United States Environmental Offi ce 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 indepe ndent 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 scientif ically 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 Es trogenic 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 (C 15 

### H 24 

### O) is produced from cyclic intermediates in the refinement of petroleum and coal-tar crudes. It is manufactured by alkyl ating 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; a nd 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, br anched, 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 nonylphe nol 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 20 o 

### 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 (pK a 

### ) of 10.7 " 1.0 and a log octanol/water partition coefficient (log K ow 

### ) of 3.80 to 4.77 (Roy F. Weston Inc. 1990). The water solubility of nonylphenol is pH-dependent; 4,600 μ g/L at pH 5.0, 6,237 μ g/L at pH 7.0, 11,897 μ g/L at pH 9.0. Nonylphenol is soluble in seawater at 3,630 μ g/L and is soluble in many organic solvents (Roy F. Weston Inc. 1990). Ahel and Giger (1993) measured the solubility of nonylphenol at different temperatures in distilled water and demonstrated a nearly linear increase in solubility between 2 E C (4,600 μ g/L) and 25 E C (6,350 μ g/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 b een published (Bennie 1999). Shackelford et al. (1983) reported 4-nonylphenol at average concentrations ranging from 2 to 1,617 μ g/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 μ g/L. They found nonylphenol in all sediment samples with concentrations ranging
from 0.17 to 72 μ g/g (dry weight). Nonylphenol and its ethoxylates have been found in treatment plant wastewaters (Ellis et al. 1982, Gige r 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 μ g/L in the
runoff as a result of aircraft deicer and antii cer 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 a nd 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 μ g/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 μ g/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 diethoxyl ates 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 (anaerobi c) 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 re viewed by Maguire (1999). Gaffney (1976) determined that 1 mg/L nonylphenol did not degrade during 135-hr incubation with domestic wastewater. In industrial wastewater, nonylphe nol concentration was unchanged after 24 hr incubation, but decreased by 45 percent after 135 hr. Staples et al. (2001) determined that nonylphenol at 13 mg/L and 22 E C was mineralized to CO 2 

### within 35 days in aerobic systems inoculated with sludge from a waste treatment plant. No intermediate compounds were formed and the calculated half-life for nonylphenol was 8.2 days. Sundaram and Szeto (1981) studied nonylphenol fate incubated in open and closed containers of stream and pond waters. They found no degradation of nonylphenol incubated in open containers of the pond or stream waters. Th e observed half-life of 2.5 days, was attributed to volatilization. Incubation of nonylphenol in pond or stream waters in closed containers resulted in formation of a breakdown product. The observed half-life of nonylphenol in pond and stream water were estimated at 16.5 and 16.3 days, respectively. The same authors incubated nonylphenol in pond water with sediment present and found about 50 percent of the nonylphenol in the sediment after 10 days. About 80 percent of the nonylphenol in the sediment was degraded in 70 days. No degradation of nonylphenol occurred when autoclaved (sterilized) water and sediment samples were used. Staples et al. (1999) measured a half-life of 20 days for nonylphenol in water (31 mg/L) at 22 E C. They suggested that the temperature of water and the initial concentration of the nonylphenol both affect the degradation rate of the chemical. Ahel et al. (1994 a,b) studied the fate a nd transport of alkylphenol polyethoxylates (AP n EO)

### 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 μ g/L range. The
concentration of AP n EO degradates varied with time of day reflecting fluctuations in wastewater treatment plant discharge. Concentrations of AP n EO degradates also varied seasonally, being found at higher concentrations in the winter due to lower water temperature. Nonylphenol had less seasonal variability than other AP n EO 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 AP n EO 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 AP n EOs degrade through either the loss of ethylene oxide units to form low-molecu lar weight ethoxylates or through the formation of carboxylated ethoxylates ultimately terminating in CO 2 

### and water. Nonylphenol is formed during anaerobic breakdown of the AP n EOs and is therefore a minor component of wastewater
treatment effluents because of aerobic conditi ons 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 responsib le 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. Di ssipation from the sediment was estimated to be 50 and 95 percent at 66 and 401 days, respectiv ely. Hale et al. (2000) measured nonylphenol
concentrations in sediments below wastewater outfalls and found one site that had a sediment concentration of 54,400 μ g/kg more than twenty years after the treatment plant ceased operation. Bennett and Metcalfe (1998; 2000) found that nonyl phenol was widely distributed in lower

### 5 Great Lakes sediments and reached 37,000 μ g/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 AP n EOs in marine sediments and found that degradation occurred in oxic and anoxic conditions. They reported no clear evidence for net formation of nonylphenol from AP n EOs 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 K ow 

### 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-de rived 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 al gae). 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 μ g/kg
and a mean value of 12.0 μ g/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 tw o 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 μ g/kg). Neither nonylphenol or its ethoxylates were detected in the sediment s 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 f ound) at concentrations that ranged from 2 -
15.3 μ g/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 su ch 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 chr onic 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 pr ogram, 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 eff ects 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 devel oped, 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 Thei r 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 Offi ce 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 us ing 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 μ g/L for an amphipod ( Hyalella azteca ) to 774 μ g/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 CaCO 3 

### , respectively. An LC50 of 20.7 μ g/L was calculated in the lower hardness water and 150 μ g/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 μ g/L from a test that had the solutions renewed daily and Comber et al. (1993) reported an EC50 of 190 μ g/L in a static test. The Daphnia magna SMAV is 140.6 μ g/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 μ g/L, nymphs of the dragonfly Ophiogomphus sp. had an LC50 of 596 μ g/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 μ g/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 μ g/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 μ g/L for the fountain darter ( Etheostoma rubrum ) to 289.3 μ g/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 μ g/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 th reatened or endangered fish species. Acute toxicity test results were based on static tests with unmeasured nonylphenol concentrations and the LC50s ranged from 110 μ g/L for the fountain darter, Etheostoma rubrum , to a geometric
mean of 289.3 μ g/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 μ g/L (Brooke 1993a) to 360

μ g/L (Dwyer et al. 1995). The tests conduc ted 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 μ g/L). One test was available for the bluegill ( Lepomis macrochirus ) and the LC50 was 209 μ g/L (Brooke 1993a). Freshwater Species Mean Acute Values (SM AV) 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 anot her. 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 μ g/L (calculated using the procedure described in the “Guidelines”). This FAV is essentially the same as the lowest freshwater SMAV of 55.72

μ g/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 μ g/L for the winter flounder, Pleuronectes americanus, to 209.8 μ g/L for the sheepshead minnow, Cyprinodon variegatus (Lussier et al. 2000; Ward and Boeri 1990b), a di fference 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 test s 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 μ g/L; Lussier et al. 2000) compared to when they were not fed (43

μ g/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 μ g/L; Lussier et al. 2000) was lower than when the organisms were not fed (310 μ g/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) te sts 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 μ g/L; Lussier et al. 2000), the copepod,
Acartia tonsa (LC50 = 190 μ g/L; Kusk and Wollenberger 1999), American lobster, Homarus
americanus (LC50 = 71 μ g/L; Lussier et al. 2000), mud crab, Dyspanopeus sayii (LC50 >195

μ g/L; Lussier et al. 2000) and the amphipods, Leptocheirus plumulosus (LC50 = 61.6; Lussier et al. 2000) and Eohaustorius estuarius (LC50 = 138 μ g/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 μ g/L as the average of three tests (299, 194, 189 μ g/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 μ g/L. The sensitivity of the saltwater fish inland silversides ( Menidia

### 12 beryllina ), was intermediate (LC50 = 70 μ g/L) among the three saltwater fish species tested. Saltwater Species Mean Acute Values (SM AV) 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). Base d 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 μ g/L (calculated using the procedure described in the “Guidelines”). This FAV is lower than the lowest SMAV of 17 μ g/L for the the winter flounder,
Pleuronectes americanus

.

### 13 3. CHRONIC TOXICITY TO AQUATIC ANIMALS

3.1. Freshwater

The chronic toxicity of nonylphenol was determin ed 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 μ g/L, but not at
88.7 μ g/L, and survival was reduced at 377 μ g/L, but not at 202 μ g/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 μ g/L was calculated. However, the animals had received food and according to the Guidelines acute tests with this species must not receive food during an acute toxicity test if the test is to be used to compute an Acute-Chronic Ratio (ACR). Fliedner (1993) exposed 4 to 24 hr-old Daphnia magna neonates to nonylphenol for 22 days
in a 20 E C life-cycle test. Test solutions were renewed three times each week during which a 52.2 to 65.5 % decrease in nonylphenol concentration was measured. Mean measured nonylphenol test concentrations were: 0, 0, 1.55, 1.34, 3.45, 10.70, and 47.81 μ g/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 concen tratoins. 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 μ g/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 μ g/L, but not at 116 μ g/L. Survival was reduced to 60 percent at 215 μ g/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 μ g/L) and upper (215 μ g/L) chronic limits, for this test was 157.9 μ g/L. Dividing the acute value (104 μ g/L), determined from a companion test for this species (Brooke 1993a; Table 1) by the chronic value (157.9 μ g/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 surviv al, reproduction or growth at concentrations

# 24 μ g/L. The number of live young produced was significantly reduced at concentrations > 39

μ g/L when compared to controls. Growth was reduced at concentrations > 71 μ g/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 μ g/L) and upper (39 μ g/L) chronic limits, for this test was 30.59 μ g/L. Dividing the acute value (190 μ g/L), determined in in a companion test for this species (Table 1) by the chronic value (30.59 μ g/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 μ g/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 duri ng 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

μ g/L, respectively. The chronic value, calculated as the geometric mean of the NOEC and the LOEC, is 61.82 μ g/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 w ith embryos and fry of the rainbow trout, Oncorhynchus mykiss (Brooke 1993a). Five nonylphenol exposure concentrations were tested, ranging from 6.0 to 114 μ g/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 μ g/L). At the end of

### 15 the test, survival was significantly reduced at concentrations > 23.1 μ g/L but not at 10.3 μ g/L. Growth (both weight and length) was a more se nsitive chronic endpoint than survival. At the end of the test, the fish were significantly s horter (14 %) and weighed less (30 %, dry weight) than control fish at nonylphenol concentrations > 10.3 μ g/L, but not at 6.0 μ g/L. Based on growth, the NOEC and LOEC determined in this study were 6.0 and > 10.3 μ g/L, respectively. The chronic value, calculated as the geometric mean of the NOEC and the LOEC, is 7.861 μ g/L for rainbow trout. Dividing the acute value (221 μ g/L), determined in in a companion test for this species (Table 1) by the chronic value (7.861 μ g/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 μ g/L. Embryos in the control group and those in the three lowest nonylphenol exposure concentrations (2.8, 4.5, and 7.4 μ g/L) began to hatch on the third day of exposure, while the two higher concentrations (14 and 23 μ g/L) began hatching on the fourth da y. Growth (length or weight) of nonylphenol exposed fish was not significantly differe nt from the control organisms at any of the nonylphenol treatment concentrations. Survival of th e fish at the end of the test was significantly reduced at nonylphenol concentrations > 14 μ g/L. Fish survival averaged 56.7 % at 23 μ g/L nonylphenol, 66.7 % at 14 μ g/L nonylphenol, and 76.7 % at 7.4 μ g/L nonylphenol. At
concentrations < 7.4 μ g/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 μ g/L and 14 μ g/L, respectively. The chronic value, calculated as the geometric mean of the NOEC and the LOEC, is 10.18 μ g/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 μ g/L, but there was a 18 % reduction in survival and a 53% reduction in reproduction at 9.1 μ g/L. Effects on survival at the highest concentration tested (21 μ g/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 μ g/L nonylphenol, but no difference in growth for mysids exposed to 3.9 μ g/L nonylphenol. Based on growth, the NOEC and LOEC determined in this study were 3.9 μ g/L and 6.7 μ g/L, respectively. The chronic value, calculated as the geometric mean of the NOEC and the LOEC, is 5.112 μ g/L for mysids (Table 2). Dividing the acute value (43 μ g/L), determined in in a companion test for this species (Table 1) by the chronic value (5.112 μ g/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 μ g/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 μ g/L, but was
not affected at 9.46 μ g/L. Based on reproduction, the NOEC and LOEC determined in this study
were 9.46 μ g/L and 15.28 μ g/L, respectively. The chronic value, calculated as the geometric mean of the NOEC and the LOEC, is 12.02 μ g/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). Tw o 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 demonstra ting 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 gene ral 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 μ g/L) is closest to both the freshwater FAV (55.49

μ g/L) and the saltwater FAV (13.93 μ g/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 accepta ble 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 μ g/L based on cell counts. At the end of the toxicity test, algae from the highest exposure concentration (720 μ g/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 μ g/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 con centration of nonylphenol with survivors (120 μ g/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 μ g/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 nonylphe nol 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 μ g/L nonylphenol, respectively. Brooke (1993b) exposed juvenile fathead minnow ( Pimephales promelas ) and juvenile bluegill ( Lepomis macrochirus ) to nonylphenol at five concentrations for four and tw enty-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 con centrations 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 sublet hal concentrations of nonylphenol ranging from
0.4 to 3.4 μ g/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 14 

### C-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 e ffects 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 μ g/L) and Chlamydomonas reinhardtii (6,250 μ g/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 μ g/L (Weinberger and Iyengar 1983; Prasad 1986). Protozoa were affected in the concentration range from 50 to 747 μ g/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 μ g/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

μ g/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-hardne ss 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 μ g/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 μ g/L, which agreed with LC50s for other cladoceran species. The cladoceran, Ceriodaphnia dubia, gave similar LC50
results of 276 and 225 μ g/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 μ g/L (Table 2). England and Bussard (1993) reported an EC50 and an LC50 for larva of the midge, Chironomus tentans , of 95 and 119 μ g/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 orga nisms were fed, Brooke (1993b) measured a 96-hr LC50 for the fathead minnow, Pimephales promelas , of 138 μ g/L and a 96-hr LC50 for the bluegill, Lepomis macrochirus , of 135 μ g/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 μ g/L for the Atlantic salmon, Salmo salar , in a 96-hr exposure and Lech et al (1996) re ported 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 μ g/L for brook trout juveniles ( Salvelinus fontinalis ), a 96-hr LC50 of 230 μ g/L for rainbow trout juveniles ( Oncorhynchus mykiss ) and a 32-day LC50 of > 40 μ g/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 μ g/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 μ g/L of nonylphenol (Ashfield et al. 1998). They found reduced growth in fish exposed to 1 μ g/L for 22 days and grown for 86 days beyond treatment. Growth was not reduced in the 10 μ g/L treatment but was in the 50 μ g/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 μ g/L treatments on

### 22 day 55 of the study, but not at the 1 μ g/L. At day 466, the fish exposed to 10 μ g/L recovered the growth reductions seen earlier and only the 30 μ g/L exposed fish showed reduced (approximately 25%) growth. Five fish species (rainbow trout, Lahontan cu tthroat 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 μ g/L were observed in the rainbow trout, Lahontan cutthroat trout and Apache trout. The lack of a clear
connection between this sublethal biochemi cal endpoint and population relevant effectsprecludes 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 reporte d 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 μ g/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 al so 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 observe d 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 μ g/L

### 23 nonylphenol but were negatively affected at 76 μ g/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 μ g/L. The fish (bluegill) in the mesocosms (Liber et al. 1999) were unaffected at nonylphenol exposures < 76 μ g/L, but survival was reduced at 243 μ g/L. In one exposure replicate with a mean nonlyphenol concentration of 93 μ g/L, survival of the fish was reduced after 20 days of exposure indicating that concentrations near 100 μ g/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 μ g/L. The zooplankton in the study were not affected at concentrations ranging from
19 to 44 μ g/L (mean of 30 μ g/L), but species richness was affected at concentrations >30 μ g/L.
Effects observed in these mesocosm studies were all above the freshwater Final Chronic Value
of 5.920 μ g/L.

6.2. Saltwater

Additional data on the lethal and sublethal e ffects 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 μ g/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 μ g/L (Billinghurst et al. 1998). The soft-shell clam, Mya arenaria , showed no adverse effects on survival from a 360-hr exposure at 700 μ g/L (McLeese et al. 1980b). Granmo et al. (1989) report LC50s of 3,000 μ g/L
and 500 μ g/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 μ g/L (Granmo et al. 1989) and caused effects on attachment activity at 22 μ g/L (Etoh et al. 1997). Lussier et al. (2000) tested a number of saltw ater 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 μ g/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 μ g/L for the copepod, Acartia tonsa , exposed to nonylphenol in a synthetic media. Nice et al (2000) reported developmental effects at 100 μ g/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 μ g/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 μ g/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 μ g/L exposure group. The authors calculated a zero population growth value ( λ ) of 19 μ g/L for the 55-day multigenerational test. The chronic values from the 28-day exposure in this study was 12.02 μ g/L (Table 2) and from a similar 28-day study by Ward and Boeri (1991b) was 5.112 μ g/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 μ g/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 μ g/L for a 24-hr exposure to 320 μ g/L for a 72-hr exposure. Threespine stickleback, Gasterosteus aculeatus , exposed to a commercial mixture of nonylphenol had a 96-hr LC50 of 370 μ g/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 μ g/L for 96 hr were all abnormally developed at 10 days post-fertilization. The LC50 for the same exposure period was 5,444 μ g/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 Es trogenic 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 endpoi nts 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 μ g/L, but estrogenic effects were reported to have occurred in the amphipod Corophium volutator (Brown et al. 1999) at 10 μ g/L
and in the larva of Chironomus riparius (Hahn et al. 2002) at 2,000 μ g/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, 17 β -estradiol. Ren et al. (1996a) demonstrated significant increases in vitellogenin production in rainbow trout exposed to nonylphenol at 100 μ g/L for 72 hr. In another study, Ren et al. (1996b) demonstrated that nonylphenol could stimulate the production of vitellogenin mRNA (which precedes vitelloge nin protein synthesis) within 4 hr at 10 μ g/L. Similarly, Lech et al. (1996) observed a signifi cant increase in vitellogenin mRNA at 72 hr in
rainbow trout at 14.14 μ g/L nonylphenol. Vitellogenin was induced in green swordfish, Xiphophorus helleri , by exposure to 4 μ g/L of technical grade nonylphe nol (Kwak et al. 2001). Jobling et al. (1996) demonstrated significant in creases in vitellogenin in male rainbow trout,

### 27 Oncorhynchus mykiss , at three weeks of exposure to 20.3 and 54.3 μ g/L of nonylphenol. Similar results were reported in another study with rainbow trout, plasma vitellogenin was increased after 21 days exposure to 50 μ g/L of nonylphenol (Tremblay and Van Der Kraak 1998). Harris et al. (2001) also observed increased plasma vite llogenin levels in female rainbow trout exposed
to 8.3 and 85.6 μ g/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 μ g/L of nonylphenol (Pedersen et al. 1999) or in Atlantic salmon, Salmo trutta , exposed for 30 days to 20 μ g/L (Moore et al. 2003). The influence of exposure route on nonylphenol-induced vitellogenin m RNA 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 μ g/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 th e reproductive season, concentrations of > 0.3 to
0.4 μ g/L did appear to reduce fecundity. However, fish exposed to 0.09 and 0.1 μ g/L produced more eggs than control fish. These data appear to produce a U-shaped dose-response and indicate a possible hormetic response of fecund ity to nonylphenol. Nonylphenol concentrations
of 0.05 to 3.4 μ g/L did not significantly change vitellogenin concentrations in the blood of males, and raised the 17 β -estradiol titers in the blood of male and female fish at most treatment
concentrations > 0.05 μ g/L. An increase in the number of Sertoli cells may have occurred in the male fathead minnow exposed to nonylphenol at 1.6 μ g/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 μ g/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 μ g/L of nonylphenol (Gray and Metcalf 1997). The sex ratio shifted in favor of females at the highest exposure con centration. 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 concentra tion for vitellogenin induction was 11.6 μ g/L. Yokota et al. (2001) conducted a two-genera tion 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 (F 1 

### ) generation. Five exposure concentrations of nonylphenol in quadruplicate (4.2, 8.2, 17.7, 51.5,
and 183 μ g/L) and water-only and solvent controls were used. In the F 0 

### generation, egg hatchability was reduced (46.7%) by 183 μ g/L nonylphenol exposure, survival was significantly decreased at 60 days post-hatch by nonylphenol exposures > 17.7 μ g/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 μ g/L treatment, with 20% of fish displaying male characteristics externally having ovo-testis tissues. In fish from the 51.5 μ g/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 μ g/L nonylphenol, but was not observed in the fish with ovotestis exposed to 51.5 μ g/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 μ g/L, but not significantly, and GSI of fe male fish was increased significantly by exposure to nonylphenol concentrations > 8.2 μ g/L. The effects of nonylphenol on F 1 

### 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 μ g/L to favor females 1:2. Induction of ovo-testis was observed at lower concentrations of nonylphenol in the F 1 

### generation than in the F 0 

### generation. Ovo-testis were
observed in the 8.2 μ g/L exposure group (10%) and in the 17.7 μ g/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 F 0 

### generation in the
17.7 μ g/L treatment. The overall results indicate a LOEC of 17.7 μ g/L and a NOEC of 8.2 μ g/L

### 29 with a chronic value, calculated as the geometric mean of the NOEC and LOEC, of 12.05 μ g/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 μ g/L over a 4 month period. Mortality rate in the progeny was significantly reduced by parental exposure to both 1 and 10 μ g/L nonylphenol and hatching in the progency was reduced by parental exposure to 10 μ g/L nonylphenol. Vitellogenin was induced (approximately 10-fold) in adult male fish exposed to both 1 and 10 μ g/L nonylphenol. In the male progeny of parental fish exposed to 10 μ g/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 pl asma estradiol concentrations were not different from control levels. Testicular tissue of nonylphe nol exposed adult male fish was not affected by nonylphenol. Sex ratios of the offspring of e xposed 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 belo w the freshwater Final Chronic Value of 6.5965

μ g/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 μ g/L for the amphipod Hyalella azteca to 774 μ g/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 μ g/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 characteristic s such as hardness and pH. The freshwater
Final Acute Value is 55.49 μ g/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 μ g/L based on growth, and the fathead minnow, Pimephales promelas , had a chronic value of 10.18 μ g/L based on survival. Two species of freshwater cladocerans were tested and chronic values ranged from 22.62 to 157.9 μ g/L based on reproduction. One species of freshwater midge was tested and its chronic value was 61.82 μ g/L based on survival.
Data were available to calculate a Final Ac ute-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 μ g/L for the winter flounder, Pleuronectes americanus , to 209.8 μ g/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 μ g/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 μ g/L based on reduced growth and 12.02 μ g/L based on a reproductive endpoint.
Data were available to calculate a Final Ac ute-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 a quatic plants, one freshwater alga and one saltwater diatom, were available. Algae were as sensitive as animals, showing effect concentrations that ranged from 27 μ g/L for the freshwater alga to 410 μ g/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 μ g/L.

8.4. Bioaccumulation Data

Nonylphenol bioaccumulation in aquatic organisms is less than would be predicted from the log K ow 

### 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 μ g/L more than once every three years on the average and if the four-day average concentration of nonylphenol does not exceed 6.6 μ g/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 μ g/L more than once every three years on the average and if the four-day average concentration of nonylphenol does not exceed 1.7 μ g/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-s pecific 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 th e 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 considera tions 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 1000Nonylphenol 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 1000Nonylphenol 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 1000Nonylphenol 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 Method a Chemical pH LC 50 or EC 50 ( : g/L) Species Mean Acute Value b ( : 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 Method a Chemical pH LC 50 or EC 50 ( : g/L) Species Mean Acute Value b ( : 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 Method a Chemical pH LC 50 or EC 50 ( : g/L) Species Mean Acute Value b ( : 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 Method a Chemical pH LC 50 or EC 50 ( : g/L) Species Mean Acute Value b ( : 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 Method a Chemical pH LC 50 or EC 50 ( : g/L) Species Mean Acute Value b ( : 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 fro m the underlined number(s) in the preceding column.

### 44 Table 2. Chronic Toxicity of Nonylphenol to Aquatic Animals

Species Test a Chemical pH Chronic Limits ( μ g/L) b Chronic Value ( μ g/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 lif e-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 ( μ g/L) Chronic Value ( μ g/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

Rank a Genus Mean Acute Value

( μ g/L) Species Species Mean Acute Value

( μ g/L) b Species Mean Acute-Chronic Ratio c

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

Rank a Genus Mean Acute Value

( μ g/L) Species Species Mean Acute Value

( μ g/L) b Species Mean Acute-Chronic Ratio c 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 μ g/L
Criterion Maximum Concentration = 55.49 ÷ 2 = 27.75 μ g/L Final Acute-Chronic Ratio = 8.412 (see text)
Final Chronic Value = 55.49 μ g/L ÷ 8.412 = 6.5965 μ g/L

Saltwater
Final Acute Value = 13.93 μ g/L
Criterion Maximum Concentration = 13.93 ÷ 2 = 6.965 μ g/L Final Acute-Chronic Ratio = 8.412 (see text)
Final Chronic Value = 13.93 μ g/L ÷ 8.412 = 1.6560 μ g/L

### 49 Table 4. Toxicity of Nonylphenol to Aquatic Plants

Species Chemical pH Duration (days) Effect Concentration ( μ g/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% 30 a 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. ( μ g/L) a pH Duration (days) Tissue Percent Lipids BCF

or BAF b Normalized BCF or BAF c 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. ( μ g/L) a pH Duration (days) Tissue Percent Lipids BCF

or BAF b Normalized BCF or BAF c 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. ( μ g/L) a pH Duration (days) Tissue Percent Lipids BCF

or BAF b Normalized BCF or BAF c 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 14 Clabeled 5.9 - 16 Whole body 1.6 2,740 1,712 Ekelund et al. 1990

Blue mussel, Mytilus edulis 14 Clabeled 6.2 - 16 Whole body 1.9 4,120 2,168 Ekelund et al. 1990

Common shrimp, Crangon crangon d 14 Clabeled 6.4 - 16 Whole body 1.4 110 78.75 Ekelund et al. 1990

Common shrimp, Crangon crangon d 14 Clabeled 7.4 - 16 Whole body 1.7 900 529.4 Ekelund et al. 1990

Three-spined stickleback, Gasterosteus aculeatus 14 Clabeled 4.8 - 16 Whole body 6.7 1,200 179.1 Ekelund et al. 1990

Three-spined stickleback, Gasterosteus aculeatus 14 Clabeled 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. ( μ g/L) a pH Duration (days) Tissue Percent Lipids BCF

or BAF b Normalized BCF or BAF c Reference

a Measured concentration of nonylphenol. b Bioconcentration factors (BCFs) and bioaccumulation factors (BAFs) are ba sed on measured concentrations of nonylphenol in water and in tissue.

c When possible, the factors were normali zed 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 ( μ g/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 macroinvertebrates 96.4% 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 ( μ g/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 ( μ g/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 ( μ g/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 ( μ g/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 ( μ g/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; F 0 3 yr-old adults), Oncorhynchus mykiss 98% 7.6 4 months (exposed 10 days/month) Reduced embryo survival; Reduced hatch;
F o Males increased vitellogenin; F 1 Females increased vitellogenin and

testosterone;

F 1 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 ( μ g/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 ( μ g/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 ♂ & ♀ 17 β -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 ( μ g/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% 30-31 a 24 hr LC50 50 Lussier et al. 2000 Coot clam, Mulinia lateralis 90% 30-31 a 48 hr LC50 50 Lussier et al. 2000 Coot clam, Mulinia lateralis 90% 30-31 a 72 hr LC50 40 Lussier et al. 2000 Blue mussel, Mytilus edulis 90% 30-31 a 96 hr LC50 3,000 Granmo et al. 1989 Blue mussel, Mytilus edulis - 32 a 360 hr LC50 500 Granmo et al. 1989 Blue mussel, Mytilus edulis - 32 a 13 days Reduced byssus strength 56 Granmo et al. 1989 Blue mussel, Mytilus edulis - 32 a 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 ( μ g/L) Reference Blue mussel, Mytilus edulis - 32 a 30 days No byssus threads formed 100 Granmo et al. 1989 Blue mussel, Mytilus edulis - 32 a 32 days Reduction in growth 56 Granmo et al. 1989 Blue mussel, Mytilus edulis - 32 a 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% 30-31 a 24 hr LC50 ~114 Lussier et al. 2000

Mysid, Americamysis bahia 90% 30-31 a 48 hr LC50 ~82 Lussier et al. 2000

Mysid, Americamysis bahia 90% 30-31 a 72 hr LC50 ~66 Lussier et al. 2000

Mysid, Americamysis bahia 90% 30-31 a 120 hr LC50 ~60 Lussier et al. 2000

Mysid, Americamysis bahia 90% 30-31 a 144 hr LC50 ~60 Lussier et al. 2000

Mysid, Americamysis bahia 90% 30-31 a 168 hr LC50 ~60 Lussier et al. 2000

Mysid, Americamysis bahia >95% 20 a 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 ( μ g/L) Reference Mysid, Americamysis bahia >95% 20 a 48 hr LC50 >47 Ward and Boeri 1990a

Mysid, Americamysis bahia >95% 20 a 72 hr LC50 44 Ward and Boeri 1990a

Pacific Oyster (embryo-larva), Crassostrea gigas - 35 a 72 hr Delayed development to D-shape stage 100 Nice et al. 2000

Copepod (10-12 d), Acartia tonsa - 18 a 48 hr LC50 synthetic media 360 280 Kusk and Wollenberger 1999 Amphipod, Leptocheirus plumulosus 90% 30-31 a 48 hr LC50 ~160 Lussier et al. 2000

Amphipod, Leptocheirus plumulosus 90% 30-31 a 72 hr LC50 ~80 Lussier et al. 2000

Amphipod, Leptocheirus plumulosus 90% 30-31 a 120 hr LC50 ~50 Lussier et al. 2000

Amphipod, Leptocheirus plumulosus 90% 30-31 a 144 hr LC50 ~40 Lussier et al. 2000

Amphipod, Leptocheirus plumulosus 90% 30-31 a 168 hr LC50 ~30 Lussier et al. 2000

Grass shrimp, Palaemonetes vulgaris 90% 30-31 a 24 hr LC50 ~125 Lussier et al. 2000

Grass shrimp, Palaemonetes vulgaris 90% 30-31 a 48 hr LC50 ~60 Lussier et al. 2000

Grass shrimp, Palaemonetes vulgaris 90% 30-31 a 72 hr LC50 ~60 Lussier et al. 2000

Grass shrimp, Palaemonetes vulgaris 90% 30-31 a 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 ( μ g/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% 30-31 a 24 hr LC50 ~140 Lussier et al. 2000

American lobster, Homarus americanus 90% 30-31 a 48 hr LC50 ~140 Lussier et al. 2000

American lobster, Homarus americanus 90% 30-31 a 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% 30-31 a 72 hr LC50 ~150 Lussier et al. 2000

Sheepshead minnow, Cyprinodon variegatu 90% 30-31 a 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 ( μ g/L) Reference Sheepshead minnow, Cyprinodon variegatu 90% 30-31 a 144 hr LC50 ~120 Lussier et al. 2000

Sheepshead minnow, Cyprinodon variegatu 90% 30-31 a 168 hr LC50 ~120 Lussier et al. 2000

Sheepshead minnow, Cyprinodon variegatu >95% 15-17 a 24 hr LC50 >420 Ward and Boeri 1990c

Sheepshead minnow, Cyprinodon variegatu >95% 15-17 a 48 hr LC50 340 Ward and Boeri 1990c

Sheepshead minnow, Cyprinodon variegatu >95% 15-17 a 72 hr LC50 320 Ward and Boeri 1990c

Killifish (embryo), Fundulus heteroclitus 85 - 90% (technical) 20 a 10 days 100% abnormal development 2,204 Kelly and Di Giulio 2000

Killifish (embryo), Fundulus heteroclitus 85 - 90% (technical) 20 a 96 hr LC50 5,444 Kelly and Di Giulio 2000

Killifish (1-day old larva), Fundulus heteroclitus 85 - 90% (technical) 20 a 96 hr LC50 (fed) 214 Kelly and Di Giulio 2000

Killifish (14-day old larva), Fundulus heteroclitus 85 - 90% (technical) 20 a 96 hr LC50 (fed) 209 Kelly and Di Giulio 2000

Killifish (28-day old larva), Fundulus heteroclitus 85 - 90% (technical) 20 a 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 ( μ g/L) Reference Threespine stickleback Gasterosteus aculeatus Commercial(parasubstituted with branched nonyl chain) 32 a 96 hr LC50 370 Granmo et al. 1991a

Inland silversides, Menidia beryllina 90% 30-31 a 24 hr LC50 ~120 Lussier et al. 2000 Inland silversides, Menidia beryllina 90% 30-31 a 48 hr LC50 ~100 Lussier et al. 2000 Inland silversides, Menidia beryllina 90% 30-31 a 72 hr LC50 ~80 Lussier et al. 2000 Inland silversides, Menidia beryllina 90% 30-31 a 120 hr LC50 ~60 Lussier et al. 2000 Inland silversides, Menidia beryllina 90% 30-31 a 144 hr LC50 ~60 Lussier et al. 2000 Inland silversides, Menidia beryllina 90% 30-31 a 168 hr LC50 ~60 Lussier et al. 2000

a Salinity (g/kg).

### 68 11. REFERENCES

Ahel, M. and W. Giger. 1993. Aqueous solubility of alkylphenols and alkylphenol polyethoxylates. Chemosphere
26:1461-1470.

Ahel, M., J. McEvoy and W. Giger. 1993. Bioaccumulati on of the lipophilic metabolites of nonionic surfactants in freshwater organisms. Environ. Pollut. 79:243-248.

Ahel, M., W. Giger and M. Koch. 1994a. Behaviour of alkylphenol polyethoxylate surfactants in the aquatic
environment

B I. Occurrence and transformation in se wage treatment. Wat. Res. 28:1131-1142.

Ahel, M., W. Giger and C. Schaffner. 1994b. Behavior of alkylphenol polyethoxylate surfactants in the aquatic
environment

B II. Occurrence and transformation in rivers. Wat. Res. 28:1143-1152.

Ahel, M., F.E. Scully, Jr., J. Hoigne and W. Giger. 1994c. Photochemical degradation of nonylphenol and nonylphenol polyethoxylates in natural waters. Chemosphere 28:1361-1368.

Ahel, M., C. Schaffner and W. Giger. 1996. Behaviour of alkylphenol polyethoxylates surfactants in the aquatic
environment

B III. Occurrence and elimination of their persistent metabolites during infiltration of river water to groundwater. Wat. Res. 30:37-46.

Ahel, M., E. Molnar, S. Ibric and W. Giger. 2000. Estrogenic metabolites of alkylphenol polyethoxylates in secondary sewage effluents and rivers. Wat. Sci. Tech. 42:15-22.

Allen, Y., A.P. Scott, P. Matthiessen, S. Haworth, J.E. Thain and S. Feist. 1999. Survey of estrogenic activity in United Kingdom estuarine and coastal waters and its effects on gonadal development of the flounder Platichthys flesus . Environ. Toxicol. Chem. 18:1791-1800.

Amato, J.R. and D.D. Wayment. 1998. Surfactant toxi city identification with a municipal wastewater. In: Environmental Toxicology and Risk Assessment. Little, E. S., A.J. DeLoney and B.M. Greenberg, Eds. ASTM STP 1333, Vol. 7, American Society for Testing a nd Materials, Philadelphia, PA. pp. 272-283.

Andersen, H.R., A. Andersson, S.F. Arnold, H. Autrup, M. Barfoed, N.A. Beresford, P. Bjerregaard, L.B. Christiansen, B. Gissel, R. Hummel, E.B. Jorgensen, B. Korsgaard, R. Le Guevel, H. Leffers, J. McLachlan, A. Moller, J.B. Nielsen, N. Olea, A. Oles-Karasko, F. Pa kdel, K.L. Pedersen, P. Pe rez, N.E. Skakkeboek, C. Sonnenschein, A.M. Soto, J.P. Sump ter, S.M. Thorpe and P. Grandjean . 1999. Comparison of short-term

### 69 estrogenicity tests for identification of hormone-disrupting chemicals. Environ. Health Perspect. 107(Suppl. 1):89-
108.

Applegate, V.C., J.H. Howell, A.E. Hall, Jr. and M.A. Smith . 1957. Toxicity of 4,346 chemicals to larval lampreys and fishes. Special Scientific Report-Fisheries, No. 207, U. S. Dept. Interior, Fish and Wildlife Service, Washington, DC. 157 pp.

Arukwe, A., L. Forlin and A. Goksoyr. 1997a. Xenobiotic and steroid biotransformation enzymes in Atlantic salmon ( Salmo salar ) liver treated with an estrogenic compound, 4-nonylphenol. Environ. Toxicol. Chem. 16:2576-
2583.

Arukwe, A., F.R. Knudsen and A. Goksoyr. 1997b. Fish z ona radiata (eggshell) protein: A sensitive biomarker for environmental estrogens. Environ. Health Perspect. 105:418-422.

Arukwe, A., T. Celius, B.T. Walther and A. Goksoyr. 1998. Plasma levels of vitellogenin and eggshell zona radiata proteins in 4-nonylphenol and o,p = nonylphenol treated juvenile Atlantic salmon ( Salmo salar ). Mar.
Environ. Res. 46:133-136.

Arukwe, A., R. Thibaut, K. Ingebrigtsen, T. Celius, A. Goksoyr and J-P. Cravedi. 2000. In vivo and in vitro metabolism and organ distribution of nonylphenol in Atlantic salmon ( Salmo salar ). Aquat. Toxicol. 49:289-304.

Ashfield, L.A., T.G. Pottinger and J.P. Sumpter. 1998. Exposure of female juvenile rainbow trout to alkylphenolic compounds results in modifications to growth and ovos omatic index. Environ. Toxicol. Chem. 17:679-686.

Baer, K.N. and K.D. Owens. 1999. Evaluation of select ed endocrine disrupting compounds on sex determination in Daphnia magna using reduced photoperiod and different feeding ra tes. Bull. Environ. Contam. Toxicol. 62:214-
221.

Baldwin, W.S., S.E. Graham, D. Shea and G.A. LeBl anc. 1997. Metabolic androgenization of female Daphnia

magna by the xenoestrogen 4-nonylphenol. E nviron. Toxicol. Chem. 16:1905-1911.

Baldwin, W.S., S.E. Graham, D. Shea and G.A. Leblanc. 1998. Altered metabolic elimination of testosterone and associated toxicity following exposure to Daphnia magna to nonylphenol polyethoxylate. Ecotoxicol. Environ. Saf.
39:104-111.

### 70 Bearden, A.P. and T.W. Schultz. 1997. Structure-activity relationships for Pimephales and Tetrahymena : a mechanism of action approach. E nviron. Toxicol. Chem. 16:1311-1317.

Bearden, A.P. and T.W. Schultz. 1998. Comparison of Tetrahymena and Pimephales toxicity based on mechanism of action. SAR QSAR Environ. Res. 9:127-153.

Bechmann, R.K. 1999. Effect of the endoc rine disrupter nonylphenol on the marine copepod Tisbe battagliai. Sci.
Total Environ. 233:33-46.

Bennett, E.R. and C.D. Metcalfe. 1998. Distribution of alkylphenol compounds in Great Lakes sediments, United
States and Canada. Environ. Toxicol. Chem. 17:1230-1235.

Bennet, E.R. and C.D. Metcalfe. 2000. Distribution of degradation products of alkylphenol ethoxylates near sewage treatment plants in the Lower Great Lakes, Nort h America.. Environ. T oxicol. Chem. 19: 784-792.

Bennie, D.T. 1999. Review of the environmental o ccurrence of alkylphenols and alkylphenol ethoxylates. Wat. Qual. Res. J. Can. 34:79-122.

Bennie, D.T., C.A. Sullivan, H. Lee, T.E. Peart a nd R.J. Maguire. 1997. Occurrence of alkylphenols and alkylphenol mono- and di-ethoxylates in natural waters of the Laurentian Great Lakes basin and the upper St.
Lawrence River. Sci. Total Environ. 193:263-275.

Bettinetti, R. and A. Provini. 2002. Toxicity of 4-nonylphenol to Tubifex tubifex and Chironomus riparius in 28- day whole-sediment tests. Ecot oxicol. Environ. Safety 53:113-121.

Billinghurst, Z., A.S. Clare, T. Fileman, M. McEvoy, J. Readman and M.H. Depledge. 1998. Inhibition of barnacle settlement by the environmental oestrogen 4-nonylphenol and the natural oestrogen 17 β oestradiol. Mar. Pollut.
Bull. 36:833-839.

Blackburn, M.A., S.J. Kirby and M.J. Waldock. 1999. Concentrations of alkyphenol polyethoxylates entering UK
estuaries. Mar. Pollut. Bull. 38:109-118.

Braaten, B., A. Granmo and R. Lange. 1972. Tissue-swelling in Mytilus edulis L. induced by exposure to a nonionic surface active agent. Norw. J. Zool. 20:137-140.

### 71 Brooke, L.T. 1993a. Acute and chronic toxicity of nonylphenol to ten species of aquatic organisms. Report to the U.S. EPA for Work Assignment No. 02 of Contract No. 68-C1 -0034. Lake Superior Research Institute, University of Wisconsin-Superior, Superior, WI. March 24. 30 pp. Amended 18 October 2005. 34 pp.

Brooke, L.T. 1993b. Nonylphenol Toxicity. Accumulation and lethality for two freshwater fishes (fathead minnow and bluegill) to nonylphenol. Report to the U.S. EPA for Work Assignment No. 1-12 of Contract No. 68-C1-0034. Lake Superior Research Institute, University of Wi sconsin-Superior, Superior, WI. September 30. 50 pp.

Brooke, L.T. 1994. Nonylphenol Analysis. Accumulation and lethality for two freshwater fishes (fathead minnow and bluegill) to nonylphenol. Report to the U.S. EPA for Work Assignment No. 2-15 of Contract No. 68-C1-0034. Lake Superior Research Institute, University of Wisconsin-Superior, Superior, WI. May 31. 49 pp.

Brown, R.J., M. Conradi and M.H. Depledge. 1999. Long-term exposure to 4-nonylphenol affects sexual differentiation and growth of the amphipod Corophium volutator (Pallas, 1766). Sci. Total Environ. 233:77-88.

Budavari, S. (Ed.). 1989. The Merck Index: An encycl opedia of chemicals, drugs, and biologicals. 11th ed. Merck and Co., Inc. Rahway, NJ.

Burkhardt-Holm, P., T. Wahli and W. Meier. 2000. Nonyl phenol affects the granulation of pattern of epidermal mucous cells in rainbow trout, Oncorhynchus mykiss . Ecotoxicol. Environ. Safe. 46:34-40.

Celius, T., T.B. Haugen, T. Grotmol and B.T. Walth er. 1999. A sensitive zonagenetic assay for rapid in vitro assessment of estrogenic potency of xenobiotics and my cotoxins. Environ. Health Perspect. 107:63-68.

Christiansen, T., B. Korsgaard and A. Jesp ersen. 1998a. Effects of nonylphenol and 17 β -estradiol on vitellogenin synthesis, testicular structure and cytology in male eelpout Zoarces viviparus . J. Exp. Biol. 201:179-192.

Christiansen, T., B. Korsgaard and A. Jespersen. 1998b. Induction of vitellogenin synthesis by nonylphenol and
17 β -estradiol and effects on the testicular structure in the eelpout Zoarces viviparus . Mar. Environ. Res. 46:141-
144.

Christiansen, T., K.L. Pedersen, B. Korsgaard and P. Bje rregaard. 1998c. Estrogenicity of xenobiotics in rainbow trout ( Oncorhynchus mykiss ) using in vivo synthesis of vitellogenin as a biomar ker. Mar. Environ. Res. 46(1-5):
137-140.

### 72 Christensen, L.J., B. Korsgaard and P. Bjerregaard. 1999. The effect of 4-nonylphenol on the synthesis of vitellogenin in the flounder Platichthys flesus . Aquat. Toxicol. 46:211-219.

Coldham, N.G., M.J. Sauer, S. Sivapathasundaram, L. Ashfield, T. Pottinger and C. Goodall. 1997. Tissue distribution, metabolism and excretion of 4-nonylphenol in ra inbow trout. J. Vet. Pharm acol. Therap. 20(Suppl. 1):
256-257.

Coldham, N.G., S. Sivapathasundaram, M. Dave, L.A. Ashf ield, T.G. Pottinger, C. Gooda ll and M.J. Sauer. 1998. Biotransformation, tissue distribution, and persisten ce of 4-nonylphenol residues in juvenile rainbow trout ( Oncorhynchus mykiss ). Drug Metabol. Dispos. 26:347-354.

Comber, M.H.I., T.D. Williams and K.M. St ewart. 1993. The effects of nonylphenol on Daphnia magna . Wat.
Res. 27:273-276.

Corsi, S.R., D.H. Zitomer, J.A. Field and D.A. Canc illa. 2003. Nonylphenol ethoxylates and other additives in aircraft deicers, antiicers, and waters receiving airport runoff. Environ. Sci. Technol. 37:4031-4037.

Dorn, P.B., J.P. Salanitro, S.H. Evans and L. Kravetz. 1993. Assessing the aquatic hazard of some branched and linear nonionic surfactants by biodegradation and t oxicity. Environ. Toxicol. Chem. 12:1751-1762.

Dwyer, F.J., L.C. Sappington, D.R. Buckler and S.B. J ones. 1995. Use of surrogate species in assessing contaminant risk to endangered and threatened speci es. EPA/600/R-96/029. National Technical Information Service, Springfield, VA. 71 pp.

Dwyer, F.J., D.K. Hardesty, C.E. Henke, C.G. Ingersoll, D.W. Whites, D.R. Mount and C.M. Bridges. 1999a. Assessing contaminant sensitivity of endandered and thr eatened species: Toxicant cl asses. EPA/600/R-99/098. National Technical Information Service, Springfield, VA. 15 pp.

Dwyer, F.J., D.K. Hardesty, C.E. Henke, C.G. Ingersoll, D.W. Whites, D.R. Mount and C.M. Bridges. 1999b. Assessing contaminant sensitivity of endangered and threaten ed species: Effluent toxicity tests. EPA/600/R-99/099. National Technical Information Service, Springfield, VA. 9 pp.

Ekelund, R., A. Bergman, A. Granmo, and M. Berggr en. 1990. Bioaccumulation of 4-nonylphenol in marine
animals--A re-evaluation. Environ. Poll. 64: 107-120.

### 73 Ekelund, R., A. Granmo, K. Magnusson and M. Berggren. 1993. Biodegradation of 4-nonylphenol in seawater and
sediment. Environ. Pollut. 79:59-61.

Ellis, D.D., C.M. Jone, R.A. Larson a nd D.J. Schaeffer. 1982. Organic constituents of mutagenic secondary effluents from wastewater treatment plants. Arch. Environ. Contam. Toxicol. 11:373-382.

England, D.E. 1995. Chronic toxicity of nonylphenol to Ceriodaphnia dubia . Report No. 41756. ABC Laboratories, Inc. Columbia, MO. 409 pp.

England, D.E. and J.B. Bussard. 1993. Toxicity of nonylphenol to the midge Chironomus tentans . Report No. 40597. ABC Laboratories, Inc., Columbia, MO. 2528 pp.

England, D.E. and J.B. Bussard. 1995. Toxicity of nonylphenol to the amphipod Hyalella azteca . Report No. 41569. ABC Laboratories, Inc., Columbia, MO. 178 pp.

Ernst, B., G. Julien, K. Doe and R. Parker. 1980. E nvironmental investigations of the 1980 spruce budworm spray program in New Brunswick. Surveillance Report EPS-5-AR -81-3, November 1980. Environment Canada, Atlantic Region, Halifax, NS.

Escher, M., T. Wahli, S. Buttner, W. Meier and P. Burkha rdt-Holm. 1999. The effect of sewage plant effluent on brown trout ( Salmo trutta fario ): a cage experiment. Aquat. Sci. 61:93-110.

Etoh, H., S. Hageshita and K. Ina. 1997. An improved assay for attachment-promoting substances of the blue mussel, Mytilus edulis galloprovincialis . J. Mar. Biotechnol. 5:24-26.

Fay, A.A., F.J. Brownawell, A.A. Elskus and A.E. Mc Elroy. 2000. Critical body residues in the marine amphipod Ampelisca abdita : Sediment exposures with nonionic organic cont aminants. Environ. Toxicol. Chem. 19:1028-1035.

Ferguson, P.L. and B.J. Brownawell. 2003. Degradation of nonylphenol ethoxylates in estuarine sediment under
aerobic and anaerobic conditions. E nviron. Toxicol. Chem. 22:1189-1199.

Fliedner, A. 1993. Daphnia magna , Reproduction test (OECD No. 202). Fraunhofer

B Institute fur Umweltchemie und Okotoxikologie, Postfach 1260, W

B 5948 Schmallenberg

B Grafschaft, Germany. Report No. UBA

B 002/4

B 22 February.

### 74 Flouriot, G., F. Pakdel, B. Ducouret and Y. Valotaire. 1995. Influence of xenobiotics on rainbow trout liver estrogen receptor and vitellogenin gene expr ession. J. Molecular Endocrin. 15:143-151.

Folmar, L., M. Hemmer, N. Denslow, K. Kroll, A. Ch eek, H. Meredith, H. Richman and G. Grau. 1998. A comparison of an in vivo fish VTG assay with YES and E-screen to rank > relative estrogenicity = of natural pharmaceutical and xenoestrogens. Am. Zool. 38:112A.

Fort, D.J. and E.L. Stover. 1997. De velopment of short-term, whole-embryo a ssays to evaluate detrimental effects on amphibian limb development and metamorphosis using Xenopus laevis . In: Environmental Toxicology and Risk Assessment: Modeling and Risk Assessment. F.J. Dwyer, T.R. Doane and M.L. Hinman (Eds.). ASTM, STP 1317, Vol. 6, American Society for Testing a nd Materials, Philadelphia, PA., pp. 376-390.

Gaffney, P.E. 1976. Carpet and rug industry case study II: Biological effects. J. Water Pollut. Control Fed.
48:2731-2737.

Gerritsen, A., N. van der Hoeven and A. Pielaat. 1998. The acute toxicity of selected alkylphenols to young and
adult Daphnia magna . Ecotoxicol. Environ. Safety. 39:227-232.

Giesy, J.P., S.L. Pierens, E.M. Snyde r, S. Miles-Richardson, V.J. Kramer, S.A. Snyder, K.M. Nichols and D.A. Villeneuve. 2000. Effects of 4-nonylphenol on fecundity of biomarkers of estrogenicity in fathead minnows ( Pimephales promelas ). Environ. Toxicol. Chem. 19:1368-1377.

Giger, W., E. Stephanou and C. Schaffner. 1981. Persis tent organic chemicals in sewage effluents: I. Identifications of nonylphenols and nonylphenol ethoxylat es by glass capillary gas chromatography/ mass spectrometry. Chemosphere 10:1253-1263.

Giger, W., P.H. Brunner and C. Schaffner. 1984. 4- Nonylphenol in sewage sludge; Accumulation of toxic
metabolites from nonionic surfactants. Science 225:623-625.

Granmo, A., R. Ekelund, K. Magnusson and M. Berggren. 1989. Lethal and sublethal toxicity of 4-nonylphenol to the common mussel ( Mytilus edulis L.). Environ. Poll.59: 115-127.

Granmo, A., R. Ekelund, M. Berggren and K. Magnusson. 1991a. Toxicity of 4-nonylphenol to aquatic organisms and potential for bioaccumulation. Proceedings, Swedish Environmental Protection Agency Seminar on Nonylphenol Ethoxylates/nonylphenol, Saltsjobade n, Sweden, February 6-8, pp. 53-75.

### 75 Granmo, A., S. Kollberg, M. Berggren, R. Ekelund, K. Magnusson, L. Renberg and C. Wahlberg. 1991b. Bioaccumulation of nonylphenol in caged mussels in an indus trial coastal area on the Swedish west coast. In: Organic Micropollutants in the Aquatic Environment. Angeletti, G. (Ed.). Proceedings of the 6 th European Symposium. Pp. 71-79.

Gray, M.A. and C.D. Metcalfe. 1997. Induc tion of testis-ova in Japanese medaka ( Oryzias latipes ) exposed to p

nonylphenol. Environ. Toxicol. Chem. 16:1082-1086.

Gross-Sorokin, M.Y., E.P.M. Grist, M. Cooke and M. Cr ane. 2003. Uptake and depuration of 4-nonylphenol by the
benthic invertebrate Gammarus pulex : How important is feeding rate? Environ. Sci. Technol. 37:2236-2241.

Hahn, T., K. Schenk and R. Schulz. 2002. Environmen tal chemicals with known endocrine potential affect yolk protein content in the aquatic insect Chironomus riparius . Environ. Pollut. 120:525-528.

Hale, R.C., C.L. Smith, P.O. de Fur, E. Harvey, E.O. Bu sh, M.J. La Guardia and G.G. Vadas. 2000. Nonylphenols
in sediments and effluents associated with diverse wa stewater outfalls. Environ. Toxicol. Chem. 19:946-952.

Hansen, F.T., V.E. Forbes and T.L. Forbes. 1999. E ffects of 4-n-nonylphenol on life-history traits and population dynamics of a polychaete. Ecol. Appl. 9:482-495.

Hansen, P., H. Dizer, B. Hock, A. Marx, J. Sherry, M. Mc Master and C. Blaise. 1998. Vitellogenin - a biomarker for endocrine disruptors. Trends Anal. Chem. 17:448-451.

Harries, J.E., D.A. Sheahan, S. Jobling, P. Matthiessen, P. Neall, J.P. Sumpter, T. Tylor and N. Zaman. 1997. Estrogenic activity in five United Kingdom rivers detected by measurement of vitellogenesis in caged male trout.
Environ. Toxicol. Chem. 16:534-542.

Harris, C.A., E.M. Santos, A Janbakhsh, T.G. Pottinger, C.R. Tyler and J.P. Sumpter. 2001. Nonylphenol affects gonadotropin levels in the pituitary gland and plasma of female rainbow trout. Environ. Sci. Technol. 35:2909-2916.

Harvilicz, H. 1999. NPE demand remains strong despite e nvironmental concerns in Europe. Chemical Market Reporter. October 18, 1999. p. 15.

Haya, K., L.E. Burridge and T.J. Benfey. 1997. The e ffect of cortisol and nonylphenol on growth and ornithine decarboxylase activity of juvenile Atlantic salmon, Salmo salar . Can. Tech. Rep. Fish. Aquat. Sci. 2192: 85-86.

### 76 Hecht, S. and B.L. Boese. 2002. Sensitivity of an infaunal amphipod, Eohaustorius estuarius , to acute waterborne exposures of 4-nonylphenol: Evidence of a toxic hangover. Environ. Toxicol. Chem. 21:816-819.

Hecht, S.A., J.S. Gunnarsson, B.L. Boese, J.O. Lamb erson, C. Schaffner, W. Giger and P.C. Jepson. 2004. Influences of sedimentary organic matter quality on th e bioaccumulation of 4-nonylphenol by estuarine amphipods.
Environ. Toxicol. Chem. 23:865-873.

Heinis, L.J., M.L. Knuth, K. Liber, B.R. Sheedy, R.L. Tunell and G.T. Ankl ey. 1999. Persistence and distribution of 4-nonylphenol following repeated application to littora l enclosures. Environ. Toxicol. Chem. 18:363-375.

Hemmer, M.J., B.L. Hemmer, C.J. Bowman, K.J. Kroll, L.C. Folmer, D. Marcovich, M.D. Hoglund and N.D. Denstow. 2001. Effects of p-nonylphenol, methoxychlor and endosulfan on vitellogenin induction and expression in sheepshead minnow ( Cyprindon variegatus ). Environ. Toxicol. Chem. 20:336-343.

Hense, B.A., I. Juttner, G. Welzl, G.F. Severin, G. Pfister, A. Behechti a nd K-W. Schramm. 2003. Effects of 4- nonylphenol on phytoplankton and periphyton in aquatic mi crocosms. Environ. Toxicol. Chem. 22:2727-2732.

Hewitt, L.M., L. Tremblay, G.J. Van Der Kraak, K.R. So lomon and M.R. Servos. 1998. Identification of the lampricide 3-trifluoromethyl-4-nitrophenol as an agonist for the rainbow trout estrogen receptor. Environ. Toxicol.
Chem. 17:425-432.

Holcombe, G.W., G.L. Phipps, M.L. Knuth and T. Felhab er. 1984. The acute toxicity of selected substituted phenols, benzenes and benzoic acid esters to fathead minnows ( Pimephales promelas ). Environ. Pollut. (Series A)
35:367-381.

Holmes, S. B. and P. D. Kingsbury. 1980. The environmental impact of nonylphenol and the Matacil 7 formulation. Part 1: Aquatic ecosystems. Report FPMX-35. Forest Pest Management Institute, Canadian Forestry Service, Sault Ste. Marie, Ontario.

Islinger, M., S. Pawlowski, H. Hollert, A. Volkl and T. Braumbeck. 1999. Measurement of vitellogenin-mRNA expression in primary cultures of rainbow trout hepatocyte s in a non-radioactive dot blot/RNAse protection-assay.
Sci. Total Environ. 233:109-122.

Jobling, S. and J.P. Sumpter. 1993. Detergent components in sewage effluent are weakly oestrogenic to fish: An in vitro study using rainbow trout ( Oncorhynchus mykiss ) hepatocytes. Aquat. Toxicol. 27:361-372.

### 77 Jobling, S., D. Sheahan, J.A. Osborne, P. Matthiessen and J. P. Sumpter. 1996. Inhibition of testicular growth in rainbow trout ( Oncorhynchus mykiss ) exposed to estrogenic alkylphenolic ch emicals. Environ. Toxicol. Chem.
15:194-202.

Jones, S.B., L.B. King, L.C. Sappington, F.J. Dwyer, M. Ellersieck and D.R. Buckler. 1998. Effects of carbaryl, permethrin, 4-nonylphenol, and copper on muscarinic choliner gic receptors in brain of surrogate and listed fish species. Comp. Biochem. Physiol. 120:405-414.

Kahl, M.D., E.A. Makynen, P.A. Kosian and G.T. Ankley. 1997. Toxicity of 4-nonylphenol in a life-cycle test with the midge Chironomus tentans . Ecotoxicol. Environ. Saf. 38:155-160.

Kannan, K., T.L. Keith, C.G. Naylor, C.A. Staples, S.A. Snyder and J.P. Giesy. 2003. Nonylphenol and nonylphenol ethoxylates in fish, sediment, and water from the Kalamazoo River, Michigan. Arch. Environ. Contam.
Toxicol. 44:77-82.

Keith, T.L., S.A. Snyder, C.G. Naylor, C.A. Staples, C. Summer, K. Kannan and J.P. Giesy. 2001. Identification and quantitation of nonylphenol ethoxylates and nonylphenol in fi sh tissues from Michigan. Environ. Sci. Technol.
35:10-13.

Kelly, S.A. and R.T. Di Giulio. 2000. Developmen tal toxicity of estrogenic alkylphenols in killifish ( Fundulus
heteroclitus ). Environ. Toxicol. Chem. 19:2564-2570.

Kinnberg, K., B. Korsgaard, P. Bjerregaard and A. Jespersen. 2000. Effects of nonylphenol and 17 β -estradiol on vitellogenin synthesis and testis morphology in male platyfish Xiphorphorus maculatus . J. Exper. Biol. 203:171-
181.

Kitajima, F., C.G. Satuito, H. Hirota, I. Katsuyama and N. Fusetani. 1995. A new screening method for antifouling substances against the young mussels Mytilus edulis galloprovincialis . Fish. Sci. 61:578-583. Kloas, W., I. Lutz and R. Einspanier. 1999. Amphibian s as a model to study endocrine disruptors: II. Estrogenic activity of environmental chemicals in vitro and in vivo. Sci. Total Environ. 225:59-68.

Knudsen, F.R. and T.G. Pottinger. 1999. Interaction of e ndocrine disrupting chemicals, singly and in combination, with estrogen-, androgen-, and corticosteroid-binding sites in rainbow trout ( Oncorhynchus mykiss ). Aquat. Toxicol.
44:159-170.

### 78 Kolpin, D.W., E.T. Furlong, M.T. Meyer, E.M. Thurman, S.D. Zaugg, L.B. Barber and H.T. Buxton. 2002. Pharmaceuticals, hormones, and other organic wastewater contaminants in U.S. streams, 1999 - 2000: A national
reconnaissance. Environ. Sci. Technol. 36:1202-1211.

Kopf, W. 1997. Wirkung endodriner stoffe in biotests mit wasserorganismen. In: Stoffe mit endokriner wirkung in wasser. Bayerisches landesamt fur wasserwirt sc haft, Institut fur Wasserforschung, Munchen. 12pp.

Kuhn, A., W.R. Munns, Jr., D. Champlin, R. McKinney, M. Tagliabue, J. Serbst and T. Gleason. 2001. Evaluation of the efficacy of extrapolation population modeling to predict the dynamics of Americamysis bahia populations in the laboratory. Environ. Toxicol. Chem. 20:213-221.

Kusk, K.O. and L. Wollenberger. 1999. Fully defined saltwat er medium for cultivation of and toxicity testing with marine copepod Acartia tonsa . Environ. Toxicol. Chem. 18:1564-1567.

Kwak, H., M. Bae, M. Lee, Y. Lee, B. Lee, K. Kang, C. Chae, H. Sung, J. Shin, J. Kim, W. Mar, Y. Sheen and M. Cho. 2001. Effects of nonylphenol, bisphenol A, and their mixture on the viviparous swordtail fish ( Xiphophorus
helleri ). Environ. Toxicol. Chem. 20:787-795.

Lamche, G. and P. Burkhardt-Holm. 2000. Nonylphenol provoke s a vesiculation of the golgi apparatus in three fish epidermis cultures. Ecotoxicol. Environ. Saf. 47:137-148.

Larsson, D.G.J., M. Adolfsson-Erici, J. Parkkonen, M. Pettersson, A.H. Be rg, P. Olsson and L. Forlin. 1999. Ethinyloestradiol - an undesired fish c ontraceptive? Aquat. Toxicol. 45:91-97.

Lavado, R., R. Thibaut, D. Raldua, R. Martin and C. Port e. 2004. First evidence of endocrine disruption in feral carp from the Ebro River. Toxicol. Appl. Pharm. 196:247-257.

LeBlanc, G.A., S. Mu and C.V. Rider. 2000. Embryot oxicity of the alkylphenol degradation product 4-nonylphenol
to the crustacean Daphnia magna . Environ. Health Perspect. 108:1133-1138.

Lech, J. J., S. K. Lewis and L. Ren. 1996. In vivo estrogenic activity of nonylphenol in rainbow trout. Fund. Appl.
Toxicol. 30:229-232.

Levine, C. and M.A. Cheney. 2000. Metabolic respons es to acute toxicity of alkylphenols and alkylphenol polyethoxylates in Elliptio complanata measured by calorespirometry. Environ. Toxicol. Chem. 19:1906-1910.

### 79 Lewis, M.A. 1991. Chronic and sublethal toxicities of surfactants to aquatic animals: A review and risk
assessment. Water Res. 25:101-113.

Lewis, S. K. and J. J. Lech. 1996. Uptake, disposition, and persistence of nonylphenol from water in rainbow trout ( Oncorhynchus mykiss ). Xenobiotica 26:813-819.

Liber, K., J. A. Gangl, T. D. Corry, L. J. Heinis a nd F. S. Stay. 1999. Lethality and bioaccumulation of 4- nonylphenol in bluegill sunfish in littoral enclos ures. Environ. Toxicol. Chem. 18:394-400.

Loomis, A.K. and P. Thomas. 1999. Binding characteris tics of estrogen receptor (ER) in Atlantic croaker ( Micropogonias undulatus ) testis: Different affinity for estrogens a nd xenobiotics from that of hepatic ER. Biol. Reprod. 61:51-60.

Lussier, S.M., D. Champlin, J. LiVolsi, S. Pouc her and R.J. Pruell. 2000. Acute toxicity of para -nonylphenol to
saltwater animals. Envir on. Toxicol. Chem. 19:617-621.

Lutz, I. and W. Kloas. 1999. Amphibians as a model to study endocrine disruptors: I. Environmental pollution and estrogen receptor binding. Sci. Total Environ. 25:49-57.

Lye, C.M., C.L.J. Frid, M.E. Gill, D.W. Cooper and D.M. Jones. 1999. Estrogenic alkylphenols in fish tissues, sediments, and waters from the U.K. Tyne and T ees estuaries. Environ. Sci. Technol. 33:1009-1014.

MacPhee, C. and R. Ruelle. 1969. Lethal effects of 1888 chemicals upon four species of fish from western North America. Bulletin No. 3, Forest, Wildlife and Range Experi ment Station, University of Idaho, Moscow, ID. 112 p.

Madsen, S.S., A.B. Mathiesen and B. Korsgaard. 1997. Effects of 17 β -estradiol and 4-nonylphenol on smoltification and vitellogenesis in Atlantic salmon ( Salmo salar ). Fish Physiol. Biochem. 17:303-312.

Magliulo, L., M.P. Schreibman and J.L. Cepriano. 1998. Disruption of the brain-pituitary-gonad axis of platyfish due to hormone mimicking environmental pollutants. Am. Zool. 38:112A.

Maguire, R.J. 1999. Review of the persisten ce of nonylphenol and nonylphenol ethoxylates in aquatic environments. Wat. Qual. Res. J. Can. 34:37-78.

Maki, H., H. Okamura, I. Aoyama and M. Fujita. 1998. Ha logenation and toxicity of the biodegradation products of a nonionic surfactant, nonylphenol ethoxylate. Environ. Toxicol. Chem. 17:650-654.

### 80 Manzano, M.A., J.A. Perales, D. Sales and J.M. Quiroga. 1998. Effect of concentra tion on the biodegradation of a nonylphenol polyethoxylate in river water. Bu ll. Environ. Contam. Toxicol. 61:489-496.

Manzano, M.A., J.A. Perales, D. Sales and J.M. Quiroga. 1999. The effect of temperature on the biodegradation of a nonylphenol polyethoxylate in river water. Water Res. 33:2593-2600.

Marcomini, A., B. Pavoni, A. Sfriso and A.A. Orio. 1990. Persistent metabolites of alkylphenol polyethoxylates in
the marine environment. Mar. Chem. 29:307-324.

McLeese, D.W., D.B. Sargeant, C.D. Metcalfe, V. Zit ko and L.E. Burridge. 1980a. Uptake and excretion of aminocarb, nonylphenol, and pesticides diluent 585 by mussels ( Mytilus edulis ). Bull. Environ. Contam. Toxicol.
24:575-581.

McLeese, D.W., V. Zitko, C.D. Metcalf and D.B. Sergean t. 1980b. Lethality of aminocarb and the components of the aminocarb formulation to juvenile Atlantic salmon, marine invertebrates and a freshwater clam. Chemosphere.
9:79-82.

McLeese, D.W., V. Zitko, D.B. Sargeant, L. Burridge and C.D. Metcalfe. 1981. Lethality and accumulation of alkylphenols in aquatic fauna. Chemosphere 10:723-730.

Meldahl, A. C., K. Nithipatikom and J. J. Lech. 1996. Metabolism of several 14 C-nonylphenol isomers by rainbow trout ( Oncorhynchus mykiss ): in vivo and in vitro microsomal metabolites. Xenobiotica 26:1167-1180.

Miles-Richardson, S.R., S.L. Pierens, K.M. Nichols, V.J. Kramer, E.M. Snyde r, S.A. Snyder, J.A. Render, S.D. Fitzgerald and J.P. Giesy. 1999. Effects of wate rborne exposure to 4-nonylphenol and nonylphenol ethoxylate on secondary sex characteristics and gonads of fathead minnows ( Pimephales promelas ). Environ. Res. (Series A)
80:S122-S137.

Milligan, S.R., O. Khan and M. Nash. 1998. Competitive binding of xenobiotic estrogens to rat alpha-fetoprotein and to sex steroid binding proteins in human and rainbow trout ( Oncorhynchus mykiss ) plasma. Gen. Comp.
Endocrinol. 112:89-95.

Moore, A., A.P. Scott, N. Lower, I. Katsiadaki and L. Greenwood. 2003. The effects of 4-nonylphenol and atrazine on Atlantic salmon ( Salmo salar L. ) smolts. Aquaculture 222:253-263.

### 81 Moore, S.B., R.A. Diehl, J.M. Barnhardt and G.B. Avery. 1987. Aquatic toxicities of textile surfactants. Text.
Chem. Color. 19:29-32.

Muller, R. 1980. Fish toxicity and surface tension of non-i onic surfactants: investigations of antifoam agents. J.
Fish. Biol. 16:585-589.

Naylor, C.G. 1992. Environmental fate of alkylphenol ethoxylates. Soap Cosmetics Chemical Specialities. August.

Naylor, C.G., J.P. Mieure, W.J. Adams, J.A. Weeks, F.J. Castaldi, L.D. Ogle and R.R. Romano. 1992. Alkylphenol ethoxylates in the environment. J. Am. Oil Chem. Soc. 69:695-708.

Nice, H.E., M.C. Thorndyke, D. Morrittt, S. Steel e and M. Crane. 2000. Development of the Crassostrea gigas larvae is affected by 4-nonylphenol. Mar. Pollut. Bull. 40:491-496.

Nimrod, A.C. and W.H. Benson. 1996. Estrogeni c responses to xenobiotics in channel catfish ( Ictalurus punctatus ).
Mar. Environ. Res. 42:155-160.

Nimrod, A.C. and W.H. Benson. 1997. Xenobiotic inter action with and alteration of channel catfish estrogen receptor. Toxicol. Appl. Pharmacol. 147:381-390.

Nimrod, A.C. and W.H. Benson. 1998. Reproduction and de velopment of Japanese medaka following an early life stage exposure to xenoestrogens. Aquat. Toxicol. 44: 141-156.

O = Halloran, S. L., K. Liber, J. A. Gangl and M. L. Knut h. 1999. Effects of repeat ed exposure to 4-nonylphenol on the zooplankton community in littoral enclosur es. Environ. Toxicol. Chem. 18:376-385.

Palmer, B.D., S.K. Palmer, C. Burch, J. Danyo and K.W. Selcer. 1998. Effects of endoc rine disrupting chemicals on amphibians. Am. Zool. 38( 5): Abstract No. 177A.

Patoczka, J. and G.W. Pulliam. 1990. Biodegradation and secondary effluent toxicity of ethoxylated surfactants.
Water Res. 24:965-972.

Pedersen, S.N., L.B. Christiansen, K.L. Pedersen, B. Ko rsgaard and P. Bjerregaard. 1999. In vivo estrogenic activity of branched and linear alkylphenols in rainbow trout ( Oncorhynchus mykiss ). Sci. Tot. Environ. 233:89-96.

### 82 Petit, F., P. Le Goff, J.P. Cravedi, Y. Valotaire and F. Pakdel. 1997. Two complementary bioassays for screening the estrogenic potency of xenobiotics: recombinant yeast for trout estrogen receptor and trout hepatocyte cultures. J.
Mol. Endocrinol. 19:321-335.

Petit, F., P. Le Goff, J.P. Cravedi, O. Kah, Y. Valotaire and F. Pakdel. 1999. Trout estrogen receptor sensitivity to xenobiotics as tested by different bioassays. Aquaculture. 177:353-365.

Pickford, K.A., R.E. Thomas-Jones, B. Wheals, C.R. Tyler and J.P. Sumpter. 2003. Route of exposure affects the oestrogenic response of fish to 4- tert -nonylphenol. Aq. Toxicol. 65:267-279.

Prasad, R. 1986. Effects of nonylphenol adjuvant on m acrophytes. In: Adjuvants and Agrochemicals. Chow, P.N.P. (Ed.). Vol. 1, Mode of action and physiological activity. First International Symposium. Brandon, Manitoba, Canada. August 5-7. XIV+207 pp.

Preston, B.L., T.W. Snell, T.L. Robertson and B.J. Dingmann. 2000. Use of freshwater rotifer Brachionus calyciflorus in screening assay for potential endocrine di sruptors. Environ. T oxicol. Chem. 19:2923-2928.

Purdom, C.E., P.A. Hardiman, V.J. Bye, N.C. Eno, C.R. Tyler and J.P. Sumpter. 1994. Estrogenic effects of effluents from sewage treatment works. Chem. and Ecol. 8:275-285.

Radian Corp. 1990. Nonylphenol and nonylphenol ethoxylates in river water and bottom sediments-January 1989- August 1990. Final Report to Alkylphenol and Ethoxylat es Panel, Chemical Manufacturers Association.

Reed, H.W.B. 1978. Alkylphenols. In: Kirk-Othmer: Encyclopedia of chemical technology. Grayson M. and EcKroth D. (Eds.). 3rd ed., Vol. 2, J ohn Wiley and Sons, New York. pp. 72-96.

Ren, L., D. Lattier and J. J. Lech. 1996a. Estrogenic ac tivity in rainbow trout determined with a new cDNA probe for vitellogenesis, pSG5Vg1.1. Bull. Environ. Contam. Toxicol. 56:287-294.

Ren, L., S.K. Lewis and J.J. Lech. 1996b. Effect s of estrogen and nonylphenol on the post-transcriptional regulation of vitellogenin gene expressi on. Chemico-Biol. Interact. 100:67-76.

Rice, C.D., L.E. Roszell, M.M. Banes and R.E. Arnol d. 1998. Effects of diet ary PCBs and nonyl-phenol on immune function and CYP1A activity in channel catfish, Ictalurus punctatus . Mar. Environ. Res. 46:351-354.

### 83 Routledge, E.J. and J.P. Sumpter. 1996. Estrogenic activity of surfactants and some of their degradation products assessed using a recombinant yeast screen . Environ. Toxicol. Chem. 15:241-248.

Routledge, E.J. and J.P. Sumpter. 1997. Structural feat ures of alkylphenolic chemical s associated with estrogenic activity. J. Biol. Chem. 272:3280-3288.

Roy F. Weston Inc. 1990. Determination of the vapor pressure of 4-nonylphenol. Final Report Study No. 90-047. Roy F. Weston Inc., Environmental Fate and Effects La boratory, 254 Welsh Pool Road, Lionville, PA. 15 August
1990.

Schmude, K. L., K. Liber, T. D. Corry and F. S. Stay. 1999. Effects of 4-nonylphenol on benthic macroinvertebrates and insect emergence in littoral enclosures. Environ. Toxicol. Chem. 18:386-393.

Schultz, T.W. 1997. Tetratox: Tetrahymena pyriformis population growth impairment endpoint - A surrogate for fish lethality. Toxicol. Methods. 7:289-309.

Schwaiger, J., O.H. Spieser, C. Bauer, H. Ferling, U. Ma llow, W. Kalbfus and R.D. Negele. 2000. Chronic toxicity of nonylphenol and ethinylestradiol: haematological and histopathological effects in juvenile common carp ( Cyprinus carpio ). Aquat. Toxicol. 51:69-78.

Schwaiger, J., U. Mallow, H. Ferling, S. Knoerr, T. Braunbeck, W. Kalbfus and R.D. Negele. 2002. How estrogenic is nonylphenol: A transgenerational study using rainbow trout ( Oncorhynchus mykiss ) as a test organism. Aquat. Toxicol. 59:177-189.

Seki, M., H. Yokota, M. Maeda, H. Tadokoro and K. Kobayashi. 2003. Effects of 4-nonylphenol and 4- tert

octylphenol on sex differentiation and vitellogenin induction in medaka ( Oryzias latipes ). Environ. Toxicol. Chem.
22:1507-1516.

Servos, M.R. 1999. Review of the aquatic toxicity, estrogenic responses and bioaccumulation of alkylphenols and alkylphenol polyethoxylates. Wat. Qual. Res. J. Can. 34:123-177.

Severin, G.F., G. Welzl, I. Juttner, G. Pfister and K- W. Schramm. 2003. Effects of nonylphenol on zooplankton in aquatic microcosms. Environ. Toxicol. Chem. 22:2733-2738.

### 84 Shackelford, W.M., D.M. Cline, L. Faas and G. Kurth. 1983. Evaluation of automated spectrum matching for survey identification of wastewater components by gas chromatography-mass spectrometry. PB83-182931. National Technical Information Service, Springfield, VA.

Shang, D.Y., R.W. MacDonald and M.G. Ikonomou. 1999. Persistence of nonylphenol ethoxylate surfactants and their primary degradation products in sediments from near a municipal outfall in the Strait of Georgia, British Columbia, Canada. Environ. Sci. Technol. 33:1366-1372.

Shurin, J.B. and S.I. Dodson. 1997. Sublethal toxic e ffects of cyanobacteria and nonylphenol on environmental sex determination and development in Daphnia . Environ. Toxicol. Chem. 16:1269-1276.

Sonnenschein, C. and A.M. Soto. 1998. An updated revi ew of environmental estrogen and androgen mimics and antagonists. J. Steroid Bioc hem. Molec. Biol. 65:143-150.

Soto, A.M., H. Justicia, J.W. Wray and C. Sonnenschein. 1991. p -Nonylphenol: An estrogenic xenobiotic released from modified polystyrene. Environ. Health Perspect. 92:167-173.

Soto, A.M., T.M. Lin, H. Justicia, R.M. Silvia and C. Sonnenschein. 1992. An "in culture" bioassay to assess the estrogenicity of xenobiotics (E-SCREEN). In: Advances in Modern Environmental Toxicology. Volume XXI. Chemically-induced alterations in sexual and functional de velopment: The wildlife/human connection. T. Colburn and C. Clement (Eds.). Princeton Scien tific Publishing Co., Inc., NJ. 403 pp.

Spieser, O.H., J. Schwaiger, H. Ferling, W. Kalbfus and R. Negele. 1998. Effects of nonylphenol and ethinylestradiol on swimming behavior of juvenile carp. Abstract for Annual Meeting. International Association for Great Lakes Research. Hamilton, Ontario, Canada. pp. 37-38.

Staples, C.A., J. Weeks, J.F. Hall and C.G. Naylor. 1998. Evaluation of aquatic toxicity and bioaccumulation of C8- C9-alkylphenol ethoxylates. E nviron. Toxicol. Chem. 17:2470-2480.

Staples, C.A., J.B. Williams, R.L. Blessing and P.T. Varineau. 1999. Measuring the biodegradability of nonylphenol, ether carboxylates, octylphenol, ether car boxylates, and nonylphenol. Chemosphere 38:2029-2039.

Staples, C.A., C.G. Naylor, J.B. Williams and W.E. Gledhill. 2001. Ultimate biodegradation of alkylphenol ethoxylate surfactants and their biodegradation inte rmediates. Environ. Toxicol. Chem. 20:2450-2455.

Stephan, C.E., D.I. Mount, D.J. Hansen, J.H. Gentile, G.A. Chapman and W.A. Brungs. 1985. Guidelines for

### 85 deriving numerical national water quality criteria for th e protection of aquatic organisms and their uses. http://www.epa.gov/waterscience/criteria/85guidelines.pdf
Sumpter, J.P. 1998. Xenoendocrine disrupters -- e nvironmental impacts. Toxicol. Lett. 102-103:337-342.

Sundaram, K.M.S. and S. Szeto. 1981. The dissipati on of nonylphenol in stream and pond water under simulated
field conditions. J. Environ. Sci. Health (Part B) B16(6):767-776.

Sundaram, K.M.S., S. Szeto, R. Hindle and D. MacTavish. 1980. Residues of nonylphenol in spruce foliage, forest soil, stream water and sediment after its aerial applicati on. J. Environ. Sci. Hea lth (Part B) B15(4):403-419.

Tabira, Y., M. Nakai, D. Asai, Y. Yakabe, Y. Taha ra, T. Shinmyozu, M. Noguchi, M. Takatsuki and Y. Shimohigashi. 1999. Structural requirements of para -alkylphenols to bind to estrogen receptor. Eur. J. Biochem.
262:240-245.

Tagliabue, M.D. 1993. Nonylphenol Champia results. Memo dated February 23 to S. Poucher, Science Applications International Corpor ation, Narragansett, RI. 6 pp.

Takasawa, R., H. Etoh, A. Yagi, K. Sakata and K. Ina. 1990. Nonylphenols as promising antifouling agents found by a simple bioassay method using the blue mussel, Mytilus edulis . Agric. Biol. Chem. 54:1607-1610.

Tanghe, T., G. Devriese and W. Verstraete. 1999. Nonyl phenol and estrogenic activity in aquatic environmental samples. J. Environ. Qual. 28:702-709.

Thibaut, R., L. Debrauwer, D. Rao and J. Cravedi. 1998. Characterization of biliary metabolites of 4-n-nonylphenol in rainbow trout ( Oncorhynchus mykiss ). Xenobiotica. 28:745-757.

Thibaut, R., L. Debrauwer, D. Rao and J.P. Cravedi. 1999. Urinary metabolites of 4-n-nonylphenol in rainbow trout (Oncorhynchus mykiss). Sci. Total Environ. 233:193-200.

Tsuda, T., A. Takino, M. Kojima, H. Harada, K. Muraki and M. Tsuji. 2000. 4-Nonylphenols and 4-tertoctylphenol in water and fish from rivers fl owing into Lake Biwa. Chemosphere 41:757-762.

Tremblay, L. and G. Van Der Kraak. 1998. Use of a series of homologous in vitro and in vivo assays to evaluate the endocrine modulating actions of β -sitosterol in rainbow trout. Aquat. Toxicol. 43:149-162.

### 86 Turner, A.H., F.S. Abram, V.M. Brown and H.A. Painte r. 1985. The biodegradability of two primary alcohol ethoxylate nonionic surfactants under practical conditions, and the toxicity of biodegradation products to rainbow
trout. Water Res. 19:45-51.

Uguz, C., M. Iscan, A. Erguven, B. Isgor and I. Togan. 2003. The bioaccumulation of nonyphenol and its adverse effect on the liver of rainbow trout ( Onchorynchus mykiss ). Environ. Res. 92:262-270.

University of Wisconsin-Superior. 1985. Acute t oxicities of organic chemicals to fathead minnows ( Pimephales promelas ). Volume II. D.L. Geiger, C.E. Northcott, D.J. Call and L.T. Brooke (Eds.). Center for Lake Superior Environmental Studies, University of Wisconsin-Superior, Superior, WI. 326 pp.

U.S. EPA. 1983. Water quality standards regulation. Federal Register 48:51400-51413. November 8.

U.S. EPA 1986. Chapter I - Stream design flow for stea dy-state modeling. In: Book VI - Design conditions. In: Technical guidance manual for performing waste load allocation. Office of Water, Washington, DC.
U.S. EPA. 1987. Permit writers guide to water quality-based permitting for toxic pollutants. Office of Water, Washington, DC. EPA-440/4-87-005.

U.S. EPA. 1991. Technical support documen t for water quality-based toxics control. Office of Water, Washington, DC. EPA-505/2-90-001. Available for download at: http://www.epa.gov/npdes/pubs/owm0264.pdf
U.S. EPA. 1994. Water quality standards handbook. Second Edition. Office of Water, Washington, DC. EPA-823-B94-005. Available for download at: http://www.epa.gov/waterscience/standards/handbook/

USITC. 1981. Synthetic organic chemi cals: United States production and sale s, 1980. United States International
Trade Commission Publication No. 1183. U.S. Governme nt Printing Office, Washington, DC 20402. 327 p.

USITC. 1989. Synthetic organic chemi cals: United States production and sale s, 1988. United States International
Trade Commission Publication No. 2219. U.S. Gove rnment Printing Office, Washington, DC 20402.

Varma, M.M. and D. Patel. 1988. Nonionic surfactan ts in perspective. J. Environ. Systems 18:87-96.

Veith, G.D. and O.G. Mekenyan. 1993. A QSAR approach for estimating the aquatic toxicity of soft electrophiles (QSAR for soft electrophiles). Quant. Struct. Act. Relat. 12:349-356.

### 87 Villenueve, D.L., S.A. Villalobos, T.L. Keith, E.M. Snyder, S.D. Fitzgerald and J.P. Giesy. 2002. Effects of waterborne exposure to 4-nonylphenol on plasma sex steroi d and vitellogenin concentrations in sexually mature male carp ( Cyprinus carpio ). Chemosphere 47:15-28.

Ward, T.J. and R.L. Boeri. 1990a. Acute static toxicity of nonylphenol to the freshwater alga Selenastrum capricornutum . EnviroSystems Study No. 8969-CMA. Resour ce Analysts, Inc., Hampton, NH. 41 pp.

Ward, T.J. and R.L. Boeri. 1990b. Acute flow through toxicity of nonylphenol to the mysid, Mysidopsis bahia

. Study Number 8974-CMA. EnviroSystems, Hampton, NH. 35 pp.

Ward, T.J. and R.L. Boeri. 1990c. Acute flow th rough toxicity of nonylphenol to the sheepshead minnow, Cyprinodon variegatus . Study Number 8972-CMA. EnviroSystems, Hampton, NH. 34 pp.

Ward, T.J. and R.L. Boeri. 1990d. Acute sta tic toxicity of nonylphenol to the marine alga Skeletonema costatum

. Study Number 8970-CMA. EnviroSystems, Hampton, NH. 42 pp.

Ward, T.J. and R.L. Boeri. 1991a. Bioconcentration test with nonylphenol and the fathead minnow, Pimephales promelas . Study Number 8975-CMA. Envirosystems, Hampton, NH. 72 pp.

Ward, T.J. and R.L. Boeri. 1991b. Chr onic toxicity of nonylphenol to the mysid, Mysidopsis bahia . Study Number 8977-CMA. EnviroSystems, Hampton, NH. 61 pp.

Ward, T.J. and R.L. Boeri. 1991c. Early life st age toxicity of nonylphenol to the fathead minnow, Pimephales promelas . Study Number 8979 CMA. EnviroSystems, Hampton, NH. 59 pp.

Ward, T.J. and R.L. Boeri. 1992. Toxicity of nonylphenol to the tadpole Rana catesbiana . Study Number 8981- MA. EnviroSystems, Hampton, NH. 78 pp.

Weinberger, P. and R. Greenhalgh. 1984. Some adjuvant effects on the fate of fenitrothion and aminocarb.
Environm. Toxicol. Chem. 3:325-334.

Weinberger, P. and S. Iyengar. 1983. Effects of aminocarb, fuel oil 585 and nonylphenol on the growth and development of Lemna minor L. Dev. Ecol. Environ. Qual. Shuval, H. I. (Ed.). Proc. Int. Meet. Isr. Ecol. Soc. pp.
595-607.

### 88 Weinberger, P. and M. Rea. 1981. Nonylphenol: A perturba nt additive. Canadian Tech. Rep. Fish Aquat. Sci.
990:370-381.

Weinberger, P., C. DeChacin and M. Czuba. 1987. E ffects of nonyl phenol, a pesticide surfactant, on some metabolic processes of Chlamydomonas segnis . Can. J. Bot. 65:696-702.

White, R., S. Jobling, S.A. Hoare, J.P. Sumpter and M.G. Parker. 1994. Environmentally persistent alkylphenolic compounds are estrogenic. Endocrinology 135:175-182.

Wood, E. M. 1953. The toxicity of 3400 chemicals to fi sh. Report of Fish and Wildlife Service, Kearneysville, West Virginia. EPA 560/6-87-002. Office of Toxic Substances, Washington, DC, August.

Yadetie, F., A. Arukwe, A. Goksoyr and R. Male. 1999. I nduction of hepatic estrogen receptor in juvenile Atlantic salmon in vivo by the environmental estrogen, 4-nonylphenol. Sci. Total Environ. 233:201-210.

Yokota, H., M. Seki, M. Maeda, Y. Oshima, H. Tadokoro, T. Honjo and K. Kobayashi. 2001. Life-cycle toxicity of 4-nonylphenol to medaka ( Oryzias latipes ). Environ. Toxicol. Chem. 20:2552-2560.

Yoshimura, K. 1986. Biodegradation and fish toxicity of nonionic surfactants. J. Amer. Oil Chem. Soc. 63:1590-
1596.

Yoshioka, Y. 1985. Testing for the toxicity of chemicals with Tetrahymena pyriformis . Sci. Total Environ. 43:149-
157.

Zang, L., R. Gibble and K.N. Baer. 2003. The effect s of 4-nonylphenol and ethanol on acute toxicity, embryo development, and reproduction in Daphnia magna . Ecotoxicol. Environ. Safety 55:330-337.