United States Environmental Office of Water EPA-822-R-05-005 

Protection Agency 4304T December 2005 

Aquatic Life Ambient 

Water Quality Criteria B

Nonylphenol 

FINAL 
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 
ii 

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
iii 

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 
iv 

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 
v 

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. BIOACCUMULATION ......................................................................................................... 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 
vi 

9. NATIONAL CRITERIA........................................................................................................ 34 

9.1. Freshwater ................................................................................................................... 34 

9.2. Saltwater ...................................................................................................................... 34 

10. IMPLEMENTATION ......................................................................................................... 35 

11. REFERENCES..................................................................................................................... 68 
vii 

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 
1 

1. INTRODUCTION 

1.1. Physical-Chemical Properties 

Nonylphenol (C15H24O) 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 para-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 ortho-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 
2 

specific gravity of 0.953 g/mL at 20o

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 μg/L at pH 5.0, 6,237 μg/L at pH 

7.0, 11,897 μg/L at pH 9.0. Nonylphenol is soluble in seawater at 3,630 μg/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 2EC (4,600 μg/L) and 25EC (6,350 μg/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 μg/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 μg/L. They found nonylphenol in all sediment samples with concentrations ranging 

from 0.17 to 72 μg/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 μg/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 μg/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 μg/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. 
3 

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 22EC was mineralized to CO2 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 22EC. 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 (APnEO) 
4 

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 μg/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 APnEOs 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 CO2 and water. Nonylphenol is formed 

during anaerobic breakdown of the APnEOs 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 μg/kg more than twenty years after the treatment plant ceased operation. 

Bennett and Metcalfe (1998; 2000) found that nonylphenol was widely distributed in lower 
5 

Great Lakes sediments and reached 37,000 μg/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 halflife of 60 days in marine sediments, similar to that of nonylphenol (Shang et al. 1999). Ferguson 

and Brownawell (2003) conducted degradation studies with APnEOs 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 salar) 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 
6 

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 μg/kg 

and a mean value of 12.0 μg/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 μg/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 μg/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 
7 

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 Disruptor 

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 
8 

appropriate screening and or testing protocols being considered under the Agency’s Endocrine 

Disruptor 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). 
9 

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 μg/L for an amphipod (Hyalella azteca) to 774 μg/L for a snail (Physella virgata). 

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 CaCO3, 

respectively. An LC50 of 20.7 μg/L was calculated in the lower hardness water and 150 μg/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 μg/L from a test that had the solutions renewed daily and Comber et al. 

(1993) reported an EC50 of 190 μg/L in a static test. The Daphnia magna SMAV is 140.6 μg/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 μg/L, nymphs of the 

dragonfly Ophiogomphus sp. had an LC50 of 596 μg/L and the least sensitive species tested was 

a snail, Physella virgata, which had an LC50 of 774 μ/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 μg/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 μg/L and were ranked second in sensitivity to nonylphenol 

(Dwyer et al. 1999a). 

Freshwater fish species were in the mid-range of sensitivity to nonylphenol (Figure 1). 

SMAVs ranged from 110 μg/L for the fountain darter (Etheostoma rubrum) to 289.3 μg/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 
10 

tested and had similar LC50s ranging from 140 to 270 μg/L (Dwyer et al. 1995; Brooke 1993a). 

Dwyer et al. (1995, 1999a) exposed nine species of fish 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 LC50s ranged from 110 μg/L for the fountain darter, Etheostoma rubrum, to a geometric 

mean of 289.3 μg/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 

(Pimephales promelas). LC50s for this species ranged from 128 μg/L (Brooke 1993a) to 360 

μg/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 μg/L). One test was available 

for the bluegill (Lepomis macrochirus) and the LC50 was 209 μg/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 μg/L (calculated using the procedure described in 

the “Guidelines”). This FAV is essentially the same as the lowest freshwater SMAV of 55.72 

μg/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 μg/L for the winter flounder, 

Pleuronectes americanus, to 209.8 μg/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 
11 

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 μg/L; Lussier et al. 2000) compared to when they were not fed (43 

μg/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 μg/L; Lussier et al. 2000) was 

lower than when the organisms were not fed (310 μg/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 μg/L; Lussier et al. 2000), the copepod, 

Acartia tonsa (LC50 = 190 μg/L; Kusk and Wollenberger 1999), American lobster, Homarus 

americanus (LC50 = 71 μg/L; Lussier et al. 2000), mud crab, Dyspanopeus sayii (LC50 >195 

μg/L; Lussier et al. 2000) and the amphipods, Leptocheirus plumulosus (LC50 = 61.6; Lussier et 

al. 2000) and Eohaustorius estuarius (LC50 = 138 μg/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 μg/L as the average of three tests (299, 194, 189 μg/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 μg/L. The sensitivity of the saltwater fish inland silversides (Menidia 
12 

beryllina), was intermediate (LC50 = 70 μg/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 americanus, 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 μg/L (calculated using the procedure described in the 

“Guidelines”). This FAV is lower than the lowest SMAV of 17 μg/L for the the winter flounder, 

Pleuronectes americanus. 
13 

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 μg/L, but not at 

88.7 μg/L, and survival was reduced at 377 μg/L, but not at 202 μg/L. Based upon reproductive 

impairment, the Chronic Value for C. dubia was 133.9 μ/L. At the end of 48 hr in the same test, 

effects were observed and an EC50 of 69 μg/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 20EC 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 μg/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 μg/L and the Lowest Observed Effect Concentration 

(LOEC) of 47.8 μ/L reported, the chronic value (geometric mean of the NOEC and LOEC) for 

D. magna in this test is 22.62 μ/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 D. magna

growth and reproduction were significantly affected at 215 μg/L, but not at 116 μg/L. Survival 

was reduced to 60 percent at 215 μg/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 
14 

mean of the lower (116 μg/L) and upper (215 μg/L) chronic limits, for this test was 157.9 μg/L. 

Dividing the acute value (104 μg/L), determined from a companion test for this species (Brooke 

1993a; Table 1) by the chronic value (157.9 μg/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 μg/L. The number of live young produced was significantly reduced at concentrations > 39 

μg/L when compared to controls. Growth was reduced at concentrations > 71 μg/L and survival 

of adults was reduced at concentrations > 130 μ/L. Based on reproductive impairment, the 

chronic value, calculated as the geometric mean of the lower (24 μg/L) and upper (39 μg/L) 

chronic limits, for this test was 30.59 μg/L. Dividing the acute value (190 μg/L), determined in 

in a companion test for this species (Table 1) by the chronic value (30.59 μg/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 μg/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 

μg/L, respectively. The chronic value, calculated as the geometric mean of the NOEC and the 

LOEC, is 61.82 μg/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 μg/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 μg/L). At the end of 
15 

the test, survival was significantly reduced at concentrations > 23.1 μg/L but not at 10.3 μg/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 μg/L, but not at 6.0 μg/L. Based on 

growth, the NOEC and LOEC determined in this study were 6.0 and > 10.3 μg/L, respectively. 

The chronic value, calculated as the geometric mean of the NOEC and the LOEC, is 7.861 μg/L 

for rainbow trout. Dividing the acute value (221 μg/L), determined in in a companion test for 

this species (Table 1) by the chronic value (7.861 μg/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, Pimephales promelas

(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 μg/L. 

Embryos in the control group and those in the three lowest nonylphenol exposure concentrations 

(2.8, 4.5, and 7.4 μg/L) began to hatch on the third day of exposure, while the two higher 

concentrations (14 and 23 μg/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 μg/L. Fish survival averaged 56.7 % at 23 μg/L 

nonylphenol, 66.7 % at 14 μg/L nonylphenol, and 76.7 % at 7.4 μg/L nonylphenol. At 

concentrations < 7.4 μg/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 μg/L and 14 μg/L, respectively. The chronic value, calculated as the 

geometric mean of the NOEC and the LOEC, is 10.18 μg/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 
16 

and Boeri 1991b). There was no effect on survival or reproduction at 6.7 μg/L, but there was a 

18 % reduction in survival and a 53% reduction in reproduction at 9.1 μg/L. Effects on survival 

at the highest concentration tested (21 μg/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 μg/L 

nonylphenol, but no difference in growth for mysids exposed to 3.9 μg/L nonylphenol. Based on 

growth, the NOEC and LOEC determined in this study were 3.9 μg/L and 6.7 μg/L, respectively. 

 The chronic value, calculated as the geometric mean of the NOEC and the LOEC, is 5.112 μg/L 

for mysids (Table 2). Dividing the acute value (43 μg/L), determined in in a companion test for 

this species (Table 1) by the chronic value (5.112 μg/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 μg/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 μg/L, but was 

not affected at 9.46 μg/L. Based on reproduction, the NOEC and LOEC determined in this study 

were 9.46 μg/L and 15.28 μg/L, respectively. The chronic value, calculated as the geometric 

mean of the NOEC and the LOEC, is 12.02 μg/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 
17 

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 μg/L) is closest to both the freshwater FAV (55.49 

μg/L) and the saltwater FAV (13.93 μg/L). Therefore, the SMACR for the mysid is used as the 

FACR and is 8.412. 
18 

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 μg/L based on cell counts. At the 

end of the toxicity test, algae from the highest exposure concentration (720 μg/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 μg/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 μg/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 μg/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. 
19 

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 μg/L nonylphenol, respectively. Brooke (1993b) exposed 

juvenile fathead minnow (Pimephales promelas) 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 μg/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 for M. edulis. The BCF for M. edulis was an 

estimate, because steady-state tissue concentration was not reached during 16 days of exposure. 
20 

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 μg/L) and Chlamydomonas reinhardtii (6,250 μg/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 μg/L (Weinberger and Iyengar 1983; Prasad 1986). 

Protozoa were affected in the concentration range from 50 to 747 μg/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 μg/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 

μg/L for young and adult animals, respectively (Gerritsen et al. 1998). In a 48-hr test with the 

same species, EC50s 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 μg/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, Daphnia pulex, was exposed for 48 hr in tests in which 

nonylphenol concentrations decreased more than 50 percent during the exposures (Ernst et al. 
21 

1980). The LC50s determined in the test ranged from 140 to 190 μg/L, which agreed with 

LC50s for other cladoceran species. The cladoceran, Ceriodaphnia dubia, gave similar LC50 

results of 276 and 225 μg/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 μg/L (Table 2). England and Bussard (1993) reported an 

EC50 and an LC50 for larva of the midge, Chironomus tentans, of 95 and 119 μg/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, Pimephales promelas, of 138 μg/L and a 96-hr LC50 for the 

bluegill, Lepomis macrochirus, of 135 μg/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 μg/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 μg/L for brook trout juveniles (Salvelinus fontinalis), a 96-hr LC50 of 230 μg/L for 

rainbow trout juveniles (Oncorhynchus mykiss) and a 32-day LC50 of > 40 μg/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 LC50s ranging from 560-920 μg/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 μg/L of nonylphenol (Ashfield et al. 

1998). They found reduced growth in fish exposed to 1 μg/L for 22 days and grown for 86 days 

beyond treatment. Growth was not reduced in the 10 μg/L treatment but was in the 50 μg/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 μg/L treatments on 
22 

day 55 of the study, but not at the 1 μg/L. At day 466, the fish exposed to 10 μg/L recovered the 

growth reductions seen earlier and only the 30 μg/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 μg/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 suborganismal 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 μg/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 μg/L 
23 

nonylphenol but were negatively affected at 76 μg/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 μg/L. The fish (bluegill) in the mesocosms (Liber et al. 1999) were unaffected 

at nonylphenol exposures < 76 μg/L, but survival was reduced at 243 μg/L. In one exposure 

replicate with a mean nonlyphenol concentration of 93 μg/L, survival of the fish was reduced 

after 20 days of exposure indicating that concentrations near 100 μg/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 μg/L. The zooplankton in the study were not affected at concentrations ranging from 

19 to 44 μg/L (mean of 30 μg/L), but species richness was affected at concentrations >30 μg/L. 

Effects observed in these mesocosm studies were all above the freshwater Final Chronic Value 

of 5.920 μg/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, Champia parvula, indicated 

that reproduction was not inhibited at the highest measured concentration tested, 167 μg/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 μg/L (Billinghurst et al. 

1998). The soft-shell clam, Mya arenaria, showed no adverse effects on survival from a 360-hr 

exposure at 700 μg/L (McLeese et al. 1980b). Granmo et al. (1989) report LC50s of 3,000 μg/L 

and 500 μg/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 μg/L 

(Granmo et al. 1989) and caused effects on attachment activity at 22 μg/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 LC50s for various timepoints 
24 

(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 μg/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 μg/L for the copepod, Acartia tonsa, exposed to nonylphenol 

in a synthetic media. Nice et al (2000) reported developmental effects at 100 μg/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 μg/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 μg/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 μg/L exposure group. The authors calculated a zero 

population growth value (λ) of 19 μg/L for the 55-day multigenerational test. The chronic values 

from the 28-day exposure in this study was 12.02 μg/L (Table 2) and from a similar 28-day study 

by Ward and Boeri (1991b) was 5.112 μg/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 
25 

were in general agreement with freshwater trout test results. In four tests, LC50 values ranged 

from 130 to 900 μg/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, LC50s ranged from > 420 μg/L for a 24-hr exposure to 320 μg/L 

for a 72-hr exposure. Threespine stickleback, Gasterosteus aculeatus, exposed to a commercial 

mixture of nonylphenol had a 96-hr LC50 of 370 μg/L (Granmo et al. 1991a). 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 μg/L for 96 hr were all abnormally developed at 10 days post-fertilization. The LC50 

for the same exposure period was 5,444 μg/L. Killifish larva were similar in sensitivity to 

nonylphenol exposures at post hatch ages of 1, 14, and 28 days with LC50s 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. 
26 

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 ovotestis). 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 μg/L, but estrogenic effects were 

reported to have occurred in the amphipod Corophium volutator (Brown et al. 1999) at 10 μg/L 

and in the larva of Chironomus riparius (Hahn et al. 2002) at 2,000 μg/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, 17β-estradiol. Ren et al. (1996a) demonstrated significant increases in 

vitellogenin production in rainbow trout exposed to nonylphenol at 100 μg/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 μg/L. 

Similarly, Lech et al. (1996) observed a significant increase in vitellogenin mRNA at 72 hr in 

rainbow trout at 14.14 μg/L nonylphenol. Vitellogenin was induced in green swordfish, 

Xiphophorus helleri, by exposure to 4 μg/L of technical grade nonylphenol (Kwak et al. 2001). 

Jobling et al. (1996) demonstrated significant increases in vitellogenin in male rainbow trout, 
27 

Oncorhynchus mykiss, at three weeks of exposure to 20.3 and 54.3 μg/L of nonylphenol. Similar 

results were reported in another study with rainbow trout, plasma vitellogenin was increased 

after 21 days exposure to 50 μg/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 μg/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 μg/L of nonylphenol (Pedersen et al. 1999) 

or in Atlantic salmon, Salmo trutta, exposed for 30 days to 20 μg/L (Moore et al. 2003). The 

influence of exposure route on nonylphenol-induced vitellogenin mRNA 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 μg/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 μg/L did appear to reduce fecundity. However, fish exposed to 0.09 and 0.1 μg/L produced 

more eggs than control fish. These data appear to produce a U-shaped dose-response and 

indicate a possible hormetic response of fecundity to nonylphenol. Nonylphenol concentrations 

of 0.05 to 3.4 μg/L did not significantly change vitellogenin concentrations in the blood of 

males, and raised the 17β-estradiol titers in the blood of male and female fish at most treatment 

concentrations > 0.05 μg/L. An increase in the number of Sertoli cells may have occurred in the 

male fathead minnow exposed to nonylphenol at 1.6 μg/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 μg/L, survival, growth, egg 

production, egg viability, and gonadosomatic index (GSI) were not altered. In another study 

with the same species of fish, development of ovo-testis, an intersex condition, occurred after a 
28 

three month exposure to 50 μg/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 of fish exposed to nonylphenol from fertilized egg to 60 days post-hatch, the 

lowest-observed-effect concentration for vitellogenin induction was 11.6 μg/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-days post-hatch of the second (F1) 

generation. Five exposure concentrations of nonylphenol in quadruplicate (4.2, 8.2, 17.7, 51.5, 

and 183 μg/L) and water-only and solvent controls were used. In the F0 generation, egg 

hatchability was reduced (46.7%) by 183 μg/L nonylphenol exposure, survival was significantly 

decreased at 60 days post-hatch by nonylphenol exposures > 17.7 μg/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 μg/L treatment, with 20% of fish displaying male characteristics externally 

having ovo-testis tissues. In fish from the 51.5 μg/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 μg/L nonylphenol, but was not observed in the fish with ovotestis exposed to 51.5 μg/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 μg/L, but not significantly, and GSI of female fish was increased significantly by 

exposure to nonylphenol concentrations > 8.2 μg/L. 

The effects of nonylphenol on F1 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-days post-hatch by any of the nonylphenol exposure 

concentrationsl. The sex ratio, characterized by secondary sex characteristics, changed in 

treatments > 17.7 μg/L to favor females 1:2. Induction of ovo-testis was observed at lower 

concentrations of nonylphenol in the F1 generation than in the F0 generation. Ovo-testis were 

observed in the 8.2 μg/L exposure group (10%) and in the 17.7 μg/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 F0 generation in the 

17.7 μg/L treatment. The overall results indicate a LOEC of 17.7 μg/L and a NOEC of 8.2 μg/L 
29 

with a chronic value, calculated as the geometric mean of the NOEC and LOEC, of 12.05 μg/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 μg/L over a 4 month period. Mortality rate in the progeny was 

significantly reduced by parental exposure to both 1 and 10 μg/L nonylphenol and hatching in 

the progency was reduced by parental exposure to 10 μg/L nonylphenol. Vitellogenin was 

induced (approximately 10-fold) in adult male fish exposed to both 1 and 10 μg/L nonylphenol. 

In the male progeny of parental fish exposed to 10 μg/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 

μg/L. 
30 

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 (GrossSorokin 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) 

Hansen et al. (1998) 

Kopf (1997) 

Magliulo et al. (1998) 

Muller (1980) 

Palmer et al. (1998) 

Weinberger and Rea (1981) 

Nonylphenol was a Component of a Mixture or Sediment

Ahel et al. (1993) 

Amato and Wayment (1998) 

Bettinetti and Provini (2002) 

Fay et al. (2000) 

Dwyer et al. (1999a,b) 

Escher et al. (1999) 

Hansen et al. (1999) 

Larsson et al. (1999) 

Moore et al. (1987) 

Purdom et al. (1994) 

Sundaram et al. (1980) 

Turner et al. (1985) 

Ward and Boeri (1992) 

Studies were Conducted with Ethoxylated Nonylphenols

Baldwin et al. (1998) 

Braaten et al. (1972) 

Dorn et al. (1993) 

Maki et al. (1998) 

Manzano et al. (1998, 1999) 

Patoczka and Pulliam 1999 
31 

Organisms were Dosed by Injection, Gavage or in Artifical Medium

Arukwe et al. (1997a,b;1998) 

Christiansen et al. (1998a,b,c; 

1999) 

Coldham et al. (1997, 1998) 

Haya et al. (1997) 

Madsen et al. (1997) 

Nimrod and Benson (1996; 

1997) 

Rice et al. (1998) 

Spieser et al. (1998) 

Thibaut et al. (1998) 

Weinberger et al. (1987) 

Yadetie et al. (1999) 

Experimental Model was Plasma, Enzymes, Receptors, Tissues or Cell Cultures

Andersen et al. (1999) 

Celius et al. (1999) 

Flouriot et al. (1995) 

Hewitt et al. (1998) 

Jobling et al. (1996) 

Jobling and Sumpter (1993) 

Knudsen and Pottinger (1999) 

Lamche and Burkhardt-Holm 

(2000) 

Levine and Cheney (2000) 

Loomis and Thomas (1999) 

Lutz and Kloas (1999) 

Milligan et al. (1998) 

Petit et al. (1997, 1999) 

Routledge and Sumpter (1996, 

1997) 

Soto et al. (1991, 1992) 

White et al. (1994) 

Data were Compiled from Other Source

Bearden and Schultz (1997, 

1998) 

Lewis (1991) 

Liber et al. (1999) 

Varma and Patel (1988) 

Veith and Mekenyan (1993 
32 

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 μg/L for 

the amphipod Hyalella azteca to 774 μg/L for the snail Physella virgata. Eleven species of fish 

were tested and were in the mid-range of sensitivity (SMAVs = 110 to 289.3 μg/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 μg/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 μg/L based on growth, and the fathead minnow, Pimephales promelas, 

had a chronic value of 10.18 μg/L based on survival. Two species of freshwater cladocerans 

were tested and chronic values ranged from 22.62 to 157.9 μg/L based on reproduction. One 

species of freshwater midge was tested and its chronic value was 61.82 μg/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 μg/L for the winter 

flounder, Pleuronectes americanus, to 209.8 μg/L for the sheepshead minnow, Cyprinodon 

variegatus. These two fish species were the only fish were tested. Nine different species of 
33 

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 μg/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 μg/L based on reduced growth and 12.02 μg/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 μg/L for the freshwater alga to 410 μg/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 μg/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.
34 

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 μg/L more than once every three years on the 

average and if the four-day average concentration of nonylphenol does not exceed 6.6 μg/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 μg/L more than once every three years on the 

average and if the four-day average concentration of nonylphenol does not exceed 1.7 μg/L more 

than once every three years on the average. 
35 

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). 
36 

Genus Mean Acute Values

10

100

1000

Nonylphenol Effect Concentration (μg/L)

Figure 1. Summary of Ranked Nonylphenol GMAVs

Freshwater

 Inv ertebrates

 Fish

 Amphibians

Freshwater Final Acute Value = 55.49 μg/L Nonylphenol

Criteria Maximum Concentration = 27.75 μg/L Nonylphenol

(Cumulative Fraction)
37 

Genus Mean Acute Values

1

10

100

1000

Nonylphenol Effect Concentration (μg/L)

Figure 2. Summary of Ranked Nonylphenol GMAVs

Saltwater

 Inv ertebrate

 Fish

Saltwater Final Acute Value = 13.93 μg/L Nonylphenol

Criteria Maximum Concentration = 6.965 μg/L Nonylphenol

(Cumulative Fraction)
38 

Genus Mean Chronic Values

1

10

100

1000

Nonylphenol Effect Concentration (μg/L)

Figure 3. Chronic Toxicity of Nonylphenol to Aquatic Animals

 Freshwater Inv ertebrates

 Freshwater Fish

 Saltwater Inv ertebrates (Cumulative Fraction)

Freshwater Final Chronic Value = 6.596 μg/L Nonylphenol

Saltwater Final Chronic Value = 1.656 μg/L Nonylphenol
39 

Table 1. Acute Toxicity of Nonylphenol to Aquatic Animals 

Species Methoda Chemical pH

LC50 

or EC50 

(:g/L)

Species 

Mean Acute 

Valueb

(:g/L) Reference

FRESHWATER SPECIES

Annelid (adult), 

Lumbriculus variegatus

F,M >90% 6.75 342 342 Brooke 1993a 

Snail (adult), 

Physella virgata

F,M >90% 7.89 774 774 Brooke 1993a 

Cladoceran 

(<24-hr old), 

Daphnia magna

R,M >90% 7.87 104 - Brooke 1993a 

Cladoceran 

(<24-hr old), 

Daphnia magna

S,M 91.8% 8.25 190 140.6 Comber et al. 

1993 

Midge (2nd instar), 

Chironomus tentans

F,M >95% 8.0-8.4 160 160 England and 

Bussard 1995 

Dragonfly (nymph), 

Ophiogomphus sp. 

F,M >90% 8.06 596 596 Brooke 1993a 

Amphipod, 

(juvenile, 2mm TL), 

Hyalella azteca

F,M >90% 7.80 20.7 - Brooke 1993a 

Amphipod 

(juvenile, 2-3mm TL), 

Hyalella azteca

F,M >95% 7.9-8.7 150 55.72 England and 

Bussard 1995 

Rainbow trout 

(0.67 " 0.35 g), 

Oncorhynchus mykiss

S,U 85% 7.8-7.9 190 - Dwyer et al. 

1995 

Rainbow trout 

(1.25 " 0.57 g), 

Oncorhynchus mykiss

S,U 85% 7.5-7.7 260 - Dwyer et al. 

1995 

Rainbow trout 

(0.27 " 0.07 g), 

Oncorhynchus mykiss

S,U 85% 7.9 140 - Dwyer et al. 

1995 

Rainbow trout 

(1.09 " 0.38 g), 

Oncorhynchus mykiss

S,U 85% 7.7-7.9 270 - Dwyer et al. 

1995 
40 

Table 1. Acute Toxicity of Nonylphenol to Aquatic Animals 

Species Methoda Chemical pH

LC50 

or EC50 

(:g/L)

Species 

Mean Acute 

Valueb

(:g/L) Reference

Rainbow trout 

(0.48 " 0.08 g), 

Oncorhynchus mykiss

S,U 85% 7.5-7.9 160 - Dwyer et al. 

1995 

Rainbow trout 

(0.50 " 0.21 g), 

Oncorhynchus mykiss

S,U 85% 6.5-7.9 180 - Dwyer et al. 

1995 

Rainbow trout 

(45 d), 

Oncorhynchus mykiss

F,M >90% 6.72 221 221 Brooke 1993a 

Apache trout 

(0.85 " 0.49 g), 

Oncorhynchus apache

S,U 85% 7.8-7.9 180 - Dwyer et al. 

1995 

Apache trout 

(0.38 " 0.18 g), 

Oncorhynchus apache

S,U 85% 7.3-7.7 160 169.7 Dwyer et al. 

1995 

Greenback cutthroat 

trout (0.31 " 0.17 g), 

Oncorhynchus clarki 

stomais

S,U 85% 7.5-7.6 150 - Dwyer et al. 

1995 

Lahontan cutthroat 

trout 

(0.34 " 0.08 g), 

Oncorhynchus clarki 

henshawi

S,U 85% 7.9 140 - Dwyer et al. 

1995 

Lahontan cutthroat 

trout 

(0.57 " 0.23 g), 

Oncorhynchus clarki 

henshawi

S,U 85% 7.6-7.7 220 166.6 Dwyer et al. 

1995 

Fathead minnow 

(0.32 " 0.16 g), 

Pimephales promelas

S,U 85% 7.7-8.1 210 - Dwyer et al. 

1995 

Fathead minnow 

(0.56 " 0.19 g), 

Pimephales promelas

S,U 85% 7.8-8.1 360 - Dwyer et al. 

1995 

Fathead minnow 

(0.45 " 0.35 g), 

Pimephales promelas

S,U 85% 7.6-7.8 310 - Dwyer et al. 

1995 
41 

Table 1. Acute Toxicity of Nonylphenol to Aquatic Animals 

Species Methoda Chemical pH

LC50 

or EC50 

(:g/L)

Species 

Mean Acute 

Valueb

(:g/L) Reference

Fathead minnow 

(0.40 " 0.21 g), 

Pimephales promelas

S,U 85% 7.5-7.9 330 - Dwyer et al. 

1995 

Fathead minnow 

(0.34 " 0.24 g), 

Pimephales promelas

S,U 85% 7.5-7.6 170 - Dwyer et al. 

1995 

Fathead minnow 

(0.39 " 0.14 g), 

Pimephales promelas

S,U 85% 7.8-8.2 290 - Dwyer et al. 

1995 

Fathead minnow 

(32 d), 

Pimephales promelas

F,M 99% 7.29 140 - Holcombe et 

al. 1984; 

Univ. Wisc.-

Superior 1985 

Fathead minnow 

(25-35 d), 

Pimephales promelas

F,M >90% 7.23 128 133.9 Brooke 1993a 

Bonytail chub 

(0.29 " 0.08 g), 

Gila elegans

S,U 85% 7.7-7.9 270 - Dwyer et al. 

1995 

Bonytail chub 

(0.52 " 0.09 g), 

Gila elegans

S,U 85% 7.4-7.6 310 289.3 Dwyer et al. 

1995 

Colorado squawfish 

(0.32 " 0.05 g), 

Ptychocheilus lucius

S,U 85% 8.1-8.2 240 - Dwyer et al. 

1995 

Colorado squawfish 

(0.34 " 0.05 g), 

Ptychocheilus lucius

S,U 85% 7.8-8.0 270 254.6 Dwyer et al. 

1995 

Razorback sucker (0.31 

" 0.04 g), 

Xyrauchen texanus

S,U 85% 7.8-8.1 160 - Dwyer et al. 

1995 

Razorback sucker (0.32 

" 0.07 g), 

Xyrauchen texanus

S,U 85% 7.9-8.0 190 174.4 Dwyer et al. 

1995 

Gila topminnow 

(0.219 g, 27.2 mm), 

Poeciliopsis 

occidentalis

S,U 85% 8.0 230 230 Dwyer et al. 

1999a 
42 

Table 1. Acute Toxicity of Nonylphenol to Aquatic Animals 

Species Methoda Chemical pH

LC50 

or EC50 

(:g/L)

Species 

Mean Acute 

Valueb

(:g/L) Reference

Fountain darter 

(0.062 g, 20.2 mm), 

Etheostoma rubrum

S,U 85% 8.0-8.1 110 110 Dwyer et al. 

1999a 

Greenthroat darter 

(0.133 g, 22.6 mm), 

Etheostoma lepidum

S,U 85% 8.0-8.2 190 190 Dwyer et al. 

1999a 

Bluegill (juvenile), 

Lepomis macrochirus

F,M >90% 7.61 209 209 Brooke 1993a 

Boreal toad 

(0.012 g, 9.6 mm), 

Bufo boreas

S,U 85% 7.9-8.0 120 120 Dwyer et al. 

1999a 

SALTWATER SPECIES

Coot clam 

(embryo/larva), 

Mulinia lateralis

S,U 90% 7.8-8.2 37.9 37.9 Lussier et al. 

2000 

Copepod (10-12 d), 

Acartia tonsa

S,U - - 190 190 Kusk and 

Wollenberger 

1999 

Mysid (<24-hr old), 

Americamysis bahia

F,M >95% 7.3-8.2 43 - Ward and 

Boeri 1990a 

Mysid (<24-hr old), 

Americamysis bahia

F,M 90% 7.8-8.2 60.6 51.05 Lussier et al. 

2000 

Amphipod (adult), 

Leptocheirus 

plumulosus

F,M 90% 7.8-8.2 61.6 61.6 Lussier et al. 

2000 

Amphipod (adult), 

Eohaustorius estuarius

S,U - - 138 138 Hecht and 

Boese 2002 

Grass shrimp 

(48-hr old), 

Palaemonetes vulgaris

F,M 90% 7.8-8.2 59.4 59.4 Lussier et al. 

2000 

American lobster 

(1st stage), 

Homarus americanus

R,U 90% 7.8-8.2 71 71 Lussier et al. 

2000 

Mud crab 

(4th and 5th stages), 

Dyspanopeus sayii

F,M 90% 7.8-8.2 >195 >195 Lussier et al. 

2000 
43 

Table 1. Acute Toxicity of Nonylphenol to Aquatic Animals 

Species Methoda Chemical pH

LC50 

or EC50 

(:g/L)

Species 

Mean Acute 

Valueb

(:g/L) Reference

Winter flounder 

(48-hr-old), 

Pleuronectes 

americanus

S,M 90% 7.8-8.2 17 17 Lussier et al. 

2000 

Sheepshead minnow 

(juvenile), 

Cyprinodon variegatus

F,M >95% 7.4-8.1 310 - Ward and 

Boeri 1990b 

Sheepshead minnow 

(juvenile), 

Cyprinodon variegatus

F,M 90% 7.8-8.2 142 209.8 Lussier et al. 

2000 

Inland silversides 

(juvenile), 

Menidia beryllina

F,M 90% 7.8-8.2 70 70 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. 
44 

Table 2. Chronic Toxicity of Nonylphenol to Aquatic Animals

Species Testa Chemical pH 

Chronic 

Limits 

(μg/L)b

Chronic 

Value 

(μg/L) Reference

Cladoceran, 

Ceriodaphnia dubia

LC >95% 8.3-8.6 88.7-202 133.9 England 1995 

Cladoceran, 

Daphnia magna

LC 93.1 8.04 10.7-47.8 22.62 Fliedner 1993 

Cladoceran, 

Daphnia magna

LC >90% 8.46 116-215 157.9 Brooke 1993a 

Cladoceran, 

Daphnia magna

LC 91.8% 8.25 24-39 30.59 Comber et al. 

1993 

Midge, 

Chironomus tentans

LC 95% 7.73 42-91 61.82 Kahl et al. 1997 

Rainbow trout, 

Oncorhynchus mykiss

ELS >90% 6.97 6.0-10.3 7.861 Brooke 1993a 

Fathead minnow, 

Pimephales promelas

ELS >95% 7.1-8.2 7.4-14 10.18 Ward and Boeri 

1991c 

SALTWATER SPECIES

Mysid, 

Americamysis bahia

LC >95% 7.4-8.3 3.9-6.7 5.112 Ward and Boeri 

1991b 

Mysid, 

Americamysis bahia

LC - - 9.46-15.28 12.02 Kuhn et al. 2001 

a

 LC = life-cycle or partial life-cycle; ELS = early life-stage. b

 Based upon measured concentrations of nonylphenol. 
45 

Table 2. Acute-Chronic Ratios

Acute-Chronic Ratios

Species pH 

Acute Value 

 (μg/L) 

Chronic Value 

 (μg/L) Ratio Reference

Cladoceran, 

Daphnia magna

7.87-8.46 104 157.9 0.6586 Brooke 1993a 

Cladoceran, 

Daphnia magna

8.25 190 30.59 6.211 Comber et al. 1993 

Mysid, 

Americamysis bahia

7.3-8.3 43 5.112 8.412 Ward and Boeri 

1990a, 1991b 

Rainbow trout, 

Oncorhynchus mykiss

6.72-6.97 221 7.861 28.11 Brooke 1993a 
46 

Table 3. Ranked Genus Mean Acute Values with Species Mean Acute-Chronic Ratios 

Ranka

Genus Mean 

Acute Value 

 (μg/L) Species

Species Mean 

Acute Value 

 (μg/L)b 

Species Mean 

Acute-Chronic 

 Ratioc 

FRESHWATER SPECIES

15 774 Snail, 

Physella virgata

774 - 

14 596 Dragonfly, 

Ophiogomphus sp.

596 - 

13 342 Annelid, 

Lumbriculus variegatus

342 - 

12 289.3 Bonytail chub, 

Gila elegans

289.3 - 

11 254.6 Colorado squawfish, 

Ptychocheilus lucius

254.6 - 

10 230 Gila topminnow, 

Poeciliopsis occidentalis

230 - 

9 209 Bluegill, 

Lepomis macrochirus

209 - 

8 184.2 Rainbow trout, 

Oncorhynchus mykiss

221 28.11 

 Apache trout, 

Oncorhynchus apache

169.7 - 

 Lahontan cutthroat trout, 

Oncorhynchus clarki henshawi, 

and 

166.6 - 

 Greenback cutthroat trout, 

Oncorhynchus clarki stomais

- - 

7 174.4 Razorback sucker, 

Xyrauchen texanus

174.4 - 

6 160 Midge, 

Chironomus tentans

160 - 

5 144.6 Greenthroat darter, 

Etheostoma lepidum

190 - 

 Fountain darter, 

Etheostoma rubrum

110 - 

4 140.6 Cladoceran, 

Daphnia magna 

140.6 2.023 
47 

Table 3. Ranked Genus Mean Acute Values with Species Mean Acute-Chronic Ratios 

Ranka

Genus Mean 

Acute Value 

 (μg/L) Species

Species Mean 

Acute Value 

 (μg/L)b 

Species Mean 

Acute-Chronic 

 Ratioc 

3 133.9 Fathead minnow, 

Pimephales promelas

133.9 - 

2 120 Boreal toad, 

Bufo boreas

120 - 

1 55.72 Amphipod, 

Hyalella azteca

55.72 - 

SALTWATER SPECIES

11 209.8 Sheepshead minnow, 

Cyprinodon variegatus

209.8 - 

10 >195 Mud crab, 

Dyspanopeus sayii

>195 - 

9 190 Copepod, 

Acartia tonsa

190 

8 138 Amphipod, 

Eohaustorius estuarius

138 - 

7 71 American lobster, 

Homarus americanus

71 - 

6 70 Inland silversides, 

Menidia beryllina

70 - 

5 61.6 Amphipod, 

Leptocheirus plumulosus

61.6 - 

4 59.4 Grass shrimp, 

Palaemonetes vulgaris

59.4 - 

3 51.05 Mysid, 

Americamysis bahia

51.05 8.412 

2 37.9 Coot clam, 

Mulinia lateralis

37.9 - 

1 17 Winter flounder, 

Pleuronectes americanus

17 - 

a

 Ranked from the most resistant to the most sensitive based on Genus Mean Acute Value. 

b

 From Table 1. 

c

 From Table 2. 
48 

Table 3. Ranked Genus Mean Acute Values with Species Mean Acute-Chronic Ratios

Freshwater 

Final Acute Value = 55.49 μg/L 

Criterion Maximum Concentration = 55.49 ÷ 2 = 27.75 μg/L 

Final Acute-Chronic Ratio = 8.412 (see text) 

Final Chronic Value = 55.49 μg/L ÷ 8.412 = 6.5965 μg/L 

Saltwater 

Final Acute Value = 13.93 μg/L 

Criterion Maximum Concentration = 13.93 ÷ 2 = 6.965 μg/L 

Final Acute-Chronic Ratio = 8.412 (see text) 

Final Chronic Value = 13.93 μg/L ÷ 8.412 = 1.6560 μg/L 
49 

Table 4. Toxicity of Nonylphenol to Aquatic Plants

Species Chemical pH 

Duration 

 (days) Effect

Concentration 

 (μg/L) Reference

FRESHWATER SPECIES

Green algae, 

Selenastrum 

capricornutum

>95% 7.8 4 EC50, 

number of 

cells 

410 Ward and Boeri 

1990a 

SALTWATER SPECIES

Diatom, 

Skeletonema 

costatum

>95% 30a

 4 EC50, 

number of 

cells 

27 Ward and Boeri 

1990d 

a

Salinity (g/kg). 
50 

Table 5. Bioaccumulation of Nonylphenol by Aquatic Organisms

Species Chemical

Water

Conc. 

(μg/L)a pH

Duration 

 (days) Tissue

Percent 

Lipids 

BCF 

or 

BAFb

Normalized 

BCF or 

 BAFc Reference

FRESHWATER SPECIES

Fathead 

minnow 

(0.5-1 g), 

Pimephales 

promelas

>95% 4.9 7.0-7.6 27 Whole 

body 

- 271 - Ward and 

Boeri 1991a 

Fathead 

minnow 

(0.5-1 g), 

Pimephales 

promelas

>95% 22.7 7.0-7.6 27 Whole 

body 

- 344 - Ward and 

Boeri 1991a 

Fathead 

minnow 

(4-wk old), 

Pimephales 

promelas

99% 18.4 7.62 4 Whole 

body 

4.7±1.7 751 159.8 Brooke 

1993b 

Fathead 

minnow 

(4-wk old), 

Pimephales 

promelas

99% 41.9 7.62 4 Whole 

body 

4.7±1.7 677 144.0 Brooke 

1993b 

Fathead 

minnow 

(4-wk old), 

Pimephales 

promelas

99% 82.1 7.62 4 Whole 

body 

4.7±1.7 945 201.1 Brooke 

1993b 

Fathead 

minnow 

(4-wk old), 

Pimephales 

promelas

99% 9.3 7.60 28 Whole 

body 

4.7±1.7 769 163.6 Brooke 

1993b 

Fathead 

minnow 

(4-wk old), 

Pimephales 

promelas

99% 19.2 7.60 28 Whole 

body 

4.7±1.7 984 209.4 Brooke 

1993b 
51 

Table 5. Bioaccumulation of Nonylphenol by Aquatic Organisms

Species Chemical

Water

Conc. 

(μg/L)a pH

Duration 

 (days) Tissue

Percent 

Lipids 

BCF 

or 

BAFb

Normalized 

BCF or 

 BAFc Reference

Fathead 

minnow 

(4-wk old),

Pimephales 

promelas

99% 38.1 7.60 28 Whole 

body 

4.7±1.7 876 186.4 Brooke 

1993b 

Fathead 

minnow 

(4-wk old), 

Pimephales 

promelas

99% 77.5 7.60 28 Whole 

body 

4.7±1.7 603 128.3 Brooke 

1993b 

Fathead 

minnow 

(adult), 

Pimephales 

promelas

>98% 

0.4 

1.6 

3.4 

- 42 

Whole 

body - 

203 

252 

268 

- 

- 

- 

Giesy et al. 

2000 

Bluegill 

(4-wk old), 

Lepomis 

macrochirus

99% 21.6 7.79 4 Whole 

body 

4.9±1.5 279 56.94 Brooke 

1993b 

Bluegill 

(4-wk old), 

Lepomis 

macrochirus

99% 43.9 7.79 4 Whole 

body 

4.9±1.5 257 52.45 Brooke 

1993b 

Bluegill 

(4-wk old), 

Lepomis 

macrochirus

99% 86.5 7.79 4 Whole 

body 

4.9±1.5 223 45.51 Brooke 

1993b 

Bluegill 

(4-wk old), 

Lepomis 

macrochirus

99% 5.6 7.55 28 Whole 

body 

4.9±1.5 231 47.14 Brooke 

1993b 

Bluegill 

(4-wk old), 

Lepomis 

macrochirus

99% 12.4 7.55 28 Whole 

body 

4.9±1.5 253 51.63 Brooke 

1993b 

Bluegill 

(4-wk old), 

Lepomis 

macrochirus

99% 27.6 7.55 28 Whole 

body 

4.9±1.5 250 51.02 Brooke 

1993b 
52 

Table 5. Bioaccumulation of Nonylphenol by Aquatic Organisms

Species Chemical

Water

Conc. 

(μg/L)a pH

Duration 

 (days) Tissue

Percent 

Lipids 

BCF 

or 

BAFb

Normalized 

BCF or 

 BAFc Reference

Bluegill 

(4-wk old), 

Lepomis 

macrochirus

99% 59.5 7.55 28 Whole 

body 

4.9±1.5 191 38.98 Brooke 

1993b 

Bluegill 

(juvenile), 

Lepomis 

macrochirus

96.4% 

1.0 

3.0 

30.0 

7.7 20 Whole 

body 

0.72± 

0.46 

76 

60 

37 

105.6 

83.33 

51.39 

Liber et al. 

1999 

SALTWATER SPECIES

Blue mussel, 

Mytilus 

edulis

14Clabeled 

5.9 - 16 Whole 

body 

1.6 2,740 1,712 Ekelund et 

al. 1990 

Blue mussel, 

Mytilus 

edulis

14Clabeled 

6.2 - 16 Whole 

body 

1.9 4,120 2,168 Ekelund et 

al. 1990 

Common 

shrimp, 

Crangon 

crangon

d

14Clabeled 

6.4 - 16 Whole 

body 

1.4 110 78.75 Ekelund et 

al. 1990 

Common 

shrimp, 

Crangon 

crangon

d

14Clabeled 

7.4 - 16 Whole 

body 

1.7 900 529.4 Ekelund et 

al. 1990 

Three-spined 

stickleback, 

Gasterosteus 

aculeatus

14Clabeled 

4.8 - 16 Whole 

body 

6.7 1,200 179.1 Ekelund et 

al. 1990 

Three-spined 

stickleback, 

Gasterosteus 

aculeatus

14Clabeled 

4.9 - 16 Whole 

body 

7.8 1,300 166.7 Ekelund et 

al. 1990 
53 

Table 5. Bioaccumulation of Nonylphenol by Aquatic Organisms

Species Chemical

Water

Conc. 

(μg/L)a pH

Duration 

 (days) Tissue

Percent 

Lipids 

BCF 

or 

BAFb

Normalized 

BCF or 

 BAFc Reference

a

Measured concentration of nonylphenol. 

b

Bioconcentration factors (BCFs) and bioaccumulation factors (BAFs) are based on measured concentrations of 

nonylphenol in water and in tissue. c

When possible, the factors were normalized to 1% lipid by dividing the BCFs and BAFs by the percent lipid measured 

in the test organism. d

Non-resident species. 
54 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

FRESHWATER SPECIES

Phytoplankton 

and Periphyton 

>98% 8.8 - 10.6 6 wk Dominant 

species changed 

29 - 120 Hense et al. 

2003 

Green alga, 

Chlamydomonas 

reinhardtii

- - 24 days 100% algistatic 6,250 Weinberger and 

Greenhalgh 

1984 

Floating moss, 

Salvinia molesta

- - 9 days Reduced frond 

production 

2,500 Prasad 1986 

Duckweed, 

Lemna minor

- 5.6 96 hr IC50 5,500 Weinberger and 

Iyengar 1983 

Duckweed, Lemna 

minor

- - 4 days Reduced frond 

production

125 Prasad 1986

Ciliate protozoan, 

Tetrahymena 

pyriformis

- - 24 hr EC50 460 Yoshioka 1985

Ciliate protozoan, 

Tetrahymena 

pyriformis

- 7.40 40 hr Reduced 

population 

growth 50%

747 Schultz 1997

Rotifer 

(4 to 6 hr-old 

female) 

Brachionus 

calyciflorus

Technical 7.5 96 hr Sexual 

reproduction 

reduced

50 Preston et al. 

2000

Clam (15 g), 

Anodonta 

cataractae

- - 144 hr 

(fed)

LC50 5,000 McLeese et al. 

1980b

Zooplankton 96.4% 7.5 - 8.2 20 days NOEC 

LOEC

23 

76

O’Halloran et 

al. 1999

Zooplankton >98% 8.8 - 10.4 6 wk NOEC 19 - 44 Severin et al. 

2003

Benthic macroinvertebrates96.4% 7.5 - 8.2 20 days NOEC 

LOEC

23 

76

Schmude et al. 

1999

Cladoceran 

(<24-hr old), 

Daphnia magna

- 8.0 21 days NOEC 

LOEC (reduced 

fecundity)

50 

100

Baldwin et al. 

1997
55 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Cladoceran 

(<24-hr old and 

adults), 

Daphnia magna

~85% 7.8 - 8.4 96 hr 

(fed)

MATC (young) 

MATC (adults)

302 

136

Gerritsen et al. 

1998

Cladoceran 

(<24-hr old), 

Daphnia magna

~85% 7.7+0.02 21 days No sex ratio 

change (high 

food rate) 

Increased ratio 

of males (low 

food rate)

25 

25

Baer and 

Owens 1999

Cladoceran 

(<24-hr old), 

Daphnia magna

Technical - 21 days 50% adult 

mortality 

NOEC 

(deformed 

offspring)

200.5 

44 

LeBlanc et al. 

2000

Cladoceran 

(<24-hr old), 

Daphnia magna

~85% - 48 hr EC50 234 

272 

337

Zang et al. 2003

Cladoceran 

(<24-hr old), 

Daphnia magna

~85% - 35 day LOEC >50 Zang et al. 2003 

Cladoceran 

(<36-hr old), 

Daphnia galeata 

mendotae

- - 30 day NOEC 

LOEC 

(deformed 

offspring) 

10 

50 

Shurin and 

Dodson 1997 

Cladoceran 

(>48-hr old),

Daphnia pulex

Practical 

grade 

- 48 hr LC50 140 Ernst et al. 1980 

Cladoceran 

(>48-hr old), 

Daphnia pulex

Practical 

grade 

- 48 hr LC50 176 Ernst et al. 1980 

Cladoceran 

(>48-hr old), 

Daphnia pulex

Practical 

grade 

- 48 hr LC50 190 Ernst et al. 1980 

Cladoceran 

(<24-hr old), 

Ceriodaphnia 

dubia

>95% 8.3-8.6 48 hr LC50 

(fed) 

276 England 1995 
56 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Cladoceran 

(<24-hr old), 

Ceriodaphnia 

dubia

>95% 8.3-8.6 7 days LC50 (fed) 225 England 1995 

Midge 

(2nd instar), 

Chironomus 

tentans

>95% 8.2 14 days LC50 

EC50 

119 

95 

England and 

Bussard 1993 

Sea lamprey 

(larva), 

Petromyzon 

marinus

- 7.5-8.2 14 hr LT100 5,000 Applegate et al. 

1957 

Brook trout 

(juvenile), 

Salvelinus 

fontinalis

- - 96 hr LC50 145 Holmes and 

Kingsbury 1980 

Lake trout 

(juvenile), 

Salvelinus 

naymaycush

- - 35 days LC50 

(fed) 

>40 Holmes and 

Kingsbury 1980

Brown trout 

(fingerling), 

Salmo trutta

- 7.0 2 hr LT100 5,000 Wood 1953 

Atlantic salmon 

(4 g), 

Salmo salar

- - 96 hr 

(fed) 

LC50 900 McLeese et al. 

1980b 

Atlantic salmon 

(48.3 + 2.6 mm TL), 

Salmo salar

- - 30 days No change in 

plasma 

vitellogenin or 

gill NaK ATPase 

activity or plasma 

Cl-

 and Na+

20 Moore et al. 

2003 

Chinook salmon 

(juvenile), 

Oncorhynchus 

tshawytscha

- 7.2 3 hr LT100 10,000 MacPhee and 

Ruelle 1969 

Rainbow trout 

(juvenile), 

Oncorhynchus 

mykiss

- 7.5-8.2 4 hr LT100 5,000 Applegate et al. 

1957 
57 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Rainbow trout 

(juvenile), 

Oncorhynchus 

mykiss

Practical 

grade 

- 96 hr LC50 920 Ernst et al. 1980 

Rainbow trout 

(juvenile), 

Oncorhynchus 

mykiss

Practical 

grade 

- 96 hr LC50 560 Ernst et al. 1980 

Rainbow trout 

(juvenile), 

Oncorhynchus 

mykiss

- - 96 hr LC50 230 Holmes and 

Kingsbury 1980 

Rainbow trout 

(adult males), 

Oncorhynchus 

mykiss

- - 3 wk Increased 

vitellogenin 

production 

20.3 Jobling et al. 

1996 

Rainbow trout 

(adult males), 

Oncorhynchus 

mykiss

- - 3 wk Increased 

vitellogenin 

production 

54.3 Jobling et al. 

1996 

Rainbow trout 

(50 - 200 g), 

Oncorhynchus 

mykiss

- - 72 hr LC50 193.65 Lech et al. 1996 

Rainbow trout 

(50 - 200 g), 

Oncorhynchus 

mykiss

- - 72 hr Increased 

vitellogenin 

mRNA 

14.14 Lech et al. 1996 

Rainbow trout, 

(40 - 60 g), 

Oncorhynchus 

mykiss

>99% 8 hr Tissue half-life 

 fat 19.8 hr 

 muscle 18.6 hr 

 liver 5.9 hr 

18 Lewis and Lech 

1996 

Rainbow trout, 

(40 - 60 g), 

Oncorhynchus 

mykiss

>99% 2 – 5 hr Eviscerated 

carcass 

BAF = 24.21 

18 Lewis and Lech 

1996 

Rainbow trout, 

(40 - 60 g), 

Oncorhynchus 

mykiss

>99% 2 – 5 hr Viscera 

BAF = 98.2 

18 Lewis and Lech 

1996 
58 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Rainbow trout 

(juvenile), 

Oncorhynchus 

mykiss

- - 4 hr Vitellogenin 

mRNA 

production 

10 Ren et al. 1996a 

Rainbow trout 

(juvenile), 

Oncorhynchus 

mykiss

- - 72 hr Vitellogenin 

mRNA 

production 

100 Ren et al. 1996a 

Rainbow trout 

(juvenile), 

Oncorhynchus 

mykiss

- 6.5 22 days 

35 days 

Reduced growth 

at 108 days 

Reduced growth 

at 466 days 

50 

30 

Ashfield et al. 

1998 

Rainbow trout 

(juvenile), 

Oncorhynchus 

mykiss

- - 96 hr Decreased 

number of 

muscarinic 

cholinergic 

receptors in 

brain 

220 Jones et al. 

1998 

Rainbow trout 

(35-50 g, 

immature), 

Oncorhynchus 

mykiss

- 8.0 – 8.4 21 days Increased 

vitellogenin in 

blood plasma 

50 Tremblay and 

Van Der Kraak 

1998 

Rainbow trout 

(adult males), 

Oncorhynchus 

mykiss

- - 3 wk BCF = 116 BCF 

= 88 

63 

81 

Blackburn et al. 

1999 

Rainbow trout 

(103-168 g, 

juvenile) 

Oncorhynchus 

mykiss

99% - 9 days No vitellogenin 

induction 

109 Pedersen et al. 

1999 

Rainbow trout 

(adult males), 

Oncorhynchus 

mykiss

Technical - 10 days per month 

for 4 months 

Epidermal 

mucous cell 

granulation 

1 BurkhardtHolm et al. 

2000 
59 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Rainbow trout 

(598 g; juvenile 

females), 

Oncorhynchus 

mykiss

99% - 18 wk Reduced GSI; 

Reduced HSI; 

Induced 

vitellogenin; 

Lowered 

plasma 

estradiol; 

Lowered 

plasma FSH 

85.6 

85.6 

8.3 

85.6 

8.3 

Harris et al. 

2001 

Rainbow trout 

(1667+201.6 g; 

F0 3 yr-old 

adults), 

Oncorhynchus 

mykiss

98% 7.6 4 months 

(exposed 10 

days/month)

Reduced 

embryo 

survival; 

Reduced hatch; 

Fo Males 

increased 

vitellogenin; 

F1 Females 

increased 

vitellogenin and 

testosterone; 

F1 Males 

increased 

estradiol 

1 

10 

1 

10 

10 

Schwaiger et al. 

2002 

Rainbow trout 

(6-mo-old), 

Oncorhynchus 

mykiss

- 7.2 4 wk Liver tissue 

showed 

hemorrhage and 

lymphocyte 

infiltration 

220 Uguz et al. 

2003 

Lahontan 

cutthroat trout 

(juvenile), 

Oncorhynchus 

clarki henshawi

- - 96 hr Decreased 

number of 

muscarinic 

cholinergic 

receptors in 

brain 

220 Jones et al. 

1998 

Apache trout 

(juvenile), 

Oncorhynchus 

mykiss

- - 96 hr Decreased 

number of 

muscarinic 

cholinergic 

receptors in 

brain 

>130 Jones et al. 

1998 

Northern 

squawfish 

(juvenile), 

Ptychocheilus 

oregonensis

- 7.2 3 hr LT100 10,000 MacPhee and 

Ruelle 1969 
60 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Colorado 

squawfish 

(juvenile), 

Ptychocheilus 

lucius

- - 96 hr Decreased 

number of 

muscarinic 

cholinergic 

receptors in 

brain 

>220 Jones et al. 

1998 

Goldfish 

(juvenile), 

Carassius auratus

- 7.0 5 hr LT100 5,000 Wood 1953 

Common carp 

(15.2 + 3.8 g 

juvenile), 

Cyprinus carpio

Technical 

(90% 4-

NP) 

7.6 70 days Decreased 

erythrocytes; 

Increased 

reticulocytes 

10 

 

10 

Schwaiger et al. 

2000 

Common carp 

(50-150 g mature 

males), 

Cyprinus carpio

95% 7.57 

+0.03 

28-31 days 

11 °C 

BCF = 546.5 

No change in 

17-estradiol, 

testosterone, or 

vitellogenin 

5.36 Villenueve et 

al. 2002 

Fathead minnow 

(4-wk old), 

Pimephales 

promelas

99% 7.62 4 days LC50 

(fed) 

138 Brooke 1993b 

Fathead minnow 

(4-wk old), 

Pimephales 

promelas

99% 7.60 28 days BCF = 100.4 193 Brooke 1993b 

Fathead minnow, 

Pimephales 

promelas

- - 96 hr Decreased 

number of 

muscarinic 

cholinergic 

receptors in 

brain 

>220 Jones et al. 

1998 

Fathead minnow 

(mature), 

Pimephales 

promelas

>98% - 42 days Possible 

increased 

number of 

Sertoli cells in 

males 

1.6 MilesRichardson et 

al. 1999 

Fathead minnow 

(mature), 

Pimephales 

promelas

>98% - 42 days Decreased 

fecundity 

>3.4 Giesy et al. 

2000 
61 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Fathead minnow 

(mature), 

Pimephales 

promelas

>98% - 42 days Increased ♂

vitellogenin 

>3.4 Giesy et al. 

2000 

Fathead minnow 

(mature), 

Pimephales 

promelas

>98% - 42 days Increased 

♂ & ♀ 

17β-estradiol

>0.05 

(not all test 

concentrations) 

Giesy et al. 

2000 

Bluegill 

(juvenile), 

Lepomis 

macrochirus

- 7.0 2 hr LT100 5,000 Wood 1953 

Bluegill 

(juvenile), 

Lepomis 

macrochiru

- 7.5-8.2 14 hr LT100 5,000 Applegate et al. 

1957 

Bluegill 

(4-wk old), 

Lepomis 

macrochirus

99% 7.79 4 days LC50 

(fed) 

135 Brooke 1993b 

Bluegill 

(4-wk old), 

Lepomis 

macrochirus

99% 7.55 28 days BCF=35.31 126 Brooke 1993b 

Bluegill 

(juvenile), 

Lepomis 

macrochirus

96.4% 7.7 - 7.9 20 days NOEC 

LOEC 

(survival) 

76 

243 

Liber et al. 

1999 

Southern platyfish 

(adult, 0.62 to 

1.15 g), 

Xiphophorus 

maculatus

Technical 

85% 

- 28 days Reduced GSI 960 Kinnberg et al. 

2000 

Green Swordtail 

(adult males), 

Xiphophorus 

helleri

Technical - 96 hr 

72 hr 

LC50 

Vitellogenin 

induced 

206 

4 

Kwak et al. 

2001 
62 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Green Swordtail 

(juvenile 30-d-old 

males), 

Xiphophorus 

helleri

Technical - 60 days Reduced sword 

length 

0.2 Kwak et al. 

2001 

African clawed 

frog (larva), 

Xenopus laevis

ACS 

Grade 

7.8 - 8.0 21 days NOEC 

LOEC 

(increased rate 

of tail 

resorption) 

25 

50 

Fort and Stover 

1997 

African clawed 

frog (larva), 

Xenopus laevis

- - 12 wk Increased 

female 

phenotypes 

22 Kloas et al. 

1999 

SALTWATER SPECIES

Red alga, 

Champia parvula

>95% - 2 days No effect on 

sexual 

reproduction 

167 Tagliabue 1993 

Barnacle 

(cypris larva), 

Balanus 

amphitrite

- - 48 h Reduced cyprid 

settlement 

1.0 Billinghurst et 

al. 1998 

Soft-shell clam, 

Mya arenaria

- - 360 hr No mortality 700 McLeese et al. 

1980b 

Coot clam, 

Mulinia lateralis

90% 30-31a 24 hr LC50 50 Lussier et al. 

2000 

Coot clam, 

Mulinia lateralis

90% 30-31a 48 hr LC50 50 Lussier et al. 

2000 

Coot clam, 

Mulinia lateralis

90% 30-31a 72 hr LC50 40 Lussier et al. 

2000 

Blue mussel, 

Mytilus edulis

90% 30-31a 96 hr LC50 3,000 Granmo et al. 

1989 

Blue mussel, 

Mytilus edulis

- 32a 360 hr LC50 500 Granmo et al. 

1989 

Blue mussel, 

Mytilus edulis

- 32a 13 days Reduced byssus 

strength 

56 Granmo et al. 

1989 

Blue mussel, 

Mytilus edulis

- 32a 30 days Reduced byssus 

strength 

56 Granmo et al. 

1989 
63 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Blue mussel, 

Mytilus edulis

- 32a 30 days No byssus 

threads formed 

100 Granmo et al. 

1989 

Blue mussel, 

Mytilus edulis

- 32a 32 days Reduction in 

growth 

56 Granmo et al. 

1989 

Blue mussel, 

Mytilus edulis

- 32a 24 hr No effect on 

fertilization 

200 Granmo et al. 

1989 

Blue mussel, 

Mytilus edulis

- 72 hr No effect on 

development 

200 Granmo et al. 

1989 

Blue mussel 

(40-50 mm 

length), Mytilus 

edulis

- - 50 days BCF = 350 40 Granmo et al. 

1991a,b 

Blue mussel, 

Mytilus edulis 

galloprovincialis

- - 2 days Repelled 

attachment 

22 Etoh et al. 1997 

Mysid, 

Americamysis 

bahia

90% 30-31a 24 hr LC50 ~114 Lussier et al. 

2000 

Mysid, 

Americamysis 

bahia

90% 30-31a 48 hr LC50 ~82 Lussier et al. 

2000 

Mysid, 

Americamysis 

bahia

90% 30-31a 72 hr LC50 ~66 Lussier et al. 

2000 

Mysid, 

Americamysis 

bahia

90% 30-31a 120 hr LC50 ~60 Lussier et al. 

2000 

Mysid, 

Americamysis 

bahia

90% 30-31a 144 hr LC50 ~60 Lussier et al. 

2000 

Mysid, 

Americamysis 

bahia

90% 30-31a 168 hr LC50 ~60 Lussier et al. 

2000 

Mysid, 

Americamysis 

bahia

>95% 20a 24 hr LC50 >47 Ward and Boeri 

1990a 
64 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Mysid, 

Americamysis 

bahia

>95% 20a 48 hr LC50 >47 Ward and Boeri 

1990a 

Mysid, 

Americamysis 

bahia

>95% 20a 72 hr LC50 44 Ward and Boeri 

1990a 

Pacific Oyster 

(embryo-larva), 

Crassostrea gigas

- 35a 72 hr Delayed 

development to 

D-shape stage 

100 Nice et al. 2000 

Copepod 

(10-12 d), 

Acartia tonsa

- 18a 48 hr LC50 synthetic 

media 

360 

280 

Kusk and 

Wollenberger 

1999 

Amphipod, 

Leptocheirus 

plumulosus

90% 30-31a 48 hr LC50 ~160 Lussier et al. 

2000 

Amphipod, 

Leptocheirus 

plumulosus

90% 30-31a 72 hr LC50 ~80 Lussier et al. 

2000 

Amphipod, 

Leptocheirus 

plumulosus

90% 30-31a 120 hr LC50 ~50 Lussier et al. 

2000 

Amphipod, 

Leptocheirus 

plumulosus

90% 30-31a 144 hr LC50 ~40 Lussier et al. 

2000 

Amphipod, 

Leptocheirus 

plumulosus

90% 30-31a 168 hr LC50 ~30 Lussier et al. 

2000 

Grass shrimp, 

Palaemonetes 

vulgaris

90% 30-31a 24 hr LC50 ~125 Lussier et al. 

2000 

Grass shrimp, 

Palaemonetes 

vulgaris

90% 30-31a 48 hr LC50 ~60 Lussier et al. 

2000 

Grass shrimp, 

Palaemonetes 

vulgaris

90% 30-31a 72 hr LC50 ~60 Lussier et al. 

2000 

Grass shrimp, 

Palaemonetes 

vulgaris

90% 30-31a 120 hr LC50 ~60 Lussier et al. 

2000 
65 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Shrimp, 

Crangon 

septemspinosa

>95% - 96 hr LC50 300 McLeese et al. 

1980b 

Shrimp, 

Crangon 

septemspinosa

>95% - 96 hr LC50 300 McLeese et al. 

1980b 

Shrimp, 

Crangon 

septemspinosa

>95% - 96 hr LC50 300 McLeese et al. 

1980b 

American lobster, 

Homarus 

americanus

90% 30-31a 24 hr LC50 ~140 Lussier et al. 

2000 

American lobster, 

Homarus 

americanus

90% 30-31a 48 hr LC50 ~140 Lussier et al. 

2000 

American lobster, 

Homarus 

americanus

90% 30-31a 72 hr LC50 ~140 Lussier et al. 

2000 

American lobster, 

Homarus 

americanus

>95% - 96 hr LC50 170 McLeese et al. 

1980b 

Atlantic salmon, 

Salmo salar

- - 96 hr LC50 190 McLeese et al. 

1980b 

Atlantic salmon, 

Salmo salar

- - 96 hr LC50 160 McLeese et al. 

1980b 

Atlantic salmon, 

Salmo salar

- - 96 hr LC50 130 McLeese et al. 

1980b 

Atlantic salmon, 

Salmo salar

- - 96 hr LC50 900 McLeese et al. 

1980b 

Sheepshead 

minnow, 

Cyprinodon 

variegatu

90% 30-31a 72 hr LC50 ~150 Lussier et al. 

2000 

Sheepshead 

minnow, 

Cyprinodon 

variegatu

90% 30-31a 120 hr LC50 ~125 Lussier et al. 

2000 
66 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Sheepshead 

minnow, 

Cyprinodon 

variegatu

90% 30-31a 144 hr LC50 ~120 Lussier et al. 

2000 

Sheepshead 

minnow, 

Cyprinodon 

variegatu

90% 30-31a 168 hr LC50 ~120 Lussier et al. 

2000 

Sheepshead 

minnow, 

Cyprinodon 

variegatu

>95% 15-17a 24 hr LC50 >420 Ward and Boeri 

1990c 

Sheepshead 

minnow, 

Cyprinodon 

variegatu

>95% 15-17a 48 hr LC50 340 Ward and Boeri 

1990c 

Sheepshead 

minnow, 

Cyprinodon 

variegatu

>95% 15-17a 72 hr LC50 320 Ward and Boeri 

1990c 

Killifish 

(embryo), 

Fundulus 

heteroclitus

85 - 90% 

(technical) 

20a 10 days 100% abnormal 

development 

2,204 Kelly and Di 

Giulio 2000 

Killifish 

(embryo), 

Fundulus 

heteroclitus

85 - 90% 

(technical) 

20a 96 hr LC50 5,444 Kelly and Di 

Giulio 2000 

Killifish 

(1-day old larva), 

Fundulus 

heteroclitus

85 - 90% 

(technical) 

20a 96 hr LC50 

(fed) 

214 Kelly and Di 

Giulio 2000 

Killifish 

(14-day old larva), 

Fundulus 

heteroclitus

85 - 90% 

(technical) 

20a 96 hr LC50 

(fed) 

209 Kelly and Di 

Giulio 2000 

Killifish 

(28-day old larva), 

Fundulus 

heteroclitus

85 - 90% 

(technical) 

20a 96 hr LC50 

(fed) 

260 Kelly and Di 

Giulio 2000 
67 

Table 6. Other Data on Effects of Nonylphenol on Aquatic Organisms 

Species Chemical pH Duration Effect

Concentration 

(μg/L) Reference

Threespine 

stickleback 

Gasterosteus 

aculeatus

Commercial 

(parasubstituted 

with branched 

nonyl chain) 

32a 96 hr LC50 370 Granmo et al. 

1991a 

Inland silversides, 

Menidia beryllina

90% 30-31a 24 hr LC50 ~120 Lussier et al. 

2000 

Inland silversides, 

Menidia beryllina

90% 30-31a 48 hr LC50 ~100 Lussier et al. 

2000 

Inland silversides, 

Menidia beryllina

90% 30-31a 72 hr LC50 ~80 Lussier et al. 

2000 

Inland silversides, 

Menidia beryllina

90% 30-31a 120 hr LC50 ~60 Lussier et al. 

2000 

Inland silversides, 

Menidia beryllina

90% 30-31a 144 hr LC50 ~60 Lussier et al. 

2000 

Inland silversides, 

Menidia beryllina

90% 30-31a 168 hr LC50 ~60 Lussier et al. 

2000 

a

Salinity (g/kg). 
68 

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