are ring-opening path-
ways? Why are NP and other APEO metabolites per-
sistent under anaerobic conditions when other pheno-
lic compounds are degradable? Can CAPECs be
further degraded? Can a single bacterium mineralize
APEOs or is a consortium of bacteria required? Why
do low concentrations of APEO metabolites persist in
the environment?
2. Photodegradation of APEOs and their metabolites.
Further research is necessary to determine whether
photodegradation plays an important role in the at-
tenuation of APEOs and their metabolites in the en-
vironment, and if so, to determine the pathways.
3. Analytical detection capabilities. The lack of mass
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