United States Environmental Office of Water EPA-822-R-05-005
Protection Agency 4304T December 2005
Aquatic Life Ambient
Water Quality Criteria -
Nonylphenol
FINAL
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EPA-822-R-05-005
Ambient Aquatic Life Water Quality Criteria
Nonylphenol
(CAS Registry Number 84852-15-3)
(CAS Registry Number 25154-52-3)
FINAL
December 2005
U.S. Environmental Protection Agency
Office of Water
Office of Science and Technology
Washington, DC
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NOTICE
This document has been reviewed in accordance with U.S. EPA policy and approved for
publication. Mention of trade names or commercial products does not constitute endorsement or
recommendation for use.
This document can be downloaded from EPA's website at:
http ://www. epa. gov/waterscience/criteria/aqlife.html
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FOREWORD
Section 304(a)(l) of the Clean Water Act of 1977 (P.L. 95-217) requires the
Administrator of the Environmental Protection Agency to publish water quality criteria that
accurately reflect the latest scientific knowledge on the kind and extent of all identifiable effects
on health and welfare that might be expected from the presence of pollutants in any body of
water, including ground water. This document is a revision of proposed criteria based upon
consideration of comments received from independent peer reviewers and the public. Criteria
contained in this document replace any previously published EPA aquatic life criteria for the
same pollutant(s).
The term "water quality criteria" is used in two sections of the Clean Water Act, section
304(a)(l) and section 303(c)(2). The term has a different program impact in each section. In
section 304, the term represents a non-regulatory, scientific assessment of health or ecological
effects. Criteria presented in this document are such scientific assessments. If water quality
criteria associated with specific waterbody uses are adopted by a state or tribe as water quality
standards under section 303, they become enforceable maximum acceptable pollutant
concentrations in ambient waters within that state or tribe. Water quality criteria adopted in state
or tribal water quality standards could have the same numerical values as criteria developed
under section 304. However, in many situations states or tribes might want to adjust water
quality criteria developed under section 304 to reflect local environmental conditions and human
exposure patterns. Alternatively, states or tribes may use different data and assumptions than
EPA in deriving numeric criteria that are scientifically defensible and protective of designated
uses. It is not until their adoption as part of state or tribal water quality standards that criteria
become regulatory. Guidelines to assist the states and tribes in modifying the criteria presented
in this document are contained in the Water Quality Standards Handbook (U.S. EPA 1994). The
handbook and additional guidance on the development of water quality standards and other
water-related programs of this agency have been developed by the Office of Water.
This final document is guidance only. It does not establish or affect legal rights or
obligations. It does not establish a binding norm and cannot be finally determinative of the
issues addressed. Agency decisions in any particular situation will be made by applying the
Clean Water Act and EPA regulations on the basis of specific facts presented and scientific
information then available.
Ephraim S. King, Director
Office of Science and Technology
in
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ACKNOWLEDGMENTS
Document Authors
Larry T. Brooke
Great Lakes Environmental Center
Superior, WI
Glen Thursby
U.S. Environmental Protection Agency
Office of Research and Development
National Health and Environmental Effects Research Laboratory
Narragansett, RI
U.S. EPA Document Coordinator
Frank Gostomski Office of Water
U.S. EPA Technical Reviewers
Tala Henry Office of Water
Walter Berry Office of Research and Development
Bruce Boese Office of Research and Development
Denise Champlin Office of Research and Development
Valerie Chan Office of Research and Development
Michael Lewis Office of Research and Development
Cindy Roberts Office of Research and Development
Donald Rodier Office of Prevention, Pesticides and Toxic Substances
Robert Spehar Office of Research and Development
Charles Stephan Office of Research and Development
Glen Thursby Office of Research and Development
IV
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CONTENTS
FOREWORD iii
ACKNOWLEDGMENTS iv
FIGURES vii
TABLES vii
1. INTRODUCTION 1
1.1. Physical-Chemical Properties 1
1.2. Nonylphenol in the Environment 2
1.3. Metabolism and Bioconcentration 5
1.4. Estrogenicity of Nonylphenol 6
1.5. Derivation of Aquatic Life Ambient Water Quality Criteria 8
2. ACUTE TOXICITY TO AQUATIC ANIMALS 9
2.1. Freshwater 9
2.2. Saltwater 10
3. CHRONIC TOXICITY TO AQUATIC ANIMALS 13
3.1. Freshwater 13
3.2. Saltwater 15
3.3. Acute-Chronic Ratios 16
4. TOXICITY TO AQUATIC PLANTS 18
4.1. Freshwater 18
4.2. Saltwater 18
5. BIO ACCUMULATION 19
5.1. Freshwater 19
5.2. Saltwater 19
6. OTHER DATA 20
6.1. Freshwater 20
6.2. Saltwater 23
6.3. Reproductive, Devleopmental and Estrogenic Effects of Nonylphenol 26
7. UNUSED DATA 30
8. SUMMARY 32
8.1. Freshwater Data 32
8.2. Saltwater Data 32
8.3. Plant Data 33
8.4. Bioaccumulation Data 33
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9. NATIONAL CRITERIA 34
9.1. Freshwater 34
9.2. Saltwater 34
10. IMPLEMENTATION 35
11. REFERENCES 68
VI
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FIGURES
Figure 1. Ranked Summary of Nonylphenol GMAVs - Freshwater 37
Figure 2. Ranked Summary of Nonylphenol GMAVs - Saltwater 38
Figure 3. Chronic Toxicity of Nonylphenol to Aquatic Animals 39
TABLES
Table 1. Acute Toxicity of Nonylphenol to Aquatic Animals 40
Table 2. Chronic Toxicity of Nonylphenol to Aquatic Animals 45
Table 3. Ranked Genus Mean Acute Values with Species Mean Acute-Chronic Ratios 47
Table 4. Toxicity of Nonylphenol to Aquatic Plants 50
Table 5. Bioaccumulation of Nonylphenol by Aquatic Organisms 51
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 55
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1. INTRODUCTION
1.1. Physical-Chemical Properties
Nonylphenol (CisH^O) is produced from cyclic intermediates in the refinement of petroleum
and coal-tar crudes. It is manufactured by alkylating phenol with mixed isomeric nonenes in the
presence of an acid catalyst. The resulting product is a mixture of various isomers of
nonylphenol, predominantly /ram-substituted nonylphenol, (phenol, 4-nonyl-branched, CAS No.
84852-15-3; 4-nonylphenol, CAS No. 104-40-5; and phenol, nonyl-, CAS No. 25154-52-3) with
small amounts of ort/zo-substituted phenol (2-nonylphenol, CAS No. 136-83-4), and trace
amounts of 2,4-dinonylphenol (phenol, dinonyl, branched, CAS No. 84962-08-3). Additional
isomers, which represent the numerous branched structures that occur within the nonyl (nine
carbon) group, add to the complexity of the compound. Commercial nonylphenol is most
accurately described by CAS number 84852-15-3 (phenol, 4-nonyl-branched), but CAS numbers
104-40-5 (phenol, 4-nonyl-) and 25154-52-3 (phenol, nonyl) have also been used to describe
these compounds commercially. The criteria derived in this document address the CAS numbers
84852-15-3 and 25154-52-3.
There is little direct use for nonylphenol except as a mixture with diisobutyl phthalate to
color fuel oil for taxation purposes and with acylation to produce oxime as an agent to extract
copper. Most nonylphenol is used as an intermediate in the production of other chemicals.
Notably, nonionic surfactants of the nonylphenol ethoxylate type are produced through
etherification of nonylphenol by condensation with ethylene oxide in the presence of a basic
catalyst. The nonionic surfactants are used as oil soluble detergents and emulsifiers that can be
sulfonated or phosphorylated to produce anionic detergents, lubricants, antistatic agents, high
performance textile scouring agents, emulsifiers for agrichemicals, antioxidants for rubber
manufacture, and lubricant oil additives (Reed 1978).
Nonylphenol is produced in large quantities in the United States. Production was 147.2
million pounds (66.8 million kg) in 1980 (USITC 1981), 201.2 million pounds (91.3 million kg)
in 1988 (USITC 1989), 230 million pounds (104 million kg) in 1998 (Harvilicz 1999), and
demand is increasing about 2 percent annually. Nonylphenol is a pale yellow highly viscous
liquid with a slight phenolic odor, an approximate molecular weight of 215.0 to 220.4 g/mole, a
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specific gravity of 0.953 g/mL at 20°C (Budavari 1989), and a vapor pressure of 4.55 x 10"3
(+3.54 x 10"3) Pa (Roy F. Weston Inc. 1990). It has a dissociation constant (pKa) of 10.7+1.0
and a log octanol/water partition coefficient (log Kow) of 3.80 to 4.77 (Roy F. Weston Inc. 1990).
The water solubility of nonylphenol is pH-dependent; 4,600 ug/L at pH 5.0, 6,237 ug/L at pH
7.0, 11,897 ug/L at pH 9.0. Nonylphenol is soluble in seawater at 3,630 ug/L and is soluble in
many organic solvents (Roy F. Weston Inc. 1990). Ahel and Giger (1993) measured the
solubility of nonylphenol at different temperatures in distilled water and demonstrated a nearly
linear increase in solubility between 2°C (4,600 ug/L) and 25°C (6,350 ug/L).
1.2. Nonylphenol in the Environment
Nonylphenol and nonylphenol ethoxylates have been found in the environment and a review
of studies describing their distribution has been published (Bennie 1999). Shackelford et al.
(1983) reported 4-nonylphenol at average concentrations ranging from 2 to 1,617 ug/L in eleven
water samples associated with various industrial sources. Bennie et al. (1997) measured
nonylphenol in water in 25 percent of sites sampled in the Great Lakes at concentrations from
0.01 to 0.92 ug/L. They found nonylphenol in all sediment samples with concentrations ranging
from 0.17 to 72 ug/g (dry weight). Nonylphenol and its ethoxylates have been found in
treatment plant wastewaters (Ellis et al. 1982, Giger et al. 1981) and in sewage sludges (Giger et
al. 1984). In a study of airport runoff, nonylphenol was measured at 0.98 and 7.67 ug/L in the
runoff as a result of aircraft deicer and antiicer fluid use (Corsi et al. 2003). A study was
conducted of thirty river reaches in the continental U.S. in 1989 and 1990 to determine the
frequency and concentrations of nonylphenol and its ethoxylates in water and sediments.
Nonylphenol was found in approximately 30 percent of the water samples with concentrations
ranging from about 0.20 to 0.64 ug/L. Approximately 71 percent of the sampling sites had
measurable concentrations of nonylphenol in the sediments at concentrations ranging from about
10 to 2,960 ug/kg. Ethoxylates of nonylphenol were found in 59 to 76 percent of the water
samples, with amounts varying by extent of ethoxylation (Naylor 1992, Naylor et al. 1992,
Radian Corp. 1990).
Most nonylphenol enters the environment as 4-alkylphenol polyethoxylate surfactants which
are degraded to 4-alkylphenol mono- and diethoxylates in active sewage sludges (Giger et al.
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1984). It was theorized by Giger et al. (1984) that further transformation of 4-alkylphenol
mono- and diethoxylates to 4-nonylphenol is favored by anaerobic environments. They
conducted experiments with stabilized (anaerobic) and raw (aerobic) sewage sludge and found
that concentrations of 4-nonylphenol increased four to eight times in the stabilized versus two
times in the raw sludge, a finding which supported their theory.
A reconnaissance of 95 organic wastewater contaminants in 139 U.S. streams conducted in
1999-2000 revealed that nonylphenol was one of the most commonly occurring contaminants
and was measured at higher concentrations than most of the other contaminants (Kolpin et al.
2002). Selection of streams sampled was biased toward streams susceptible to contamination. A
number of studies on the persistence of nonylphenol in sewage treatment plant effluents and the
environment have been conducted and are reviewed by Maguire (1999). Gaffney (1976)
determined that 1 mg/L nonylphenol did not degrade during 135-hr incubation with domestic
wastewater. In industrial wastewater, nonylphenol concentration was unchanged after 24 hr
incubation, but decreased by 45 percent after 135 hr. Staples et al. (2001) determined that
nonylphenol at 13 mg/L and 22 °C was mineralized to CC>2 within 35 days in aerobic systems
inoculated with sludge from a waste treatment plant. No intermediate compounds were formed
and the calculated half-life for nonylphenol was 8.2 days.
Sundaram and Szeto (1981) studied nonylphenol fate incubated in open and closed
containers of stream and pond waters. They found no degradation of nonylphenol incubated in
open containers of the pond or stream waters. The observed half-life of 2.5 days, was attributed
to volatilization. Incubation of nonylphenol in pond or stream waters in closed containers
resulted in formation of a breakdown product. The observed half-life of nonylphenol in pond
and stream water were estimated at 16.5 and 16.3 days, respectively. The same authors
incubated nonylphenol in pond water with sediment present and found about 50 percent of the
nonylphenol in the sediment after 10 days. About 80 percent of the nonylphenol in the sediment
was degraded in 70 days. No degradation of nonylphenol occurred when autoclaved (sterilized)
water and sediment samples were used. Staples et al. (1999) measured a half-life of 20 days for
nonylphenol in water (31 mg/L) at 22 °C. They suggested that the temperature of water and the
initial concentration of the nonylphenol both affect the degradation rate of the chemical.
Ahel et al. (1994 a,b) studied the fate and transport of alkylphenol polyethoxylates (APwEO)
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and their degradates in the Glatt River system in Northern Switzerland from the Greifensee to
the Rhine River. Water samples were collected at eight sites along the river hourly over several
seasons. They found nonylphenol concentrations to be lower than other degradates and
nonylphenol concentrations were most commonly detected in the 1 to 3 ug/L range. The
concentration of APnEO degradates varied with time of day reflecting fluctuations in wastewater
treatment plant discharge. Concentrations of APnEO degradates also varied seasonally, being
found at higher concentrations in the winter due to lower water temperature. Nonylphenol had
less seasonal variability than other APnEO degradates. Nonylphenol was the predominant
nonylphenolic compound found in sediments in this study. Sediment concentrations were 364 to
5,100 times those found in the river water.
Ahel and co-workers also reported that the abundance of particular APnEO degradates is
dependent on the conditions in the treatment plants studied along the Glatt River system (Ahel et
al. 1994a; Ahel et al. 2000). Under aerobic conditions, the APwEOs degrade through either the
loss of ethylene oxide units to form low-molecular weight ethoxylates or through the formation
of carboxylated ethoxylates ultimately terminating in CC>2 and water. Nonylphenol is formed
during anaerobic breakdown of the APwEOs and is therefore a minor component of wastewater
treatment effluents because of aerobic conditions present during treatment. Another study by
Ahel et al. (1996) demonstrated that nonylphenol can be reduced in ground water. The authors
propose that biological processes are responsible provided that the ground water temperature
does not become too cold for biological degradation. It has also been demonstrated (Ahel et al.
1994c) that nonylphenol can be degraded by photochemical processes. In bright summer sun,
nonylphenol near the water surface has a half-life of 10-15 hr.
Heinis et al. (1999) studied the distribution and persistence of nonylphenol in natural pond
systems in the temperate climate zone. They reported that nonylphenol partitioned to the pond
enclosure wall material, macrophytes, and sediments within two days. After 440 days, the
primary sink for nonylphenol was the sediment. Dissipation from the sediment was estimated to
be 50 and 95 percent at 66 and 401 days, respectively. Hale et al. (2000) measured nonylphenol
concentrations in sediments below wastewater outfalls and found one site that had a sediment
concentration of 54,400 ug/kg more than twenty years after the treatment plant ceased operation.
Bennett and Metcalfe (1998; 2000) found that nonylphenol was widely distributed in lower
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Great Lakes sediments and reached 37,000 ug/kg in sediments near sewage treatment plants.
It appears that degradation of nonylphenol in sea water and saltwater sediments may be
slower than in fresh water and freshwater sediments. Ekelund et al. (1993) found that
nonylphenol degradation rate was initially slow in sea water, but increased after microorganism
adaptation occurred. Approximately 50 percent of the nonylphenol was degraded after 58 days.
In marine sediments, the initial rate of degradation faster than in sea water, but after 58 days
about the same percentage of nonylphenol was degraded. Ethoxylated nonylphenol has a half-
life of 60 days in marine sediments, similar to that of nonylphenol (Shang et al. 1999). Ferguson
and Brownawell (2003) conducted degradation studies with APwEOs in marine sediments and
found that degradation occurred in oxic and anoxic conditions. They reported no clear evidence
for net formation of nonylphenol from APnEOs under anaerobic conditions during the 120 day
study, but they speculated that the time scale of their study may not have been long enough to
make the observation.
1.3. Metabolism and Bioconcentration
Nonylphenol is metabolized by hepatic cytochrome P450 enzymes in the rainbow trout
(Oncorhynchus mykiss) and bile from the fish contained the glucuronic acid conjugates of
nonylphenol (Meldahl et al. 1996; Thibaut et al. 1999). Arukwe et al. (2000) found that bile was
the major route of nonylphenol excretion with a half-life of 24 to 48 hrs following either
waterborne or dietary exposures.
The log Kow of nonylphenol ranges from 3.80 to 4.77, indicating that moderate
bioaccumulation in aquatic organisms may be expected. However, reported laboratory
bioconcentration factors (BCFs) and field-derived bioaccumulation factors (BAFs) do not
support the expected accumulations in tissues, indicating that some nonylphenol is metabolized.
Bioconcentration was measured in two saltwater organisms, the blue mussel (Mytilus edulis) and
Atlantic salmon (Salmo salaf) by McLeese et al. (1980a). The estimated BCF for the blue
mussel ranged from 1.4 to 7.9 and the estimated BCF for Atlantic salmon was 75 (McLeese et al.
1981). Hecht et al. (2004) reported nonylphenol BCFs for the three marine amphipod species,
Eohaustorius estuarium, Grandidierella japonica and Corophidum salmonis, of 154, 185, and 46
to 133, respectively. Ahel et al. (1993) measured the bioconcentration of nonylphenol for
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several species in rivers in Switzerland. They determined a BCF for algae of 487 (converted to a
wet weight basis assuming 95 percent water in algae). Nonylphenol did not biomagnify in the
food chain in the system studied; rather BCFs in fish and ducks were lower than in the algae.
Keith et al. (2001) measured nonylphenol in fish tissues of seven species from the Kalamazoo
River and in water at the river's confluence with Lake Michigan. They found 41 percent of the
tissue samples had measurable concentrations of nonylphenol with a range of 3.3 to 29.1 ug/kg
and a mean value of 12.0 ug/kg. A followup study was conducted in the same river (Kannan et
al. 2003) to further examine the occurrence of nonylphenol and nonylphenol ethoxylates in fish,
water and sediments and their association with two wastewater treatment plants. Ten fish from
near each treatment plant were analyzed for nonylphenol and ethoxylated nonylphenol. Only
one fish contained a measurable concentration of nonylphenol (3.4 ug/kg). Neither nonylphenol
or its ethoxylates were detected in the sediments collected upstream of the treatment plants.
However, five of twenty-four (21 %) sediment samples collected from below the treatment
plants contained nonylphenol (no ethoxylates were found) at concentrations that ranged from 2 -
15.3 ug/kg dry weight. Downstream of one treatment plant, neither nonylphenol nor
nonylphenol ethoxylates were measured above the method detection limit. Nonylphenol
concentrations extracted from sediments in the Venice, Italy lagoon were higher in areas with
large masses of decomposing macroalgae (primarily Ulva rigida) than in areas not associated
with the decomposition (Marcomini et al. 1990). This may suggest that nonylphenol
bioaccumulated by the macroalgae was transferred to the sediment as the algae died and
decomposed.
1.4. Estrogenicity of Nonylphenol
There are several review articles that describe the estrogenicity of nonylphenol (Servos 1999;
Sonnenschein and Soto 1998; Sumpter 1998). The majority of studies using aquatic species
models report results for molecular or biochemical endpoints such as induction of the egg
protein, vitellogenin, or are in vitro studies such as receptor binding assays. These types of
studies and endpoints do not meet the data acceptability requirements outlined in EPA's
Guidelines for Deriving Numerical National Water Quality Criteria for the Protection of
Aquatic Organisms and Their Uses (Stephan et al. 1985) and hence were not used deriving
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ambient water quality criteria. However, studies identified in the literature search describing
effects of nonylphenol on molecular and biochemical endpoints and activity in in vitro bioassays
are discussed in Section 6 of this document.
Whole organism endpoints such as reproductive and growth effects are used to derive aquatic
life ambient water quality criteria for nonylphenol. To the extent that such endpoints reflect the
integration of molecular, biochemical and tissue-level effects at the whole organism level, the
nonylphenol criteria address the estrogenicity of nonylphenol. For example, while vitellogenin
is a commonly used biomarker indicative of exposure to estrogenic compounds, measurement of
this molecular/biochemical endpoint alone does not necessarily indicate adverse effect on
population relevant endpoints such as survival, growth and reproduction. However, several
studies have demonstrated that vitellogenin induction can be accompanied by decreased
fecundity (egg production) of breeding pairs of fathead minnows exposed chronically to
estrogenic compounds (Ankley et al.). The chronic toxicity studies used in deriving the
nonylphenol criteria (Table 6) included assessment of effects on growth and reproduction
endpoints in aquatic organisms. Hence, to the extent that these endpoints are the result of effects
on the endocrine system (although this was not definitively demonstrated in any of the tests by
use of a concommittant measure of a estrogen-receptor specific endpoint), the estrogenic effects
of nonylphenol have been considered in deriving the aquatic life ambient water quality criteria
for nonylphenol.
EPA has activities underway to develop scientific methods for considering endocrine effects,
such as the estrogenicity of nonylphenol, in Agency risk assessments. Under the Federal Food,
Drug and Cosmetic Act (FFDCA), as amended by the Food Quality Protection Act (FQPA), EPA
is required to develop a screening program to determine whether certain substances (including
all pesticide active and other ingredients) "may have an effect in humans that is similar to an
effect produced by a naturally-occurring estrogen, or other such endocrine effects as the
Administrator may designate". Following the recommendations of its Endocrine Disrupter
Screening and Testing Advisory Committee (EDSTAC), EPA determined that there was
scientific basis for including, as part of the program, the androgen and thyroid hormone systems,
in addition to the estrogen hormone system. EPA also adopted EDSTAC's recommendation that
the Program include evaluations of potential effects in aquatic life and wildlife. When the
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appropriate screening and or testing protocols being considered under the Agency's Endocrine
Disrupter Screening Program have been developed, nonylphenol may be subjected to additional
screening and or testing to better characterize effects related to endocrine systems.
1.5. Derivation of Aquatic Life Ambient Water Quality Criteria
A comprehension of the "Guidelines for Deriving Numerical National Water Quality Criteria
for the Protection of Aquatic Organisms and Their Uses" (Stephan et al. 1985), hereafter referred
to as the Guidelines, is necessary to fully understand the text, tables, and calculations presented
in this criteria document. Results of intermediate calculations are presented to four significant
figures to prevent round-off error in subsequent calculations, not to reflect the precision of the
value. Final criteria values are presented to two significant figures.
Nonylphenol has been studied for its acute and chronic toxicity to aquatic organisms and
results of many studies are summarized in a review article by Staples et al. (1998). This review
article also addresses the ability of nonylphenol to bioaccumulate in aquatic organisms. Much of
the data reported in the review article has been used in this document, as well as some newer
data, to derive the aquatic life ambient water quality criteria. The latest comprehensive literature
search for information used in developing this document was conducted in November 1999.
Subsequently, forty-three newer studies have subsequently been identified and included. Data
and analysis included in the U.S. EPA's Office of Pollution Prevention and Toxics nonylphenol
risk assessment have also been evaluated in deriving the aquatic life criteria for nonylphenol.
Freshwater criteria were derived using nonylphenol of CAS numbers 25154-52-3 and 84852-15-
3; saltwater criteria were derived using only nonlyphenol of CAS number 84852-15-3.
Whenever adequately justified, a national criterion may be replaced by a site-specific
criterion (U.S. EPA 1983), which may include not only site-specific criterion concentrations
(U.S. EPA 1994), but also site-specific averaging periods and frequencies of allowed excursions
(U.S. EPA 1991).
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2. ACUTE TOXICITY TO AQUATIC ANIMALS
2.1. Freshwater
The acute toxicity of nonylphenol to freshwater animals has been determined for 18 species
and 2 subspecies representing 15 genera (Table 1). Species Mean Acute Values (SMAV) ranged
from 55.72 ug/L for an amphipod (Hyalella aztecd) to 774 ug/L for a snail (Physella virgatd).
The most sensitive freshwater species tested was the amphipod, Hyalella azteca (Tables 1
and 3). Brooke (1993a) and England and Bussard (1995) tested this species under similar
conditions, except for water hardness levels which were 51.5 and 148-154 mg/L as CaCOs,
respectively. An LC50 of 20.7 ug/L was calculated in the lower hardness water and 150 ug/L in
the higher hardness water. Insufficient data exist to demonstrate an effect of water hardness on
the toxicity of nonylphenol; therefore, the results are given equal weight for determining the
SMAV. Data for one cladoceran species (Daphnia magna) are available. Brooke (1993a)
reported an EC50 of 104 ug/L from a test that had the solutions renewed daily and Comber et al.
(1993) reported an EC50 of 190 ug/L in a static test. The Daphnia magna SMAV is 140.6 ug/L.
The least sensitive freshwater species to nonylphenol toxicity were also invertebrates (Tables
1 and 3). The annelid worm (Lumbriculus variegatus) had an LC50 of 342 ug/L, nymphs of the
dragonfly Ophiogomphus sp. had an LC50 of 596 ug/L and the least sensitive species tested was
a snail, Physella virgata, which had an LC50 of 774 u/L (Brooke 1993a). The lower sensitivity
to nonylphenol occurs even though this species of snail does not have an operculum and would
not be able to completely enclose its body and thus protect itself against nonylphenol exposure.
The midge, Chironomus tentans, had an LC50 of 160 ug/L (England and Bussard 1995),
indicating intermediate sensitivity among invertebrate species tested (Figure 1).
The only amphibian toxicity test available was for the boreal toad, Bufo boreas. The toad
tadpoles had a 96-hr LC50 of 120 ug/L and were ranked second in sensitivity to nonylphenol
(Dwyeretal. 1999a).
Freshwater fish species were in the mid-range of sensitivity to nonylphenol (Figure 1).
SMAVs ranged from 110 ug/L for the fountain darter (Etheostoma rubrum) to 289.3 ug/L for the
bonytail chub (Gila elegans). Three trout species of the genus Oncorhynchus (rainbow trout,
apache trout, and cutthroat trout) and two subspecies of the species Oncorhynchus clarki were
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tested and had similar LCSOs ranging from 140 to 270 ug/L (Dwyer et al. 1995; Brooke 1993a).
Dwyer et al. (1995, 1999a) exposed nine species offish that were classified as
threatened/endangered or were surrogates of threatened or endangered fish species. Acute
toxicity test results were based on static tests with unmeasured nonylphenol concentrations and
the LCSOs ranged from 110 ug/L for the fountain darter, Etheostoma rubmm, to a geometric
mean of 289.3 ug/L calculated from two tests with the bonytail chub. In addition to the test
conducted by Dwyer et al. (1995), two additional tests were available for the fathead minnows
(Pimephalespromelas). LCSOs for this species ranged from 128 ug/L (Brooke 1993a) to 360
ug/L (Dwyer et al. 1995). The tests conducted using flow-through exposure conditions
(Holcombe et al. 1984; Brooke 1993a), which is preferable to static exposure conditions
(Stephan et al. 1985) were used in calculating the SMAV (158.9 ug/L). One test was available
for the bluegill (Lepomis macrochirus) and the LC50 was 209 ug/L (Brooke 1993a).
Freshwater Species Mean Acute Values (SMAV) and Genus Mean Acute Values (GMAV)
were derived from available acute values (Tables 1 and 3, respectively). GMAVs were available
for 15 genera; the most sensitive was the amphipod, H. azteca, which was 13.9 times more
sensitive than the least sensitive species, a snail P. virgata (Figure 1). The four most sensitive
species were within a factor of 2.5 of one another. Based on available data for freshwater
organisms summarized in Table 1 and the GMAVs presented in Table 3, the freshwater Final
Acute Value (FAV) for nonylphenol is 55.69 ug/L (calculated using the procedure described in
the "Guidelines"). This FAV is essentially the same as the lowest freshwater SMAV of 55.72
ug/L for the amphipod H. azteca.
2.2. Saltwater
The acute toxicity of nonylphenol to saltwater animals has been determined for 8
invertebrate and 3 fish species (Table 1). SMAVs ranged from 17 ug/L for the winter flounder,
Pleuronectes americanus, to 209.8 ug/L for the sheepshead minnow, Cyprinodon variegatus
(Lussier et al. 2000; Ward and Boeri 1990b), a difference of 12.3-fold. Fish (winter flounder),
bivalves (coot clam, Mulinia lateralis) and crustaceans (the mysid, Americamysis bahia) were
the most sensitive species.
Data for nine of the thirteen saltwater test values reported in Table 1 were from a single
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multi-species test (Lussier et al. 2000). Nonylphenol concentrations were measured in seven of
the nine tests (Table 1), with measurements made at test initiation and at the end of the test (48
or 96 hr). Test organisms were fed brine shrimp, Artemia sp., during chemical exposure because
the tests were designed to extend beyond the usual 48- or 96-hr acute test interval to 168 hr. The
extended exposure time required feeding to ensure survival of animals not affected by
nonylphenol. Normally, data gathered from tests in which organisms are fed are not acceptable
for use in deriving Final Acute Values. However, the brine shrimp fed during the tests were
"reference grade" and not likely to change the exposure to nonylphenol. Further, additional tests
conducted in a different laboratory are available for two of the saltwater species such that
toxicity results obtained when the testing is conducted with and without food added can be
compared. In a 96-hr test with the mysid, the estimated LC50 was somewhat higher when the
organisms were fed (60.6 ug/L; Lussier et al. 2000) compared to when they were not fed (43
ug/L; Ward and Boeri 1990a) during the study. In contrast, in a 96-hr test with the sheepshead
minnow, the LC50 determined when the organisms were fed (142 ug/L; Lussier et al. 2000) was
lower than when the organisms were not fed (310 ug/L; Ward and Boeri 1990a) during the study.
These data indicate that feeding during the tests did not consistently increase or decrease the
LC50 estimates, and therefore feeding is assumed not to have altered the results in these tests.
Hence, the data from the Lussier et al. (2000) tests were used in deriving a saltwater Final Acute
Value.
Acute toxicity test data were available for a number of other saltwater species. Invertebrates
tested include: coot clam, Mulinia lateralis (LC50 = 37.9 ug/L; Lussier et al. 2000), the copepod,
Acartia tonsa (LC50 =190 ug/L; Kusk and Wollenberger 1999), American lobster, Homarus
americanus (LC50 = 71 ug/L; Lussier et al. 2000), mud crab, Dyspanopeus sayii (LC50 >195
ug/L; Lussier et al. 2000) and the amphipods, Leptocheirusplumulosus (LC50 = 61.6; Lussier et
al. 2000) and Eohaustorius estuarius (LC50 =138 ug/L; Hecht and Boese 2002). The test with
the amphipod E. estuarius (Hecht and Boese 2002) was conducted as a 96-hr test with a mean
LC50 for toxicity measured at 227 ug/L as the average of three tests (299, 194, 189 ug/L). The
ability of the surviving organisms to bury themselves in sediment at 96 hr when placed on
sediment was combined with the number of survivors to calculate an EC50. The mean EC50 for
the three tests was 138 ug/L. The sensitivity of the saltwater fish inland silversides (Menidia
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beryllina)., was intermediate (LC50 = 70 ug/L) among the three saltwater fish species tested.
Saltwater Species Mean Acute Values (SMAV) and Genus Mean Acute Values (GMAV)
were derived from available acute values (Tables 1 and 3, respectively). GMAVs were available
for 11 genera; the most sensitive was the winter flounder, Pleuronectes americamis, which was
12.3 times more sensitive than the least sensitive species, the sheepshead minnow, Cyprinodon
variegates (Table 1 and 3). GMAVs for the four most sensitive saltwater species differ by a
factor of only 3.5 (Table 3 and Figure 2). Based on available data for freshwater organisms
summarized in Table 1 and the GMAVs presented in Table 3, the freshwater Final Acute Value
(FAV) for nonylphenol is 13.93 ug/L (calculated using the procedure described in the
"Guidelines"). This FAV is lower than the lowest SMAV of 17 ug/L for the the winter flounder,
Pleuronectes americanus.
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3. CHRONIC TOXICITY TO AQUATIC ANIMALS
3.1. Freshwater
The chronic toxicity of nonylphenol was determined for 5 freshwater species, two fish and 3
invertebrates (Table 2). Concentrations of nonylphenol were measured in all the tests. England
(1995) exposed neonates of a cladoceran, Ceriodaphnia dubia, to nonylphenol for seven days in
a renewal test. The results showed a significant reproductive impairment at 202 ug/L, but not at
88.7 p,g/L, and survival was reduced at 377 ug/L, but not at 202 ug/L. Based upon reproductive
impairment, the Chronic Value for C. dubia was 133.9 u/L. At the end of 48 hr in the same test,
effects were observed and an EC50 of 69 ug/L was calculated. However, the animals had
received food and according to the Guidelines acute tests with this species must not receive food
during an acute toxicity test if the test is to be used to compute an Acute-Chronic Ratio (ACR).
Fliedner (1993) exposed 4 to 24 hr-old Daphnia magna neonates to nonylphenol for 22 days
in a 20° C life-cycle test. Test solutions were renewed three times each week during which a
52.2 to 65.5 % decrease in nonylphenol concentration was measured. Mean measured
nonylphenol test concentrations were: 0, 0, 1.55, 1.34, 3.45, 10.70, and 47.81 ug/L. No effects
were observed during the study on mortality, the number of offspring per female, or the mean
day of the first brood at any of the test concentratoins. A significant effect was observed on the
total number of young per concentration on day nine of the study. Based on the No Observed
Effect Concentration (NOEC) of 10.7 ug/L and the Lowest Observed Effect Concentration
(LOEC) of 47.8 u/L reported, the chronic value (geometric mean of the NOEC and LOEC) for
D. magna in this test is 22.62 u/L. An acute test with this species conducted by the same authors
was not available to calculate an ACR.
Brooke (1993a) conducted a 21-day chronic exposure for the cladoceran Daphnia magna.
Test solutions were renewed three times per week and concentrations of nonylphenol declined,
on average, 57.4 + 5.8 % between solution renewals. The author concluded that/), magna
growth and reproduction were significantly affected at 215 ug/L, but not at 116 ug/L. Survival
was reduced to 60 percent at 215 ug/L; however, this survival rate was not a significant
reduction from the control survival rate because only 80 percent of organisms survived in the
control group. Based on reproductive impairment, the chronic value, calculated as the geometric
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mean of the lower (116 ug/L) and upper (215 ug/L) chronic limits, for this test was 157.9 ug/L.
Dividing the acute value (104 ug/L), determined from a companion test for this species (Brooke
1993a; Table 1) by the chronic value (157.9 ug/L; Table 2) results in an ACR 0.6586 for D.
magna (Table 2).
A third life-cycle test (21-day exposure) with D. magna was conducted by Comber et al.
(1993). They found no significant effects in survival, reproduction or growth at concentrations
<24 ug/L. The number of live young produced was significantly reduced at concentrations > 39
ug/L when compared to controls. Growth was reduced at concentrations > 71 ug/L and survival
of adults was reduced at concentrations > 130 u/L. Based on reproductive impairment, the
chronic value, calculated as the geometric mean of the lower (24 ug/L) and upper (39 ug/L)
chronic limits, for this test was 30.59 ug/L. Dividing the acute value (190 ug/L), determined in
in a companion test for this species (Table 1) by the chronic value (30.59 ug/L; Table 2) results
in an ACR of 6.211 for D. magna. Calculating the geometric mean of the two ACRs for D.
magna (0.6586 and 6.211) results in a species mean acute-chronic ration (SMACR) of 2.023 for
D. magna.
The midge, Chironomus tentans, was exposed in a continuous-flow diluter to nonylphenol
from <24-hr old larva through emergence (53 days) as adults (Kahl et al. 1997). Nominal
exposure concentrations ranged from 12.5 to 200 ug/L, but mean measured concentrations were
lower. Neither growth nor reproductive endpoints (sex ratio, emergence pattern, and egg
production and viability) were negatively affected at any of the exposure concentrations. There
was a significant effect on survival of larvae during the first 20 days of exposure, but no effect
after 20 days. Based on survival at 20 days, the NOEC and LOEC for this study were 42 and 91
ug/L, respectively. The chronic value, calculated as the geometric mean of the NOEC and the
LOEC, is 61.82 ug/L for this test. A companion acute toxicity test was not conducted; therefore,
an ACR can not be calculated for this species.
A 91-day early life-stage test was conducted with embryos and fry of the rainbow trout,
Oncorhynchus mykiss (Brooke 1993a). Five nonylphenol exposure concentrations were tested,
ranging from 6.0 to 114 ug/L in the flow-through test. Time to hatch and percent survival at
hatch were not affected by the nonylphenol concentrations tested; however, nearly all of the
larvae were abnormal at the two highest exposure concentrations (> 53.0 ug/L). At the end of
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the test, survival was significantly reduced at concentrations > 23.1 ug/L but not at 10.3 ug/L.
Growth (both weight and length) was a more sensitive chronic endpoint than survival. At the
end of the test, the fish were significantly shorter (14 %) and weighed less (30 %, dry weight)
than control fish at nonylphenol concentrations > 10.3 ug/L, but not at 6.0 ug/L. Based on
growth, the NOEC and LOEC determined in this study were 6.0 and > 10.3 ug/L, respectively.
The chronic value, calculated as the geometric mean of the NOEC and the LOEC, is 7.861 ug/L
for rainbow trout. Dividing the acute value (221 ug/L), determined in in a companion test for
this species (Table 1) by the chronic value (7.861 ug/L; Table 2) results in an ACR of 28.11 for
rainbow trout.
An early-life-stage toxicity test was available for the fathead minnow, Pimephalespromelas
(Ward and Boeri 1991c). Embryos and larvae were exposed under continuous-flow conditions
for a total of 33 days to five concentrations of nonylphenol that ranged from 2.8 to 23 ug/L.
Embryos in the control group and those in the three lowest nonylphenol exposure concentrations
(2.8, 4.5, and 7.4 ug/L) began to hatch on the third day of exposure, while the two higher
concentrations (14 and 23 ug/L) began hatching on the fourth day. Growth (length or weight) of
nonylphenol exposed fish was not significantly different from the control organisms at any of the
nonylphenol treatment concentrations. Survival of the fish at the end of the test was significantly
reduced at nonylphenol concentrations > 14 ug/L. Fish survival averaged 56.7 % at 23 ug/L
nonylphenol, 66.7 % at 14 ug/L nonylphenol, and 76.7 % at 7.4 ug/L nonylphenol. At
concentrations < 7.4 ug/L survival of nonylphenol exposed fish did not differ from the control
fish survival, which averaged 86.7 %. Based on survival, the NOEC and LOEC determined in
this study were 7.4 ug/L and 14 ug/L, respectively. The chronic value, calculated as the
geometric mean of the NOEC and the LOEC, is 10.18 ug/L for fathead minnow (Table 2). A
companion acute toxicity test was not conducted; therefore, an ACR can not be calculated for
this species.
3.2. Saltwater
Two chronic toxicity tests with phenol have been conducted with the same saltwater animal
species. A 28-day chronic toxicity test was conducted with mysids, Americamysis bahia (Ward
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and Boeri 1991b). There was no effect on survival or reproduction at 6.7 ug/L, but there was a
18 % reduction in survival and a 53% reduction in reproduction at 9.1 ug/L. Effects on survival
at the highest concentration tested (21 ug/L) were observed before the end of the third week of
the test. Test organisms of each sex were measured separately for length and weight. The data
show no obvious difference between the length of male and female mysids for all of the
concentrations tested, therefore growth analysis was based on combined length data for both
sexes. Growth (length) was the most sensitive endpoint for mysids. There was a 7% (statistically
significant relative to control animals) reduction in the length of mysids exposed to 6.7 ug/L
nonylphenol, but no difference in growth for mysids exposed to 3.9 ug/L nonylphenol. Based on
growth, the NOEC and LOEC determined in this study were 3.9 ug/L and 6.7 ug/L, respectively.
The chronic value, calculated as the geometric mean of the NOEC and the LOEC, is 5.112 ug/L
for mysids (Table 2). Dividing the acute value (43 ug/L), determined in in a companion test for
this species (Table 1) by the chronic value (5.112 ug/L; Table 2) results in an ACR of 8.412 for
the mysid, Americamysis bahia.
A second 28-day life cycle test with mysids, Americamysis bahia (Kuhn et al. 2001) was
conducted using the ASTM standardized life-cycle test methods. Time to first brood release
appeared dose dependent, but was not statistically significant. Growth of the female mysid was
dose dependent and was significantly affected at concentrations > 27.56 ug/L. The most
sensitive endpoint for this test was a reproduction. The average number of young per available
female reproductive days was significantly reduced at test concentration > 15.28 ug/L, but was
not affected at 9.46 ug/L. Based on reproduction, the NOEC and LOEC determined in this study
were 9.46 ug/L and 15.28 ug/L, respectively. The chronic value, calculated as the geometric
mean of the NOEC and the LOEC, is 12.02 ug/L. A companion acute toxicity test was not
conducted: therefore, an ACR can not be calculated from this test.
3.3. Acute-Chronic Ratios
Three nonylphenol ACRs, determined from the fourth (Daphnia magna) and eighth (rainbow
trout) most sensitive freshwater species tested and the third (mysid) most sensitive saltwater
species tested, are available (Table 3). Two ACRs (0.6586 and 6.211; Table 2) were available
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for the cladoceran Daphnia magna, which differed by a factor of approximately 9.4-fold. The
species mean ACR, calculated as the geometric mean of the two values, is 2.023. Acute-chronic
ratios for the acutely sensitive mysid, A. bahia., was 8.412 and the moderately acutely sensitive
rainbow trout, Oncorhynchus mykiss, was 28.11. An ACR of 0.515 would be calculated from
the tests of England (1995) with the cladoceran, Ceriodaphnia dubia. However, the organisms
were fed during the acute test and data demonstrating that feeding did not significantly affect the
acute value were not available. According to the Guidelines, acute tests with this species must
be done without food present in the test solutions. Therefore, the C. dubia ACR was not used.
The three valid species mean ACRs (2.023, 8.412 and 28.11) differed by 13.9-fold (Table 3).
The Guidelines stipulate that if the species mean acute-chronic ratio seems to increase or
decrease as the SMAV increases, the Final Acute-Chronic Ratio (FACR) should be based on the
acute-chronic ratios for species whose SMAVs are close to the Final Acute Value (FAV).
Examination of the SMACRs (Table 3) relative to the SMAVs indicates that the more acutely
sensitive species (A. bahia and D. magna) have 3 to 14-fold lower SMACRs than for the less
acutely sensitive rainbow trout, indicating a general trend of increasing SMACR with increasing
SMAV. Therefore, the FACR should be based on the SMACR for species whose SMAVs are
close to the FAV. The mysid SMAV (51.05 ug/L) is closest to both the freshwater FAV (55.49
ug/L) and the saltwater FAV (13.93 ug/L). Therefore, the SMACR for the mysid is used as the
FACR and is 8.412.
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4. TOXICITY TO AQUATIC PLANTS
4.1. Freshwater
Acceptable data on the toxicity of nonylphenol to freshwater plants were available for one
species of algae (nonvascular plant) and no acceptable toxicity data are available for vascular
plants (Table 4). Ward and Boeri (1990a) exposed the green alga, Selenastrum capricornutum,
to nonylphenol for four days. They calculated an EC50 of 410 ug/L based on cell counts. At the
end of the toxicity test, algae from the highest exposure concentration (720 ug/L) were
transferred to fresh media solution. During the next seven days, cell counts increased
exponentially, indicating that nonylphenol treatment at this concentration for four days did not
have a persistent algistatic effect.
4.2. Saltwater
Acceptable data on the toxicity of nonylphenol to saltwater plants were available for one
species of marine algae and no acceptable toxicity data are available for vascular plants (Table
4). The EC50 value for vegetative growth of the planktonic diatom, Skeletonema costatum, was
27 ug/L (Ward and Boeri 1990d). Although this value was lower than nearly all of the acute
values for animals, it is for vegetative growth, which can recover rapidly. Skeletonema
transferred from the highest nominal concentration of nonylphenol with survivors (120 ug/L)
into control medium grew to a 76-fold increase in cells/mL within 48 hr (Ward and Boeri
1990d), demonstrating that nonylphenol treatment at this concentration for four days did not
have a persistent algistatic effect.
Based on the vegetative growth test using the saltwater planktonic diatom, Skeletonema
costatum, the Final Plant Value for nonylphenol is 27 ug/L. This plant species is more sensitive
to nonylphenol than any tested species of freshwater animal and more sensitive than all but one
tested saltwater animal species.
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5. BIOACCUMULATION
5.1. Freshwater
Acceptable data on the bioconcentration of nonylphenol in two species of freshwater animals
were available (Table 5). Ward and Boeri (1991a) measured the whole body concentrations of
nonylphenol in juvenile fathead minnows at two exposure concentrations after 27 days of
exposure. The bioconcentration factors (BCFs) were 271 and 344 (not lipid normalized)
following exposure to 4.9 and 22.7 ug/L nonylphenol, respectively. Brooke (1993b) exposed
juvenile fathead minnow (Pimephalespromelas) and juvenile bluegill (Lepomis macrochirus) to
nonylphenol at five concentrations for four and twenty-eight days. Lipid concentrations were
measured (Brooke 1994) for the test fish and the bioconcentration results were lipid normalized
which reduced the bioconcentration factors from 4.7 to 4.9 times. Nonylphenol concentrations
that proved lethal to the organisms were not used to calculate bioconcentration factors. The
short-term (4-day) tests showed that tissue concentrations reached steady-state within two days
in both the fathead minnow and the bluegill. Therefore, there was good agreement between the
results obtained in the 4-day and 28-day tests. Lipid-normalized BCFs for the fathead minnow
ranged from 128.3 to 209.4 (Table 5) and for the bluegill ranged from 38.98 to 56.94. Giesy et
al. (2000) measured the nonylphenol concentrations in whole bodies of the fathead minnow
following a 42-day exposure. Exposure to sublethal concentrations of nonylphenol ranging from
0.4 to 3.4 ug/L resulted in BCFs ranging from 203 to 268
5.2. Saltwater
Bioconcentration factors are available for three species of saltwater animals, Mytilus edulis,
Crangon crangon and Gasterosteus aculeatus (Ekelund et al. 1990;Table 5). Crangon crangon
is a non-resident species, but the data are included because so little bioaccumulation data are
available. Organisms were exposed to 14C-labeled nonylphenol (CAS number not provided) for
16 days followed by an elimination period of 32 days. Lipid-normalized BCFs based on wet
weight ranged from 78.75 for C. crangon to 2,168 forM edulis. The BCF forM edulis was an
estimate, because steady-state tissue concentration was not reached during 16 days of exposure.
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6. OTHER DATA
6.1. Freshwater
Additional data on the lethal and sublethal effects of nonylphenol on freshwater species that
do not meet the data requirements described in the Guidelines (Stephan et al. 1985) for use in
deriving aquatic life ambient water quality criteria are summarized in Table 6.
Three plant species (Chlamydomonas reinhardtii, Salvinia molesta and Lemna minor) were
exposed in studies using media solutions that were not described. The effect levels determined
for Salvinia molesta (2,500 ug/L) and Chlamydomonas reinhardtii (6,250 ug/L), indicates that
these plant species are less sensitive to nonylphenol than animals (Prasad 1986; Weinberger and
Greenhalgh 1984). Effect concentrations reported for the duckweed, Lemna minor, were highly
variable, ranging from 125 to 5,500 ug/L (Weinberger and lyengar 1983; Prasad 1986).
Protozoa were affected in the concentration range from 50 to 747 ug/L (Preston et al. 2000,
Schultz 1997, Yoshioka 1985).
Additional data on acute and chronic toxicity of a variety of invertebrates are summarized in
Table 6. McLeese et al. (1980b) reported an LC50 of 5,000 ug/L for a clam, Anodonta
cataractae, following a 144-hr exposure. The test organisms were fed in this test and the
toxicity value is higher than those reported in Table 1 for similar species. In an acute test (96-hr)
in which the cladoceran, Daphnia magna was fed, effect levels were reported as 136 and 302
ug/L for young and adult animals, respectively (Gerritsen et al. 1998). In a 48-hr test with the
same species, ECSOs ranged from 234 to 337 (Zang et al. 2003). Three 21-day tests with
Daphnia magna (Baer and Owens 1999, Baldwin et al. 1997, LeBlanc et al. 2000), an additional
D. magna test of 35-day duration in a high-hardness medium (Zang et al. 2003), and a 30-day
test with D. galeata mendotae (Shurin and Dodson 1997) are included in this section because
nonylphenol concentrations in the test water were not measured in these chronic tests. Negative
effects on survival or reproduction were observed in all three tests with typical water hardness
(i.e., between 25 and 200 ug/L). The results from these tests with D. magna (Table 6) agree
reasonably well with those from tests with D. magna in which nonylphenol concentrations were
measured (Table 2). Another cladoceran, Daphniapulex, was exposed for 48 hr in tests in which
nonylphenol concentrations decreased more than 50 percent during the exposures (Ernst et al.
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1980). The LCSOs determined in the test ranged from 140 to 190 ug/L, which agreed with
LCSOs for other cladoceran species. The cladoceran, Ceriodaphnia dubia, gave similar LC50
results of 276 and 225 ug/L following exposure to nonylphenol for 48 hr and 7 days, respectively
(England 1995). The LC50 values reported in this table for the species are slightly higher than
the chronic value for the species of 134 ug/L (Table 2). England and Bussard (1993) reported an
EC50 and an LC50 for larva of the midge, Chironomus tentans, of 95 and 119 ug/L,
respectively. These values, determined when the organisms were fed, are less than the than
values reported in another study by the same authors in which organisms were not fed during the
test (Table 1).
In a pair of tests in which the test organisms were fed, Brooke (1993b) measured a 96-hr
LC50 for the fathead minnow, Pimephalespromelas, of 138 ug/L and a 96-hr LC50 for the
bluegill, Lepomis macrochirus, of 135 ug/L. The LC50 values for these species from tests in
which the fish were fed, agree well with data from tests in which the fish were not fed (Table 1).
McLeese et al. (1980b) reported an LC50 of 900 ug/L for the Atlantic salmon, Salmo salar, in a
96-hr exposure and Lech et al (1996) reported an LC50 of 193.65 for rainbow trout,
Oncorhyncus mykiss, in a 72-hr exposure. Holmes and Kingsbury (1980) reported a 96-hr LC50
of 145 ug/L for brook trout juveniles (Salvelinus fontinalis), a 96-hr LC50 of 230 ug/L for
rainbow trout juveniles (Oncorhynchus mykiss) and a 32-day LC50 of > 40 ug/L for lake trout
juveniles (Salvelinus naymaycush). Fish were fed during these studies, but the resulting toxicity
values are similar to comparable studies reported for salmonids in Table 1. Ernst et al (1980)
reported 96-hr LCSOs ranging from 560-920 ug/L for rainbow trout exposed to practical grade
nonylphenol. A number of older studies were identified that report time to lethality (LT100)
values for a number of freshwater species exposed to very high concentrations of nonylphenol
(Applegate et al. 1957; MacPhee and Ruelle 1969; Wood 1952)
A long-term study was conducted with rainbow trout, Onchorynchus mykiss, exposing
female fish immediately after hatch to 1, 10, and 30 or 50 ug/L of nonylphenol (Ashfield et al.
1998). They found reduced growth in fish exposed to 1 ug/L for 22 days and grown for 86 days
beyond treatment. Growth was not reduced in the 10 ug/L treatment but was in the 50 ug/L
treatment. In a second study in which exposure was for 35 days and grow-out was for 431 days
beyond the last treatment day, reduced growth was observed at the 10 and 30 ug/L treatments on
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day 55 of the study, but not at the 1 ug/L. At day 466, the fish exposed to 10 ug/L recovered the
growth reductions seen earlier and only the 30 ug/L exposed fish showed reduced
(approximately 25%) growth.
Five fish species (rainbow trout, Lahontan cutthroat trout, Apache trout, Colorado squawfish
and fathead minnow) were exposed to nonylphenol for 96 hr to determine if nonylphenol
inhibited brain acetylcholoinesterase enzymes. AChE inhibition was measured indirectly as a
decrease in the number of muscarinic cholinergic receptors which is a compensatory response to
acetycholine buildup (Jones et al. 1998). Responses at exposure concentrations < 220 ug/L were
observed in the rainbow trout, Lahontan cutthroat trout and Apache trout. The lack of a clear
connection between this sublethal biochemical endpoint and population relevant effects
precludes the use of these results as core data. An effect of nonylphenol on another sub-
organismal endpoint, histology of epidermal mucous cells, was observed following intermittent
exposure to technical grade nonylphenol (Burkhardt-Holm et al. 2000). Other histochemical or
biochemical changes have been reported following exposure to nonylphenol including
hemorrhage and lymphocyte infiltration in liver tissue of rainbow trout (Ugaz et al. 2003) and
blood cell composition in carp (Schwaiger et al. 2000).
Brooke (1993b) measured the bioconcentration of nonylphenol in the fathead minnow and
bluegill at concentrations near lethality. The fathead minnow BCF was 100.4 and the bluegill
BCF was 35.31. The values were slightly less than the BCFs measured in the fish from lower
exposure concentrations (Table 5). Blackburn et al. (1999) reported BCFs for adult male
rainbow trout of 116 and 88 following 3 weeks exposure to 63 and 81 ug/L nonylphenol (purity
unknown), respectively. Lewis and Lech (1996) found that bioconcentration of nonylphenol
after short-term exposure (2-24 hr) was higher in rainbow trout viscera (BCF = 98.2) than in the
remainder of the carcass (BCF = 24.21). They also measured the half-life of nonylphenol in
various tissues and found that fat and muscle similarly depurated nonylphenol to half
concentrations in about 19 hr. The liver depurated to half concentrations in about 6 hr.
Mesocosm studies were conducted with nonylphenol in which zooplankton, benthic
macroinvertebrates, and fish were observed for effects. Zooplankton populations (O'Halloran et
al. 1999) and benthic macroinvertebrate populations (Schmude et al. 1999) exposured to four
concentrations of nonylphenol for 20 days showed no negative effects at the 23 ug/L
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nonylphenol but were negatively affected at 76 ug/L nonylphenol. Various species of
zooplankton and macroinvertebrates exhibited differences in sensitivity to nonylphenol. The
authors of the zooplankton study stated that a MATC for the protection of all zooplankton taxa is
approximately 10 ug/L. The fish (bluegill) in the mesocosms (Liber et al. 1999) were unaffected
at nonylphenol exposures < 76 ug/L, but survival was reduced at 243 ug/L. In one exposure
replicate with a mean nonlyphenol concentration of 93 ug/L, survival of the fish was reduced
after 20 days of exposure indicating that concentrations near 100 ug/L may be the toxicity
threshold for this species. Hense et al. (2003) and Severin et al. (2003) conducted microcosm
studies in Germany using 6-week exposures to nonylphenol. They found changes in
phytoplankton species composition, but no change in biomass with nonylphenol concentrations
up to 120 ug/L. The zooplankton in the study were not affected at concentrations ranging from
19 to 44 ug/L (mean of 30 ug/L), but species richness was affected at concentrations >30 ug/L.
Effects observed in these mesocosm studies were all above the freshwater Final Chronic Value
of 5.920 ug/L.
6.2. Saltwater
Additional data on the lethal and sublethal effects of nonylphenol on saltwater species that do
not meet the data requirements described in the Guidelines (Stephan et al. 1985) for use in
deriving aquatic life ambient water quality criteria are summarized in Table 6.
Results from a sexual reproduction test with red alga species, Champiaparvula, indicated
that reproduction was not inhibited at the highest measured concentration tested, 167 ug/L
(Tagliabue 1993). Cypris larva of the barnacle, Balanus amphitrite, were exposed to
nonylphenol for 48 hr and the settlement of the larva was reduced at 1.0 ug/L (Billinghurst et al.
1998). The soft-shell clam, Mya arenaria, showed no adverse effects on survival from a 360-hr
exposure at 700 ug/L (McLeese et al. 1980b). Granmo et al. (1989) report LCSOs of 3,000 ug/L
and 500 ug/L at 96-hr and 360-hr, respectively, for the blue mussel, Mytilus edulis. Nonylphenol
also reduced growth and byssus thread strength in the blue mussel at concentrations of > 56 ug/L
(Granmo et al. 1989) and caused effects on attachment activity at 22 ug/L (Etoh et al. 1997).
Lussier et al. (2000) tested a number of saltwater invertebrates including coot clam, mysid,
amphipod, grass shrimp, and American lobster and determined LCSOs for various timepoints
23
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(Table 6). Results from other studies with mysid (Ward and Boeri, 1990a) and American lobster
(McLeese et al 1980b) are similar to those reported by Lussier et al (2000). The LC50 value
(300 ug/L) reported by McLeese et al. (1980b) for the shrimp, Crangon septemspinosa, is higher
than the grass shrimp values reported by Lussier et al (2000). Kusk and Wollenberger (1999)
determined a 48-hr LC50 (280-360 ug/L for the copepod, Acartia tonsa, exposed to nonylphenol
in a synthetic media. Nice et al (2000) reported developmental effects at 100 ug/L nonylphenol
on the Pacific oyster (Crassostrea gigas) exposed for 72-hr. Nonylphenols have also been
reported to have antifouling activity, but the test results are qualitative (Takasawa et al. 1990;
Kitajima et al. 1995).
A fifty-five-day flow-through test with the mysid, Americamysis bahia, was conducted by
Kuhn et al. (2001) to evaluate the efficacy of an age-classified projection matrix model for
predicting population behavior. Organisms were exposed for more than three generations to
nonylphenol. The measured mean concentrations of nonylphenol used for the 55-day exposure
were 5.79, 7.56, 10.88, 15.75, 21.44, 33.19, and 106.00 ug/L. Thirty individuals were used in
each replicate exposure chamber and the age distribution consisted of 15 (24-h newly hatched), 8
(8-d-old juveniles), 4 (17-d-old adults), 2 (23-d-old adults), and 1 (31-d-old adult) test
organisms. Several generations were possible in this test (control organisms produced first
brood in 14 days). It appears that the control populations grew in number of individuals for the
first 28 to 36 days, then stabilized. Population growth was reduced from day 8 and beyond in all
of the nonylphenol treated groups. The population exposed to 5.79 ug/L nonylphenol grew at
the same rate as the control animals for the first 21 days, but then the rate fell below the control
rate. The populations exposed to higher nonylphenol concentrations all decreased from day 8
and beyond. There appears to be a trend (not significant) in the shift in the sex ratios for the
various treatments. At the end of the test, the sex ratios were one-third female in the control
groups and half female in the 33.19 ug/L exposure group. The authors calculated a zero
population growth value (X) of 19 ug/L for the 55-day multigenerational test. The chronic values
from the 28-day exposure in this study was 12.02 ug/L (Table 2) and from a similar 28-day study
by Ward and Boeri (1991b) was 5.112 ug/L (Table 2), which are lower than the predicted value
for "population protection" from the 55-day multigenerational test.
McLeese et al. (1980b) reported 96-hr test results for the Atlantic salmon, Salmo salar, that
24
-------
were in general agreement with freshwater trout test results. In four tests, LC50 values ranged
from 130 to 900 ug/L. Ward and Boeri (1990c) found similar toxicity results for sheepshead
minnow, Cyprinodon variegatus, exposed in brackish water as those reported for salt water
(Table 1). In brackish water, LCSOs ranged from > 420 ug/L for a 24-hr exposure to 320 ug/L
for a 72-hr exposure. Threespine stickleback, Gasterosteus aculeatus, exposed to a commercial
mixture of nonylphenol had a 96-hr LC50 of 370 ug/L (Granmo et al. 199la). Killifish (Kelly
and Di Giulio 2000) were exposed as embryos and larva to nonylphenol for 96 hrs. Even though
the solvent concentration used in the exposures exceeded the 0.5 mL/L recommended limit, the
data are included in Table 6 because the results reported for the solvent controls do not show
decreased hatching success or increased abnormalities at 10 days post-hatch. Embryos exposed
to 2,204 ug/L for 96 hr were all abnormally developed at 10 days post-fertilization. The LC50
for the same exposure period was 5,444 ug/L. Killifish larva were similar in sensitivity to
nonylphenol exposures at post hatch ages of 1, 14, and 28 days with LCSOs of 214, 209, and 260,
respectively.
Additional data on the effect of nonylphenol on saltwater species do not indicate greater
sensitivities than the data summarized in Tables 1 and 2. Some of the data presented in Table 6
(e.g., sheepshead minnow, Inland silversides) were from the same acute tests listed in Table 1
(Lussier et al. 2000; Ward and Boeri 1990a,b), but for exposure durations other than 96 hr.
25
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6.3. Reproductive, Devleopmental and Estrogenic Effects of Nonylphenol
There are several review articles that describe the estrogenic activity of nonylphenol (Servos
1999; Sonnenschein and Soto 1998; Sumpter 1998). The majority of studies describing the
estrogenic activity of nonylphenol using aquatic species models exposed in vivo measure
molecular, biochemical, or histological endpoints such as induction of the egg protein,
vitellogenin, or occurrence of egg cells within testes (a condition known as intersex or ovo-
testis). In addition, estrogenicity is commonly characterized using in vitro studies such as
estrogen receptor binding assays. Molecular, biochemical and reproductive endpoints measured
following in vivo exposures to nonylphenol and that are thought to result from estrogenic
activity of nonylphenol are summarized in this section. In vitro studies are listed in Section 7 of
this document.
The majority of reports of estrogenic effects in aquatic organism have been for fish, although
some effects in invertebrates have also been reported. Bechmann (1999) found no effects in the
marine copepod Tisbe battagliai exposed to nonylphenol at 55 ug/L, but estrogenic effects were
reported to have occurred in the amphipod Corophium volutator (Brown et al. 1999) at 10 ug/L
and in the larva of Chironomus riparius (Hahn et al. 2002) at 2,000 ug/L. The mechanism(s) by
which estrogenic effects can be produced in invertebrates that do not possess estrogen receptors
is unclear.
Vitellogenin is a protein produced in the liver of female oviparous vertebrate species and
deposited in the ovaries as the primary material for yolk in the ova. Male fish normally produce
very little vitellogenin. Islinger et al. (1999) estimated the estrogenic potential of nonylphenol to
stimulate vitellogenin production in male rainbow trout at 2,000 to 3,000 times less potent than
the natural estrogen, l?p-estradiol. Ren et al. (1996a) demonstrated significant increases in
vitellogenin production in rainbow trout exposed to nonylphenol at 100 ug/L for 72 hr. In
another study, Ren et al. (1996b) demonstrated that nonylphenol could stimulate the production
of vitellogenin mRNA (which precedes vitellogenin protein synthesis) within 4 hr at 10 ug/L.
Similarly, Lech et al. (1996) observed a significant increase in vitellogenin mRNA at 72 hr in
rainbow trout at 14.14 ug/L nonylphenol. Vitellogenin was induced in green swordfish,
Xiphophorus helleri, by exposure to 4 ug/L of technical grade nonylphenol (Kwak et al. 2001).
Jobling et al. (1996) demonstrated significant increases in vitellogenin in male rainbow trout,
26
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Oncorhynchus mykiss, at three weeks of exposure to 20.3 and 54.3 ug/L of nonylphenol. Similar
results were reported in another study with rainbow trout, plasma vitellogenin was increased
after 21 days exposure to 50 ug/L of nonylphenol (Tremblay and Van Der Kraak 1998). Harris
et al. (2001) also observed increased plasma vitellogenin levels in female rainbow trout exposed
to 8.3 and 85.6 ug/L of nonylphenol. In the same study, nonylphenol also caused changes in
several pituitary and plasma hormone levels. In contrast, vitellogenin induction was not
observed in rainbow trout exposed for 9 days to 109 ug/L of nonylphenol (Pedersen et al. 1999)
or in Atlantic salmon, Salmo tmtta, exposed for 30 days to 20 ug/L (Moore et al. 2003). The
influence of exposure route on nonylphenol-induced vitellogenin wRNA and plasma vitellogenin
production in the male fathead minnow was studied by Pickford et al. (2003). Their results
showed that exposure via water produced 10-fold higher vitellogenin induction than exposure via
the dietary route.
In a study with the fathead minnow, Giesy et al. (2000) found that nonylphenol exposures to
0.5 to 3.4 ug/L nonylphenol were not acutely toxic to the adult fish and fecundity was variously
affected over the reproductive season. When the cumulative reproduction was combined for the
two experiments during different portions of the reproductive season, concentrations of > 0.3 to
0.4 ug/L did appear to reduce fecundity. However, fish exposed to 0.09 and 0.1 ug/L produced
more eggs than control fish. These data appear to produce a U-shaped dose-response and
indicate a possible hermetic response of fecundity to nonylphenol. Nonylphenol concentrations
of 0.05 to 3.4 ug/L did not significantly change vitellogenin concentrations in the blood of
males, and raised the l?p-estradiol liters in the blood of male and female fish at most treatment
concentrations > 0.05 ug/L. An increase in the number of Sertoli cells may have occurred in the
male fathead minnow exposed to nonylphenol at 1.6 ug/L for 42 days (Miles-Richardson et al.
1999). The evidence was not complete, but indicated the possibility of increased phagocytic
action and Sertoli cell tissue in testes.
A non-resident fish species, Japanese medaka (Oryzias latipes\ was exposed to nonylphenol
for 28 days following hatch and survivors monitored for the following 55 days (Nimrod and
Benson 1998). At the highest exposure concentration of 1.93 ug/L, survival, growth, egg
production, egg viability, and gonadosomatic index (GSI) were not altered. In another study
with the same species offish, development of ovo-testis, an intersex condition, occurred after a
27
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three month exposure to 50 ug/L of nonylphenol (Gray and Metcalf 1997). The sex ratio shifted
in favor of females at the highest exposure concentration. Seki et al. (2003) found that in the
same species offish exposed to nonylphenol from fertilized egg to 60 days post-hatch, the
lowest-observed-effect concentration for vitellogenin induction was 11.6 ug/L.
Yokota et al. (2001) conducted a two-generation flow-through study with the non-resident
fish species medaka (Oryzias latipes). Concentrations of nonylphenol were measured during
exposures that began with eggs and proceeded to 60-day s post-hatch of the second (Fi)
generation. Five exposure concentrations of nonylphenol in quadruplicate (4.2, 8.2, 17.7, 51.5,
and 183 ug/L) and water-only and solvent controls were used. In the FO generation, egg
hatchability was reduced (46.7%) by 183 ug/L nonylphenol exposure, survival was significantly
decreased at 60 days post-hatch by nonylphenol exposures > 17.7 ug/L, and no differences in
growth (length or weight) were observed at 60 days post-hatch. Induction of ovo-testis was
observed in the 17.7 ug/L treatment, with 20% offish displaying male characteristics externally
having ovo-testis tissues. In fish from the 51.5 ug/L treatment, 40% had ovo-testis and all of
these fish exhibited female characteristics externally. Spermatogenesis was observed in the fish
with ovo-testis exposed to 17.7 ug/L nonylphenol, but was not observed in the fish with ovo-
testis exposed to 51.5 ug/L nonylphenol. Fecundity of paired fish during the reproductive phase
(days 71 to 103 post-hatch) was not affected by nonylphenol treatments. GSI of male fish was
reduced at 17.7 ug/L, but not significantly, and GSI of female fish was increased significantly by
exposure to nonylphenol concentrations > 8.2 ug/L.
The effects of nonylphenol on FI fish from Yokota et al. (2001) were also reported. No
embryological abnormalities or hatching failures of fertilized eggs were observed in any
treatments. Growth was not affected at 60-day s post-hatch by any of the nonylphenol exposure
concentrationsl. The sex ratio, characterized by secondary sex characteristics, changed in
treatments > 17.7 ug/L to favor females 1:2. Induction of ovo-testis was observed at lower
concentrations of nonylphenol in the FI generation than in the F0 generation. Ovo-testis were
observed in the 8.2 ug/L exposure group (10%) and in the 17.7 ug/L exposure group (25%).
However, all fish with ovo-testis displayed external male characteristics and the degree of
development of oocytes in each ovo-testis was not as severe as that in the FO generation in the
17.7 ug/L treatment. The overall results indicate a LOEC of 17.7 ug/L and a NOEC of 8.2 ug/L
28
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with a chronic value, calculated as the geometric mean of the NOEC and LOEC, of 12.05 ug/L.
The chronic value for this study is in good agreement with the Table 2 data for resident
freshwater species, rainbow trout and fathead minnow.
A multi-generational exposure to nonylphenol has been conducted with rainbow trout by
Schwaiger et al. (2002). Adult rainbow trout of both sexes were exposed intermittently to
nonylphenol at 1 and 10 ug/L over a 4 month period. Mortality rate in the progeny was
significantly reduced by parental exposure to both 1 and 10 ug/L nonylphenol and hatching in
the progency was reduced by parental exposure to 10 ug/L nonylphenol. Vitellogenin was
induced (approximately 10-fold) in adult male fish exposed to both 1 and 10 ug/L nonylphenol.
In the male progeny of parental fish exposed to 10 ug/L nonylphenol, no effects were observed
on plasma vitellogenin or testosterone concentrations, but plasma estradiol concentrations were
elevated. In the female progeny of the same parental fish, plasma vitellogenin and plasma
testosterone concentrations were elevated, but plasma estradiol concentrations were not different
from control levels. Testicular tissue of nonylphenol exposed adult male fish was not affected
by nonylphenol. Sex ratios of the offspring of exposed fish were also unaffected by parental
exposure to nonylphenol.
As summarized in this section, the ability of nonylphenol to induce estrogenic effects has
seldom been reported at concentrations below the freshwater Final Chronic Value of 6.5965
ug/L.
29
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7. UNUSED DATA
Data from some studies were not used in this document, as they did not meet the criteria for
inclusion as specified in the Guidelines (Stephan et al. 1985). The reader is referred to the
Guidelines for further information regarding these criteria.
Results were not used when the test organism is not resident to North America (Gross-
Sorokin et al. 2003; Yoshimura 1986). Tsuda et al. (2000) measured tissue concentrations from
feral fish, but water concentrations greatly varied.
Test Organism or Test Material were Not Adequately Described
Folmar et al. (1998) Magliulo et al. (1998) Weinberger and Rea (1981)
Hansen et al. (1998) Muller (1980)
Kopf (1997) Palmer et al. (1998)
Nonylphenol was a Component of a Mixture or Sediment
Ahel et al. (1993) Escher et al. (1999) Sundaram et al. (1980)
Amato and Wayment (1998) Hansen et al. (1999) Turner et al. (1985)
Bettinetti and Provini (2002) Larsson et al. (1999) Ward and Boeri (1992)
Fay et al. (2000) Moore et al. (1987)
Dwyer et al. (1999a,b) Purdom et al. (1994)
Studies were Conducted with Ethoxylated Nonylphenols
Baldwin et al. (1998) Dorn et al. (1993) Manzano et al. (1998, 1999)
Braaten et al. (1972) Maki et al. (1998) Patoczka and Pulliam 1999
30
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Organisms were Dosed by Injection, Gavage or in Artifical Medium
Arukwe et al. (1997a,b;1998) Madsen et al. (1997) Thibaut et al. (1998)
Christiansen et al. (1998a,b,c; Nimrod and Benson (1996; Weinberger et al. (1987)
1999) 1997) Yadetieetal. (1999)
Coldham et al. (1997, 1998) Rice et al. (1998)
Haya et al. (1997) Spieser et al. (1998)
Experimental Model was Plasma, Enzymes, Receptors, Tissues or Cell Cultures
Andersen et al. (1999) Lamche and Burkhardt-Holm Routledge and Sumpter (1996,
Celiusetal. (1999) (2000) 1997)
Flouriot et al. (1995) Levine and Cheney (2000) Soto et al. (1991, 1992)
Hewitt et al. (1998) Loomis and Thomas (1999) White et al. (1994)
Jobling et al. (1996) Lutz and Kloas (1999)
Jobling and Sumpter (1993) Milligan et al. (1998)
Knudsen and Pottinger (1999) Petit et al. (1997, 1999)
Data were Compiled from Other Source
Bearden and Schultz (1997, Lewis (1991) Varma and Patel (1988)
1998) Liber etal. (1999) Veith and Mekenyan( 1993
-------
8. SUMMARY
8.1. Freshwater Data
Acute toxicity of nonylphenol was tested in 18 freshwater species and 2 subspecies from 15
genera (Figure 1 and Table 3). Species Mean Acute Values (SMAV) ranged from 55.72 ug/L for
the amphipod Hyalella azteca to 774 ug/L for the snail Physella virgata. Eleven species offish
were tested and were in the mid-range of sensitivity (SMAVs = 110 to 289.3 ug/L) of tested
species. The four most sensitive tested freshwater species were comprised of two invertebrate
species and two vertebrate species (Figure 1). No relationships have been demonstrated between
nonylphenol toxicity and water quality characteristics such as hardness and pH. The freshwater
Final Acute Value is 55.49 ug/L which is equal to the LC50 for the most sensitive tested species,
Hyalella azteca.
Chronic toxicity of nonylphenol was tested in 5 freshwater species from 5 genera (Figure 3
and Table 3). Two freshwater fish were tested; the rainbow trout, Oncorhynchus mykiss, had a
chronic value of 7.861 ug/L based on growth, and the fathead minnow, Pimephalespromelas,
had a chronic value of 10.18 ug/L based on survival. Two species of freshwater cladocerans
were tested and chronic values ranged from 22.62 to 157.9 ug/L based on reproduction. One
species of freshwater midge was tested and its chronic value was 61.82 ug/L based on survival.
Data were available to calculate a Final Acute-Chronic Ratio (FACR) for a freshwater
cladoceran, Daphnia magna, saltwater mysid, Americamysis bahia., and rainbow trout,
Oncorhynchus mykiss. The Final Acute-Chronic Ratio for nonylphenol was the ACR for A
bahia because SMARs increased with increasing SMAV and the SMAV for A. bahia is closest
to the freshwater and saltwater FAV. The FACR for nonylphenol is 8.412.
8.2. Saltwater Data
Acute toxicity of nonylphenol was tested in 11 saltwater species from 11 genera (Figure 2
and Table 3). Species Mean Acute Values (SMAV) ranged from 17 ug/L for the winter
flounder, Pleuronectes americanus, to 209.8 ug/L for the sheepshead minnow, Cyprinodon
variegatus. These two fish species were the only fish were tested. Nine different species of
32
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invertebrates were tested. The four most sensitive tested saltwater species were comprised of
three invertebrate species and one fish species (Figure 1). No relationships have been
demonstrated between nonylphenol toxicity and water quality characteristics such as hardness
and pH. The saltwater Final Acute Value is 13.93 ug/L.
Chronic toxicity of nonylphenol was tested on one saltwater species (Figure 3 and Table 3).
The saltwater species tested was the mysid, Americamysis bahia, which was also the most
sensitive of all species tested, both freshwater and saltwater. Two tests were available for A
bahia, with chronic values of 5.112 ug/L based on reduced growth and 12.02 ug/L based on a
reproductive endpoint.
Data were available to calculate a Final Acute-Chronic Ratio (FACR) for a freshwater
cladoceran, Daphnia magna, saltwater mysid, Americamysis bahia, and rainbow trout,
Oncorhynchus mykiss. The Final Acute-Chronic Ratio for nonylphenol was the ACR for A
bahia because SMARs increased with increasing SMAV and the SMAV for A. bahia is closest
to both the freshwater and saltwater FAV. The FACR for nonylphenol is 8.412.
8.3. Plant Data
Nonylphenol toxicity data for 2 species of aquatic plants, one freshwater alga and one
saltwater diatom, were available. Algae were as sensitive as animals, showing effect
concentrations that ranged from 27 ug/L for the freshwater alga to 410 ug/L for the saltwater
diatom. Based on the vegetative growth endpoint in saltwater planktonic diatom Skeletonema
costatum, the Final Plant Value for nonylphenol is 27 ug/L.
8.4. Bioaccumulation Data
Nonylphenol bioaccumulation in aquatic organisms is less than would be predicted from the
log Kow of nonylphenol. Nonylphenol is metabolized in animals which may account for the
lower than expected BCFs. In freshwater fish, lipid-normalized BCFs ranged from 39 to 209.
Bioaccumulation in saltwater organisms is apparently greater, with lipid-normalized BCFs 79 to
2,168.
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9. NATIONAL CRITERIA
9.1. Freshwater
The procedures described in the "Guidelines for Deriving Numerical National Water Quality
Criteria for the Protection of Aquatic Organisms and Their Uses" (Stephan et al. 1985) indicate
that, except possibly where a locally important species is very sensitive, freshwater aquatic
organisms and their uses should not be affected unacceptably if the one-hour average
concentration of nonylphenol does not exceed 28 ug/L more than once every three years on the
average and if the four-day average concentration of nonylphenol does not exceed 6.6 ug/L more
than once every three years on the average.
9.2. Saltwater
The procedures described in the "Guidelines for Deriving Numerical National Water Quality
Criteria for the Protection of Aquatic Organisms and Their Uses" (Stephan et al. 1985) indicate
that, except possibly where a locally important species is very sensitive, freshwater aquatic
organisms and their uses should not be affected unacceptably if the one-hour average
concentration of nonylphenol does not exceed 7.0 ug/L more than once every three years on the
average and if the four-day average concentration of nonylphenol does not exceed 1.7 ug/L more
than once every three years on the average.
34
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10. IMPLEMENTATION
As discussed in the Water Quality Standards Regulation (U.S. EPA 1983) and the Foreword
to this document, a water quality criterion for aquatic life has regulatory impact only after it has
been adopted in a state or tribal water quality standard. Such a standard specifies a criterion for
a pollutant that is consistent with a particular designated use. With the concurrence of the U.S.
EPA, states and tribes designate one or more uses for each body of water or segment thereof and
adopt criteria that are consistent with the use(s) (U.S. EPA 1994, 1987). In each standard a state
or tribe may adopt the national criterion, if one exists, or, if adequately justified, a site-specific
criterion (if the site is an entire state, the site-specific criterion is also a state-specific criterion).
Site-specific criteria may include not only site-specific criterion concentrations (U.S. EPA
1994), but also site-specific, and possibly pollutant-specific, durations of averaging periods and
frequencies of allowed excursions (U.S. EPA 1991). The averaging periods of "one hour" and
"four days" were selected by the U.S. EPA on the basis of data concerning how rapidly some
aquatic species react to increases in the concentrations of some pollutants, and "three years" is
the Agency's best scientific judgment of the average amount of time aquatic ecosystems should
be provided between excursions (Stephan et al. 1985; U.S. EPA 1991). However, various
species and ecosystems react and recover at greatly different rates. Therefore, if adequate
justification is provided, site-specific and/or pollutant-specific concentrations, durations, and
frequencies may be higher or lower than those given in national water quality criteria for aquatic
life.
Use of criteria, which have been adopted into state or tribal water quality standards, for
developing water quality-based permit limits requires selection of an appropriate wasteload
allocation model. Although dynamic models are preferred for the application of these criteria
(U.S. EPA 1991), limited data or other considerations might require the use of a steady-state
model (U.S. EPA 1986). Guidance on mixing zones and the design of monitoring programs is
also available (U.S. EPA 1987, 1991).
35
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1000-,
D)
c
o
"<3
2
4-1
C
-------
1000 T
D)
c
o
'•5
re
-------
Figure 3. Chronic Toxicity of Nonylphenol to Aquatic Animals
1000 q
D)
C
o
"•5
ffi
-------
Table 1. Acute Toxicity of Nonylphenol to Aquatic Animals
Species
Method" Chemical pH
LC50
or ECSO
fue/U
Species
Mean Acute
Valueb
fug/U
Reference
Annelid (adult), F,M
Lumbriculus variegatus
Snail (adult), F,M
Physella virgata
Cladoceran R,M
(<24-hr old),
Daphnia magna
Cladoceran S,M
(<24-hr old),
Daphnia magna
Midge (2nd instar), F,M
Chironomus tentans
Dragonfly (nymph), F,M
Ophiogomphus sp.
Amphipod, F,M
(juvenile, 2mm TL),
Hyalella azteca
Amphipod F,M
(juvenile, 2-3mm TL),
Hyalella azteca
Rainbow trout S,U
(0.67 ± 0.35 g),
Oncorhynchus mykiss
Rainbow trout S,U
(1.25 ± 0.57 g),
Oncorhynchus mykiss
Rainbow trout S,U
(0.27 ± 0.07 g),
Oncorhynchus mykiss
Rainbow trout S,U
(1.09 ± 0.38 g),
Oncorhynchus mykiss
FRESHWATER SPECIES
>90% 6.75 342
>90% 7.89 774
>90% 7.87 104
91.8% 8.25 190
>95% 8.0-8.4 160
342
Brooke 1993a
>90% 8.06
85%
7.9
596
>90% 7.80 20.7
>95% 7.9-8.7 150
85% 7.8-7.9 190
85% 7.5-7.7 260
140
85% 7.7-7.9 270
774 Brooke 1993a
Brooke 1993a
140.6 Comber et al.
1993
160 England and
Bussard 1995
596 Brooke 1993a
Brooke 1993a
55.72 England and
Bussard 1995
Dwyer et al.
1995
Dwyer et al.
1995
Dwyer et al.
1995
Dwyer et al.
1995
39
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Table 1. Acute Toxicity of Nonylphenol to Aquatic Animals
Species
Rainbow trout
(0.48 ± 0.08 g),
Oncorhynchus mykiss
Rainbow trout
(0.50 ± 0.21 g),
Oncorhynchus mykiss
Rainbow trout
(45 d),
Oncorhynchus mykiss
Apache trout
(0.85 ± 0.49 g),
Oncorhynchus apache
Apache trout
(0.38 ± 0.18g),
Oncorhynchus apache
Greenback cutthroat
trout (0.31 ± 0.17g),
Oncorhynchus clarki
stomais
Lahontan cutthroat
trout
(0.34 ± 0.08 g),
Oncorhynchus clarki
henshawi
Lahontan cutthroat
trout
(0.57 ± 0.23 g),
Oncorhynchus clarki
henshawi
Fathead minnow
(0.32 ±0.16 g),
Pimephales promelas
Fathead minnow
(0.56 ± 0.19g),
Pimephales promelas
Fathead minnow
(0.45 ± 0.35 g),
Pimephales promelas
Method"
S,U
s,u
F,M
S,U
S,U
S,U
S,U
S,U
S,U
s,u
s,u
Species
LC50 Mean Acute
or ECSO Valueb
Chemical pH ([ig/L) ([isfL) Reference
85% 7.5-7.9 160 - Dwyeretal.
1995
85% 6.5-7.9 180 - Dwyeretal.
1995
>90% 6.72 221 221 Brooke 1993a
85% 7.8-7.9 180 - Dwyeretal.
1995
85% 7.3-7.7 160 169.7 Dwyeretal.
1995
85% 7.5-7.6 150 - Dwyeretal.
1995
85% 7.9 140 - Dwyeretal.
1995
85% 7.6-7.7 220 166.6 Dwyeretal.
1995
85% 7.7-8.1 210 - Dwyeretal.
1995
85% 7.8-8.1 360 - Dwyeretal.
1995
85% 7.6-7.8 310 - Dwyeretal.
1995
40
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Table 1. Acute Toxicity of Nonylphenol to Aquatic Animals
Species Method"
Fathead minnow S,U
(0.40 ± 0.21 g),
Pimephales promelas
Fathead minnow S,U
(0.34 ± 0.24 g),
Pimephales promelas
Fathead minnow S,U
(0.39 ±0. 14 g),
Pimephales promelas
Fathead minnow F,M
(32 d),
Pimephales promelas
Fathead minnow F,M
(25-35 d),
Pimephales promelas
Bonytail chub S,U
(0.29 ± 0.08 g),
Gila elegans
Bonytail chub S,U
(0.52 ± 0.09 g),
Gila elegans
Colorado squawfish S,U
(0.32 ± 0.05 g),
Ptychocheilus lucius
Colorado squawfish S,U
(0.34 ± 0.05 g),
Ptychocheilus lucius
Razorback sucker (0.3 1 S,U
± 0.04 g),
Xyrauchen texanus
Razorback sucker (0.32 S,U
± 0.07 g),
Xyrauchen texanus
Gila topminnow S,U
(0.219 g, 27.2mm),
LC50
or ECSO
Chemical pH (jig/D
85% 7.5-7.9 330
85% 7.5-7.6 170
85% 7.8-8.2 290
99% 7.29 140
>90% 7.23 128
85% 7.7-7.9 270
85% 7.4-7.6 310
85% 8.1-8.2 240
85% 7.8-8.0 270
85% 7.8-8.1 160
85% 7.9-8.0 190
85% 8.0 230
Species
Mean Acute
Valueb
([iz/D Reference
Dwyer et al.
1995
Dwyer et al.
1995
Dwyer et al.
1995
Holcombe et
al. 1984;
Univ. Wisc.-
Superior 1985
133.9 Brooke 1993a
Dwyer et al.
1995
289.3 Dwyer et al.
1995
Dwyer et al.
1995
254.6 Dwyer et al.
1995
Dwyer et al.
1995
174.4 Dwyer et al.
1995
230 Dwyer et al.
1999a
Poeciliopsis
occidentalis
41
-------
Table 1. Acute Toxicity of Nonylphenol to Aquatic Animals
Species Method"
Fountain darter S,U
(0.062 g, 20.2 mm),
Etheostoma rubrum
Greenthroat darter S,U
(0. 133 g, 22.6mm),
Etheostoma lepidum
Bluegill (juvenile), F,M
Lepomis macrochirus
Boreal toad S,U
(0.0 12 g, 9.6mm),
Bufo b areas
Coot clam S,U
(embryo/larva),
Mulinia lateralis
Copepod (10-12 d), S,U
Acartia tonsa
Mysid(<24-hrold), F,M
Americamysis bahia
Mysid(<24-hrold), F,M
Americamysis bahia
Amphipod (adult), F,M
Leptocheirus
plumulosus
Amphipod (adult), S,U
Eohaustorius estuarius
Grass shrimp F,M
(48-hr old),
Palaemonetes vulgaris
American lobster R,U
(1st stage),
Homarus americanus
Mud crab F,M
(4th and 5th stages),
LC50
or ECSO
Chemical pH ([ig/L)
85% 8.0-8.1 110
85% 8.0-8.2 190
>90% 7.61 209
85% 7.9-8.0 120
SALTWATER SPECIES
90% 7.8-8.2 37.9
190
>95% 7.3-8.2 43
90% 7.8-8.2 60.6
90% 7.8-8.2 61.6
138
90% 7.8-8.2 59.4
90% 7.8-8.2 71
90% 7.8-8.2 >195
Species
Mean Acute
Valueb
fue/U
110
190
209
120
37.9
190
-
51.05
61.6
138
59.4
71
>195
Reference
Dwyer et al.
1999a
Dwyer et al.
1999a
Brooke 1993a
Dwyer et al.
1999a
Lussier et al.
2000
Kusk and
Wollenberger
1999
Ward and
Boeri 1990a
Lussier et al.
2000
Lussier et al.
2000
Hecht and
Boese 2002
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Dyspanopeus sayii
42
-------
Table 1. Acute Toxicity of Nonylphenol to Aquatic Animals
Species
Winter flounder
(48-hr-old),
Pleuronectes
americanus
Method"
S,M
Sheepshead minnow F,M
(juvenile),
Cyprinodon variegatus
Sheepshead minnow F,M
(juvenile),
Cyprinodon variegatus
Inland silversides F,M
(juvenile),
Menidia beryllina
Chemical
90%
>95%
90%
90%
pH
7.8-8.2
7.4-8.1
7.8-8.2
7.8-8.2
LC50
or ECSO
fue/U
17
310
142
70
Species
Mean Acute
Valueb
fue/U
17
-
209.8
70
Reference
Lussier et al.
2000
Ward and
Boeri 1990b
Lussier et al.
2000
Lussier et al.
2000
a S = static; R = renewal; F = flow-through; M = measured; U = unmeasured.
b Each Species Mean Acute Value was calculated from the underlined number(s) in the preceding column.
43
-------
Table 2. Chronic Toxicity of Nonylphenol to Aquatic Animals
Species
Cladoceran,
Ceriodaphnia dubia
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia magna
Midge,
Chironomus tentans
Rainbow trout,
Oncorhynchus mykiss
Fathead minnow,
Pimephales promelas
Mysid,
Americamysis bahia
Mysid,
Americamysis bahia
Test"
LC
LC
LC
LC
LC
ELS
ELS
Chemical
>95%
93.1
>90%
91.8%
95%
>90%
>95%
_pH
8.3-8.6
8.04
8.46
8.25
7.73
6.97
7.1-8.2
Chronic
Limits
fue/LV
88.7-202
10.7-47.8
116-215
24-39
42-91
6.0-10.3
7.4-14
Chronic
Value
fue/U
133.9
22.62
157.9
30.59
61.82
7.861
10.18
Reference
England 1995
Fliedner 1993
Brooke 1993a
Comber et al.
1993
Kahl et al. 1997
Brooke 1993a
Ward and Boeri
1991c
SALTWATER SPECIES
LC
LC
>95%
_
7.4-8.3
_
3.9-6.7
9.46-15.28
5.112
12.02
Ward and Boeri
1991b
Kuhnetal. 2001
a LC = life-cycle or partial life-cycle; ELS = early life-stage.
b Based upon measured concentrations of nonylphenol.
44
-------
Table 2. Acute-Chronic Ratios
Acute-Chronic Ratios
Species
Cladoceran,
Daphnia magna
Cladoceran,
Daphnia magna
Mysid,
Americamysis bahia
Rainbow trout,
Oncorhynchus mykiss
Acute Value Chronic Value
pH (ug/L) fug/U
7.87-8.46 104 157.9
7.3-8.3
6.72-6.97
190
43
221
30.59
5.112
7.861
Ratio Reference
0.6586 Brooke 1993a
6.211 Comber etal. 1993
8.412
Ward and Boeri
1990a, 1991b
28.11 Brooke 1993a
45
-------
Table 3. Ranked Genus Mean Acute Values with Species Mean Acute-Chronic Ratios
ink3
15
14
13
12
11
10
9
8
7
6
5
4
Genus Mean
Acute Value
(ug/L) Species
FRESHWATER SPECIES
774 Snail,
Physella virgata
596 Dragonfly,
Ophiogomphus sp.
342 Annelid,
Lumbriculus variegatus
289.3 Bonytail chub,
Gila elegans
254.6 Colorado squawfish,
Ptychocheilus lucius
230 Gila topminnow,
Poeciliopsis occidentalis
209 Bluegill,
Lepomis macrochims
184.2 Rainbow trout,
Oncorhynchus mykiss
Apache trout,
Oncorhynchus apache
Lahontan cutthroat trout,
Oncorhynchus clarki henshawi,
and
Greenback cutthroat trout,
Oncorhynchus clarki stomais
174.4 Razorback sucker,
Xyrauchen texanus
160 Midge,
Chironomus tentans
1 44 . 6 Greenthroat darter,
Etheostoma lepidum
Fountain darter,
Etheostoma rubrum
140.6 Cladoceran,
Daphnia magna
Species Mean Species Mean
Acute Value Acute-Chronic
(u2/L)b Ratio0
774
596
342
289.3
254.6
230
209
221 28.11
169.7
166.6
-
174.4
160
190
110
140.6 2.023
46
-------
Table 3. Ranked Genus Mean Acute Values with Species Mean Acute-Chronic Ratios
Rank3
3
2
1
11
10
9
8
7
6
5
4
3
2
1
Genus Mean
Acute Value
fue/U
133.9
120
55.72
209.8
>195
190
138
71
70
61.6
59.4
51.05
37.9
17
Species Mean Species Mean
Acute Value Acute-Chronic
Species (u2/L)b Ratio0
Fathead minnow,
Pimephales promelas
Boreal toad,
Bufo boreas
Amphipod,
Hyalella azteca
SALTWATER SPECIES
Sheepshead minnow,
Cyprinodon variegatus
Mud crab,
Dyspanopeus sayii
Copepod,
Acartia tonsa
Amphipod,
Eohaustorius estuarius
American lobster,
Homarus americanus
Inland silversides,
Menidia beryllina
Amphipod,
Leptocheirus plumulosus
Grass shrimp,
Palaemonetes vulgaris
Mysid,
Americamysis bahia
Coot clam,
Mulinia lateralis
Winter flounder,
133.9
120
55.72
209.8
>195
190
138
71
70
61.6
59.4
51.05 8.412
37.9
17
Pleuronectes americanus
a Ranked from the most resistant to the most sensitive based on Genus Mean Acute Value.
b From Table 1.
c From Table 2.
47
-------
Table 3. Ranked Genus Mean Acute Values with Species Mean Acute-Chronic Ratios
Freshwater
Final Acute Value = 55.49 ug/L
Criterion Maximum Concentration = 55.49 + 2 = 27.75 ug/L
Final Acute-Chronic Ratio = 8.412 (see text)
Final Chronic Value = 55.49 ug/L + 8.412 = 6.5965 ug/L
Saltwater
Final Acute Value =13.93 ug/L
Criterion Maximum Concentration = 13.93 + 2 = 6.965 ug/L
Final Acute-Chronic Ratio = 8.412 (see text)
Final Chronic Value = 13.93 ug/L + 8.412 = 1.6560 ug/L
48
-------
Table 4. Toxicity of Nonylphenol to Aquatic Plants
Species
Chemical pH
Duration Concentration
(days) Effect (ug/L) Reference
Salinity (g/kg).
FRESHWATER SPECIES
Green algae,
Selenastmm
capricornutum
>95%
4 EC50,
number of
cells
410
Ward and Boeri
1990a
SALTWATER SPECIES
Diatom,
Skeletonema
costatum
>95%
30a
EC50,
number of
cells
27
Ward and Boeri
1990d
49
-------
Table 5. Bioaccumulation of Nonylphenol by Aquatic Organisms
Species
Water
Cone.
Chemical (ug/L)a
Duration
BCF Normalized
Percent or BCF or
pH
(days) Tissue Lipids BAFb BAFC Reference
FRESHWATER SPECIES
Fathead >95% 4.9 7.0-7.6 27
minnow
(0.5-1 g),
Pimephales
promelas
Fathead >95% 22.7 7.0-7.6 27
minnow
(0.5-1 g),
Pimephales
promelas
Fathead 99% 18.4 7.62 4
minnow
(4-wk old),
Pimephales
promelas
Fathead 99% 41.9 7.62 4
minnow
(4-wk old),
Pimephales
promelas
Fathead 99% 82.1 7.62 4
minnow
(4-wk old),
Pimephales
promelas
Fathead 99% 9.3 7.60 28
minnow
(4-wk old),
Pimephales
promelas
Fathead 99% 19.2 7.60 28
minnow
(4-wk old),
Pimephales
promelas
Whole - 271
body
Whole - 344
body
Whole 4.7±1.7 751
body
Whole 4.7±1.7 677
body
Whole 4.7±1.7 945
body
Whole 4.7±1.7 769
body
Whole 4.7±1.7 984
body
Ward and
Boeri 1991a
Ward and
Boeri 1991a
159.8 Brooke
1993b
144.0 Brooke
1993b
201.1 Brooke
1993b
163.6 Brooke
1993b
209.4 Brooke
1993b
50
-------
Table 5. Bioaccumulation of Nonylphenol by Aquatic Organisms
Water
Cone.
BCF Normalized
Duration Percent or BCF or
Species Chemical (ug/L~)a pH
Fathead 99% 38.1 7.60
minnow
(4-wk old),
Pimephales
promelas
Fathead 99% 77.5 7.60
minnow
(4-wk old),
Pimephales
promelas
Fathead 0.4
minnow >98% 1.6
(adult), 3.4
Pimephales
promelas
Bluegill 99% 21.6 7.79
(4-wk old),
Lepomis
macrochirus
Bluegill 99% 43.9 7.79
(4-wk old),
Lepomis
macrochirus
Bluegill
(4-wk old), 99% 86.5 7.79
Lepomis
macrochirus
Bluegill 99% 5.6 7.55
(4-wk old),
Lepomis
macrochirus
Bluegill 99% 12.4 7.55
(4-wk old),
Lepomis
macrochirus
Bluegill 99% 27.6 7.55
(4-wk old),
Lepomis
macrochirus
(days) Tissue Lipids BAFb
28 Whole 4.7±1.7 876
body
28 Whole 4.7±1.7 603
body
Whole 203
42 body - 252
268
4 Whole 4.9±1.5 279
body
4 Whole 4.9±1.5 257
body
4 Whole 4.9±1.5 223
body
28 Whole 4.9±1.5 231
body
28 Whole 4.9±1.5 253
body
28 Whole 4.9±1.5 250
body
BAFC Reference
186.4 Brooke
1993b
128.3 Brooke
1993b
.
Giesy et al.
2000
56.94 Brooke
1993b
52.45 Brooke
1993b
45.51 Brooke
1993b
47.14 Brooke
1993b
51.63 Brooke
1993b
51.02 Brooke
1993b
51
-------
Table 5. Bioaccumulation of Nonylphenol by Aquatic Organisms
Water
BCF Normalized
Cone.
Species Chemical (ug/L~)a
Bluegill 99% 59.5
(4-wk old),
Lepomis
macrochirus
Bluegill
(juvenile),
Lepomis
macrochirus
Blue mussel,
Mytilus
edulis
Blue mussel,
Mytilus
edulis
Common
shrimp,
Crangon
crangond
Common
shrimp,
Crangon
crangond
Three-spined
stickleback,
Gasterosteus
aculeatus
Three-spined
stickleback,
Gasterosteus
aculeatus
1.0
96.4% 3.0
30.0
14C- 5.9
labeled
14C- 6.2
labeled
14C- 6.4
labeled
14C- 7.4
labeled
14C- 4.8
labeled
14C- 4.9
labeled
Duration Percent or
pH (days) Tissue Lipids BAFb
7.55 28 Whole 4.9±1.5 191
body
76
7.7 20 Whole 0.72± 60
body 0.46 37
SALTWATER SPECIES
16 Whole 1.6 2,740
body
16 Whole 1.9 4,120
body
16 Whole 1.4 110
body
16 Whole 1.7 900
body
16 Whole 6.7 1,200
body
16 Whole 7.8 1,300
body
BCF or
BAFC Reference
38.98 Brooke
1993b
105.6
83.33 Liber etal.
51.39 1999
1,712 Ekelundet
al. 1990
2,168 Ekelundet
al. 1990
78.75 Ekelund et
al. 1990
529.4 Ekelund et
al. 1990
179.1 Ekelundet
al. 1990
166.7 Ekelund et
al. 1990
52
-------
Table 5. Bioaccumulation of Nonylphenol by Aquatic Organisms
Water BCF Normalized
Cone. Duration Percent or BCF or
Species Chemical (ug/L)a nH (days) Tissue Lipids BAFb BAFC Reference
""Measured concentration of nonylphenol.
bBioconcentration factors (BCFs) and bioaccumulation factors (BAFs) are based on measured concentrations of
nonylphenol in water and in tissue.
°When possible, the factors were normalized to 1% lipid by dividing the BCFs and BAFs by the percent lipid measured
in the test organism.
dNon-resident species.
53
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Species
Chemical
pH
Concentration
(ug/L) Reference
FRESHWATER
Phytoplankton >98% 8.8 - 10.6
and Periphyton
Green alga,
Chlamydomonas
reinhardtii
Floating moss,
Salvinia molesta
Duckweed, - 5.6
Lemna minor
Duckweed, Lemna
minor
Ciliate protozoan,
Tetrahymena
pyriformis
Ciliate protozoan, - 7.40
Tetrahymena
pyriformis
Rotifer Technical 7.5
(4 to 6 hr-old
female)
Brachionus
calyciflorus
Clam (15 g),
Anodonta
cataractae
Zooplankton 96.4% 7.5 - 8.2
Zooplankton >98% 8.8 - 10.4
Benthic macro- 96.4% 7.5 - 8.2
invertebrates
Cladoceran - 8.0
(<24-hr old),
Daphnia magna
6wk
24 days
9 days
96 hr
4 days
24hr
40hr
96 hr
144 hr
(fed)
20 days
6wk
20 days
21 days
SPECIES
Dominant
species changed
100% algistatic
Reduced frond
production
IC50
Reduced frond
production
EC50
Reduced
population
growth 50%
Sexual
reproduction
reduced
LC50
NOEC
LOEC
NOEC
NOEC
LOEC
NOEC
LOEC (reduced
fecundity)
29 - 120
6,250
2,500
5,500
125
460
747
50
5,000
23
76
19-44
23
76
50
100
Hense et al.
2003
Weinberger and
Greenhalgh
1984
Prasad 1986
Weinberger and
lyengar 1983
Prasad 1986
Yoshioka 1985
Schultz 1997
Preston et al.
2000
McLeese et al.
1980b
O'Halloran et
al. 1999
Severin et al.
2003
Schmude et al.
1999
Baldwin et al.
1997
54
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Concentration
Species
Cladoceran
(<24-hr old and
adults),
Daphnia magna
Cladoceran
(<24-hr old),
Daphnia magna
Cladoceran
(<24-hr old),
Daphnia magna
Cladoceran
(<24-hr old),
Daphnia magna
Cladoceran
(<24-hr old),
Daphnia magna
Cladoceran
(<36-hr old),
Daphnia galeata
mendotae
Cladoceran
(>48-hr old),
Daphnia pulex
Cladoceran
(>48-hr old),
Daphnia pulex
Cladoceran
(>48-hr old),
Daphnia pulex
Cladoceran
(<24-hr old),
Ceriodaphnia
dubia
Chemical pH
-85% 7.8 - 8.4
-85% 7.7+0.02
Technical
-85%
-85%
Practical
grade
Practical
grade
Practical
grade
>95%
8.3-8.6
Duration
96 hr
(fed)
21 days
21 days
48 hr
35 day
30 day
48 hr
48hr
48 hr
48 hr
Effect
MATC (young)
MATC (adults)
No sex ratio
change (high
food rate)
Increased ratio
of males (low
food rate)
50% adult
mortality
NOEC
(deformed
offspring)
EC50
LOEC
NOEC
LOEC
(deformed
offspring)
LC50
LC50
LC50
LC50
(fed)
fue/U
302
136
25
25
200.5
44
234
272
337
>50
10
50
140
176
190
276
Reference
Gerritsen et al.
1998
Baer and
Owens 1999
LeBlanc et al.
2000
Zang et al. 2003
Zang et al. 2003
Shurin and
Dodson 1997
Ernst etal. 1980
Ernst etal. 1980
Ernst etal. 1980
England 1995
55
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Concentration
Species Chemical pH
Cladoceran >95% 8.3-8.6
(<24-hr old),
Ceriodaphnia
dubia
Midge >95% 8.2
(2nd instar),
Chironomus
tentans
Sea lamprey - 7.5-8.2
(larva),
Petromyzon
marinus
Brook trout
(juvenile),
Salvelinus
fontinalis
Lake trout
(juvenile),
Salvelinus
naymaycush
Brown trout - 7.0
(fingerling),
Salmo trutta
Atlantic salmon
(4g),
Salmo salar
Atlantic salmon
(48.3+2.6mmTL),
Salmo salar
Chinook salmon - 7.2
(juvenile),
Oncorhynchus
tshawytscha
Rainbow trout - 7.5-8.2
(juvenile),
Oncorhynchus
mykiss
Duration Effect
7 days LC50 (fed)
14 days LC50
EC50
14 hr LT100
96 hr LC50
35 days LC50
(fed)
2 hr LT100
96 hr LC50
(fed)
30 days No change in
plasma
vitellogenin or
gillNaKATPase
activity or plasma
Cl'andNa+
3 hr LT100
4 hr LT100
(ug/L) Reference
225 England 1995
119 England and
95 Bussard 1993
5,000 Applegate et al.
1957
145 Holmes and
Kingsbury 1980
>40 Holmes and
Kingsbury 1980
5,000 Wood 1953
900 McLeese et al.
1980b
20 Moore et al.
2003
10,000 MacPhee and
Ruelle 1969
5,000 Applegate et al.
1957
56
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Rainbow trout
(juvenile),
Oncorhynchus
mykiss
Rainbow trout
(juvenile),
Oncorhynchus
mykiss
Rainbow trout
(juvenile),
Oncorhynchus
mykiss
Rainbow trout
(adult males),
Oncorhynchus
mykiss
Rainbow trout
(adult males),
Oncorhynchus
mykiss
Rainbow trout
(50-200 g),
Oncorhynchus
mykiss
Rainbow trout
(50 - 200 g),
Oncorhynchus
mykiss
Rainbow trout,
(40 - 60 g),
Oncorhynchus
mykiss
Rainbow trout,
(40 - 60 g),
Oncorhynchus
mykiss
Rainbow trout,
(40 - 60 g),
Oncorhynchus
mykiss
Chemical
Practical
grade
Practical
grade
>99%
>99%
>99%
Duration
96hr
96 hr
96 hr
3wk
3wk
72hr
72hr
8hr
2-5hr
2-5hr
Effect
LC50
LC50
LC50
Increased
vitellogenin
production
Increased
vitellogenin
production
LC50
Increased
vitellogenin
mRNA
Tissue half-life
fat 19.8 hr
muscle 18.6 hr
liver 5.9 hr
Eviscerated
carcass
BAF = 24.21
Viscera
BAF = 98.2
Concentration
(ug/L) Reference
920 Ernst etal. 1980
560 Ernst etal. 1980
230 Holmes and
Kingsbury 1980
20.3 Joblingetal.
1996
54.3
193.65
14.14
18
18
18
Jobling et al.
1996
Lech etal. 1996
Lech etal. 1996
Lewis and Lech
1996
Lewis and Lech
1996
Lewis and Lech
1996
57
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Species
Rainbow trout
(juvenile),
Oncorhynchus
mykiss
Rainbow trout
(juvenile),
Oncorhynchus
mykiss
Rainbow trout
.(juvenile),
Oncorhynchus
mykiss
Rainbow trout
(juvenile),
Oncorhynchus
mykiss
Rainbow trout
(35-50 g,
immature),
Oncorhynchus
mykiss
Rainbow trout
(adult males),
Oncorhynchus
mykiss
Rainbow trout
(103-168 g,
juvenile)
Oncorhynchus
mykiss
Rainbow trout
(adult males),
Oncorhynchus
mykiss
Chemical pH
6.5
8.0-8.4
99%
Technical
Duration
4hr
72 hr
22 days
35 days
96 hr
21 days
3wk
9 days
Effect
Vitellogenin
mRNA
production
Vitellogenin
mRNA
production
Reduced growth
at 108 days
Reduced growth
at 466 days
Decreased
number of
muscarinic
cholinergic
receptors in
brain
Increased
vitellogenin in
blood plasma
BCF=116BCF
No vitellogenin
induction
10 days per month Epidermal
for 4 months mucous cell
granulation
Concentration
(ug/L) Reference
10 Renetal. 1996a
100 Renetal. 1996a
50
30
220
63
81
109
Ashfield et al.
1998
Jones et al.
1998
50 Tremblay and
Van Der Kraak
1998
Blackburn et al.
1999
Pedersen et al.
1999
Burkhardt-
Holm et al.
2000
58
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Rainbow trout
(598 g; juvenile
females),
Oncorhynchus
mykiss
Rainbow trout
(1667+201.6 g;
F0 3 yr-old
adults),
Oncorhynchus
mykiss
Rainbow trout
(6-mo-old),
Oncorhynchus
mykiss
Lahontan
cutthroat trout
(juvenile),
Oncorhynchus
clarki henshawi
Apache trout
(juvenile),
Oncorhynchus
mykiss
Northern
squawfish
(juvenile),
Ptychocheilus
oregonensis
Chemical
99%
Duration
18 wk
98%
7.6
4 months
(exposed 10
days/month)
7.2
4wk
96 hr
96 hr
7.2
3hr
Effect
Reduced GSI;
Reduced HSI;
Induced
vitellogenin;
Lowered
plasma
estradiol;
Lowered
plasma FSH
Reduced
embryo
survival;
Reduced hatch;
F0 Males
increased
vitellogenin;
Fj Females
increased
vitellogenin and
testosterone;
Fj Males
increased
estradiol
Liver tissue
showed
hemorrhage and
lymphocyte
infiltration
Decreased
number of
muscarinic
cholinergic
receptors in
brain
Decreased
number of
muscarinic
cholinergic
receptors in
brain
LT100
Concentration
(ug/L) Reference
85.6
85.6
8.3
85.6
Harris et al.
2001
8.3
1 Schwaiger et al.
2002
10
1
10
10
220 Uguz et al.
2003
220
>130
10,000
Jones et al.
1998
Jones et al.
1998
MacPhee and
Ruelle 1969
59
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Concentration
Species Chemical pH
Colorado
squawfish
(juvenile),
Ptychocheilus
lucius
Goldfish - 7.0
(juvenile),
Carassius auratus
Common carp Technical 7.6
(15.2 + 3.8 g (90% 4-
juvenile), NP)
Cyprinus carpio
Common carp 95% 7.57
(50-1 50 g mature ±°-03
males),
Cyprinus carpio
Fathead minnow 99% 7.62
(4-wk old),
Pimephales
promelas
Fathead minnow 99% 7.60
(4-wk old),
Pimephales
promelas
Fathead minnow,
Pimephales
promelas
Fathead minnow >98%
(mature),
Pimephales
promelas
Fathead minnow >98%
(mature),
Pimephales
promelas
Duration
96 hr
5hr
70 days
28-31 days
11 °C
4 days
28 days
96 hr
42 days
42 days
Effect
Decreased
number of
muscarinic
cholinergic
receptors in
brain
LT100
Decreased
erythrocytes;
Increased
reticulocytes
BCF = 546.5
No change in
17-estradiol,
testosterone, or
vitellogenin
LC50
(fed)
BCF = 100.4
Decreased
number of
muscarinic
cholinergic
receptors in
brain
Possible
increased
number of
Sertoli cells in
males
Decreased
fecundity
fue/U
>220
5,000
10
10
5.36
138
193
>220
1.6
>3.4
Reference
Jones et al.
1998
Wood 1953
Schwaiger et al.
2000
Villenueve et
al. 2002
Brooke 1993b
Brooke 1993b
Jones et al.
1998
Miles-
Richardson et
al. 1999
Giesy et al.
2000
60
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Concentration
Species Chemical pH Duration
Fathead minnow >98% - 42 days
(mature),
Pimephales
promelas
Fathead minnow >98% - 42 days
(mature),
Pimephales
promelas
Bluegill - 7.0 2hr
(juvenile),
Lepomis
macrochims
Bluegill - 7.5-8.2 14 hr
(juvenile),
Lepomis
macrochiru
Bluegill 99% 7.79 4 days
(4-wk old),
Lepomis
macrochims
Bluegill 99% 7.55 28 days
(4-wk old),
Lepomis
macrochims
Bluegill 96.4% 7.7-7.9 20 days
(juvenile),
Lepomis
macrochims
Southern platyfish Technical - 28 days
(adult, 0.62 to 85%
1.15 g),
Xiphophorus
maculatus
Green Swordtail Technical - 96 hr
(adult males), 72 hr
Xiphophorus
helleri
Effect
Increased d1
vitellogenin
Increased
c?& 9
17p-estradiol
LT100
LT100
LC50
(fed)
BCF=35.31
NOEC
LOEC
(survival)
Reduced GSI
LC50
Vitellogenin
induced
fue/U
>3.4
>0.05
(not all test
concentrations)
5,000
5,000
135
126
76
243
960
206
4
Reference
Giesy et al.
2000
Giesy et al.
2000
Wood 1953
Applegate et al.
1957
Brooke 1993b
Brooke 1993b
Liber et al.
1999
Kinnberg et al.
2000
Kwak et al.
2001
61
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Concentration
Species Chemical
Green Swordtail Technical
(juvenile 30-d-old
males),
Xiphophorus
helleri
African clawed ACS
frog (larva), Grade
Xenopus laevis
African clawed
frog (larva),
Xenopus laevis
pH Duration
60 days
7.8 - 8.0 21 days
12 wk
Effect
Reduced sword
length
NOEC
LOEC
(increased rate
of tail
resorption)
Increased
female
phenotypes
fue/U
0.2
25
50
22
Reference
Kwak et al.
2001
Fort and Stover
1997
Kloas et al.
1999
SALTWATER SPECIES
Red alga, >95%
Champia parvula
Barnacle
(cypris larva),
Balanus
amphitrite
Soft-shell clam,
Mya arenaria
Coot clam, 90%
Mulinia lateralis
Coot clam, 90%
Mulinia lateralis
Coot clam, 90%
Mulinia lateralis
Blue mussel, 90%
Mytilus edulis
Blue mussel,
Mytilus edulis
Blue mussel,
Mytilus edulis
Blue mussel,
Mytilus edulis
2 days
48 h
360 hr
30-3 la 24 hr
30-3 la 48 hr
30-3 la 72 hr
30-3 la 96 hr
32a 360 hr
32a 13 days
32a 30 days
62
No effect on
sexual
reproduction
Reduced cyprid
settlement
No mortality
LC50
LC50
LC50
LC50
LC50
Reduced byssus
strength
Reduced byssus
strength
167
1.0
700
50
50
40
3,000
500
56
56
Tagliabue 1993
Billinghurst et
al. 1998
McLeese et al.
1980b
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Granmo et al.
1989
Granmo et al.
1989
Granmo et al.
1989
Granmo et al.
1989
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Species
Blue mussel,
Mytilus edulis
Blue mussel,
Mytilus edulis
Blue mussel,
Mytilus edulis
Blue mussel,
Mytilus edulis
Blue mussel
(40-50 mm
length), Mytilus
edulis
Blue mussel,
Mytilus edulis
galloprovincialis
Mysid,
Americamysis
bahia
Mysid,
Americamysis
bahia
Mysid,
Americamysis
bahia
Mysid,
Americamysis
bahia
Mysid,
Americamysis
bahia
Mysid,
Americamysis
bahia
Mysid,
Americamysis
bahia
Chemical
pH
90%
90%
90%
90%
90%
90%
>95%
32a
32a
32a
30-3 la
30-3 la
30-3 la
30-3 la
30-3 la
30-3 la
20a
Duration
30 days
32 days
24 hr
72 hr
50 days
2 days
24hr
48hr
72hr
120hr
144hr
168 hr
24 hr
Effect
No byssus
threads formed
Reduction in
growth
No effect on
fertilization
No effect on
development
BCF = 350
Repelled
attachment
LC50
LC50
LC50
LC50
LC50
LC50
LC50
Concentration
fue/U
100
56
200
200
40
22
-114
-82
-66
-60
-60
-60
>47
Reference
Granmo et al.
1989
Granmo et al.
1989
Granmo et al.
1989
Granmo et al.
1989
Granmo et al.
1991a,b
Etohetal. 1997
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Ward and Boeri
1990a
63
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Concentration
Species Chemical pH Duration
Mysid, >95% 20a 48 hr
Americamysis
bahia
Mysid, >95% 20a 72 hr
A wipvi (~'nwi\)(?i f
j*lAA£t,A /CWA/f l/i3/i3
Pacific Oyster - 35a 72 hr
(embryo-larva),
Crassostrea gigas
Copepod - 18a 48 hr
(10-12 d),
Acartia tonsa
Amphipod, 90% 30-3 la 48 hr
Leptocheirus
plumulosus
Amphipod, 90% 30-3 la 72 hr
Leptocheirus
plumulosus
Amphipod, 90% 30-3 la 120 hr
Leptocheirus
plumulosus
Amphipod, 90% 30-3 la 144 hr
Leptocheirus
plumulosus
Amphipod, 90% 30-3 la 168 hr
Leptocheirus
plumulosus
Grass shrimp, 90% 30-3 la 24 hr
Palaemonetes
vulgaris
Grass shrimp, 90% 30-3 la 48 hr
Palaemonetes
vulgaris
Grass shrimp, 90% 30-3 la 72 hr
Palaemonetes
vulgaris
Grass shrimp, 90% 30-3 la 120 hr
Palaemonetes
vulgaris
Effect
LC50
LC50
Delayed
development to
D-shape stage
LC50 synthetic
media
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
(ug/L)
>47
44
100
360
280
-160
-80
-50
-40
-30
-125
-60
-60
-60
Reference
Ward and Boeri
1990a
Ward and Boeri
1990a
Nice et al. 2000
Kusk and
Wollenberger
1999
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
64
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Concentration
Species Chemical
Shrimp, >95%
Crangon
septemspinosa
Shrimp, >95%
Crangon
septemspinosa
Shrimp, >95%
Crangon
septemspinosa
American lobster, 90%
Homarus
americanus
American lobster, 90%
Homarus
americanus
American lobster, 90%
Homarus
americanus
American lobster, >95%
Homarus
americanus
Atlantic salmon,
Salmo salar
Atlantic salmon,
Salmo salar
Atlantic salmon,
Salmo salar
Atlantic salmon,
Salmo salar
Sheepshead 90%
minnow,
Cyprinodon
variegatu
Sheepshead 90%
minnow,
Cyprinodon
variegatu
pH Duration
96 hr
96 hr
96 hr
30-3 la 24 hr
30-3 la 48 hr
30-3 la 72hr
96hr
96hr
96hr
96 hr
96 hr
30-3 la 72 hr
30-3 la 120 hr
Effect
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
LC50
fue/U
300
300
300
-140
-140
-140
170
190
160
130
900
-150
-125
Reference
McLeese et al.
1980b
McLeese et al.
1980b
McLeese et al.
1980b
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
McLeese et al.
1980b
McLeese et al.
1980b
McLeese et al.
1980b
McLeese et al.
1980b
McLeese et al.
1980b
Lussier et al.
2000
Lussier et al.
2000
65
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Concentration
Species
Sheepshead
minnow,
Cyprinodon
variegatu
Sheepshead
minnow,
Cyprinodon
variegatu
Sheepshead
minnow,
Cyprinodon
variegatu
Sheepshead
minnow,
Cyprinodon
variegatu
Sheepshead
minnow,
Cyprinodon
variegatu
Killifish
(embryo),
Fundulus
heteroclitus
Killifish
(embryo),
Fundulus
heteroclitus
Killifish
(1-day old larva),
Fundulus
heteroclitus
Killifish
(14-day old larva),
Fundulus
heteroclitus
Killifish
(28-day old larva),
Fundulus
heteroclitus
Chemical pH Duration
90% 30-3 la 144 hr
90% 30-3 la 168 hr
>95% 15-17a 24 hr
>95% 15-17a 48 hr
>95% 15-17a 72 hr
85 - 90% 20a 10 days
(technical)
85 - 90% 20a 96 hr
(technical)
85 - 90% 20a 96 hr
(technical)
85 - 90% 20a 96 hr
(technical)
85 - 90% 20a 96 hr
(technical)
Effect
LC50
LC50
LC50
LC50
LC50
100% abnormal
development
LC50
LC50
(fed)
LC50
(fed)
LC50
(fed)
(ug/L) Reference
~120 Lussier et al.
2000
~120 Lussier et al.
2000
>420 Ward and Boeri
1990c
340 Ward and Boeri
1990c
320 Ward and Boeri
1990c
2,204 Kelly and Di
Giulio 2000
5,444 Kelly and Di
Giulio 2000
214 Kelly and Di
Giulio 2000
209 Kelly and Di
Giulio 2000
260 Kelly and Di
Giulio 2000
66
-------
Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms
Species
Threespine
stickleback
Gasterosteus
aculeatus
Inland silversides,
Menidia beryllina
Inland silversides,
Menidia beryllina
Inland silversides,
Menidia beryllina
Inland silversides,
Menidia beryllina
Inland silversides,
Menidia beryllina
Inland silversides,
Menidia beryllina
"Salinity (g/kg).
Chemical pH
Commercial 32a
(para-
substituted
with branched
nonyl chain)
90%
Duration
96 hr
Effect
LC50
90%
90%
90%
90%
90%
30-3 la
30-3 la
30-3 la
30-3 la
30-3 la
30-3 la
24hr
48hr
72 hr
120 hr
144 ru-
les hr
LC50
LC50
LC50
LC50
LC50
LC50
Concentration
(ug/L) Reference
370 Granmo et al.
1991a
-120
-100
-80
-60
-60
-60
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
Lussier et al.
2000
67
-------
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