## ASSESSMENT OF THE PERSISTENCE

AND BIOACCUMULATION POTENTIAL FOR NONYLPHENOL, OCTYLPHENOL,

AND THEIR ETHOXYLATES

FOR CATEGORIZATION AND SCREENING OF THE
CANADIAN DOMESTIC SUBSTANCE LIST (DSL)

Prepared for the Alkylphenols & Ethoxylates Research Council

By

G.M. Klecka, 1 

## C.A. Staples, 2 

## B.S. Losey, 3 

## and K.B. Woodburn 1

1 

### The Dow Chemical Company, Midland, MI

2 

### Assessment Technologies, Inc., Spotsylvania, VA

3 

### RegNet Environmental Services, Washington, DC



## September 16, 2005

### 1 

## TABLE OF CONTENTS

I. EXECUTIVE SUMMARY ...............................................................................2

II. SCOPE AND OBJECTIVES ...............................................................................2

III. DESCRIPTION OF SUBSTANCES ......................................................................3

IV. PERSISTENCE ASSESSMENT ..........................................................................8

V. BIOACCUMULATION ASSESSMENT ...............................................................18

VI. CONCLUSIONS .............................................................................................26

VII. REFERENCES ...............................................................................................28

APPENDIX: TABLES AND FIGURES ..........................................................................33

### 2 I. EXECUTIVE SUMMARY

Under the Canadian Environmental Protection Act of 1999, Environment Canada (EC) is required to

assess all substances on the Canadian Domestic Substa nces List (DSL) with respect to persistence, bioaccumulation, and inherent toxicity (PBiT) charact eristics. Applicable guidance for the evaluation of substances is given in EC (2003). Environment Cana da has provisionally categorized selected substances on the DSL for PBiT and is soliciting further informa tion to improve the technical basis of its initial

assessment.

This report is intended to inform the decision on th e categorization of persistence and bioaccumulation potential for the family of nonylphenols (NP), octylphenols (OP), and their ethoxylates by:

• Identifying additional studies either from th e published literature (or unpublished industry studies) not previously considered by Environment Canada;

• Performing a critical review and evaluation of the available data and defining the P and B

characteristics of the substances based on a weight of evidence approach; and,

• Comparing the conclusion of the analysis to those of existing risk assessment reports developed in other jurisdictions (European Union (EU), Environment Canada, etc).

The following substances are included in this assessment:

CAS RN Sort/Group
Nonylphenol 104-40-5 Organics 11066-49-2 Organics 25154-52-3 Organics

84852-15-3 UVCB Nonylphenol Ethoxylates 9016-45-9 Original Polymers 26027-38-3 Original Polymers 37205-87-1 Original Polymers 68412-54-4 UVCB Polymers 127087-87-0 UVCB Polymers Octylphenol

140-66-9 Organics 1806-26-4 Organics 27193-28-8 Organics Octylphenol Ethoxylates 9002-93-1 Original Polymers 9036-19-5 Original Polymers 68987-90-6 UVCB Polymers

The overall conclusion from this review is that all members of the family of nonylphenols, octylphenols and their ethoxylates do not meet the criteria for P or B . Although it is acknowledged that certain members of the family meet the criteria for inherent toxicity (iT) to environmental organisms, the fact that

criteria for P and B are not met indicates that none of the members of the family should be categorized as

PiT or BiT.

II. SCOPE & OBJECTIVES

The Canadian Environmental Protection Act of 1999 has mandated the categorization of substances on the Canadian Domestic Substance List (DSL) with r espect to persistence, bioaccumulation, and inherent toxicity (PBiT) characteristics by September 14, 2006. Environment Canada (EC, 2003) has published technical guidelines describing the systematic approach th at is to be used for this categorization process. Environment Canada has subsequently applied this scheme to preliminarily categorize 10,629 organic

### 3 substances and is seeking comments and additional s upporting P, B, and iT information on substances

included in this initial assessment.

For more than twenty years the Alkylphenols & Ethoxylates Research Council (APERC), its predecessor the American Chemistry Council’s APE Panel, and its member companies have been actively engaged in environmental fate and effects research on alkylphenol s and their ethoxylates. Consequently, the industry

members can contribute considerable information a nd expertise relevant to the environmental assessment

of these substances.

The objective of this report is to provide a techni cal review of the persistence and bioaccumulation potential of the family of nonylphenols, octylphenols, and their commercially relevant ethoxylates. The report will illustrate the extensive amount of experime ntal data on the environmental fate and effects, comment on studies used by Environment Canada to define pivot values, and provide an improved environmental assessment based on a weight of evidence approach.

It is important to note that the environmental behavi or, particularly the persistence and bioconcentration potential of the various NP, OP, and their ethoxylates, has been extensively summa rized in several review articles (Staples et al., 1998; Servos, 1999; Talmage, 1994). In addition, various regulatory authorities have recently published their decisions on the envir onmental properties for many of these substances in risk assessment reports (Environment Canada and H ealth Canada, 2001; European Union, 2001; United Kingdom Environment Agency, 2003) or Fact Sheets of PBT assessments (European Chemicals Bureau, 2004).

III. DESCRIPTION OF SUBSTANCES

Nonylphenol CAS RNs and Nomenclature

Nonylphenol is the commercial description for a comp lex mixture of nine-carbon alkyl-chain substituted phenols. NP is produced through the Friedel-Crafts alkylation of phenol with nonene, which, in the presence of an acid catalyst, preferentially alkylates at the para position of phenol. Commercial nonene

does not contain linear C 9 H 18 alpha -olefin; rather it is a complex mixture of highly branched, predominantly nine-carbon olefins known as propylene trimers. Therefore, the NP formed by the alkylation of phenol with propylene trimers is also a very complex mixture of branched isomers with the following approximate composition: ortho -NP (3-6%), para -NP (90-93%), and decylphenol (2-5%).

Since the para isomer predominates, the product is most accurately described as para -nonylphenol, branched ( p -NP or PNP). The following table lists the results of 

### high-resolution GC analyses of p- NP

and identifies 22 branched para -isomers, within five distinct groups. Group designations are presented based on the substitution of the alpha-carbon on the alkyl chain ( Bhatt, 1992; Kirk-Othmer, 1992; Wheeler, 1997; Thiele, 2004).

Para Isomers of Nonylphenol

Group # Isomer Type Number of

Isomers Para Isomers, %

1 Alpha-dimethyl 10 48.6%
2 Alpha-methyl, alphaethyl, beta-primary 3 8.9%

3 Alpha-methyl, betamethyl 4 24.7%

4 Alpha-methyl 2 6.6%
5 Alpha-methyl, alphapropyl 3 11.2%

Total Isomers 22 100%

### 4 The complexity of the chemistry and nomenclatur e that relates to NP has been recognized in governmental risk assessments. The EU Risk Assessment on NP simultaneously addressed CAS RN 84852-15-3 (EINECS No. 284-325-5) “4-Nonylphenol (branched)” and CAS RN 25154-52-3 (EINECS No. 246-672-0) “Nonylphenol” as equivalent commercially-relevant compounds. Furthermore, the EU Assessment recognized the following additional synonyms for NP: Isononylphenol (CAS RN 11066-49- 2); Phenol, nonyl-, branched (CAS RN 90481-04-2); and, monoalkyl (C3-C9) phenol. Note that CAS RN 90481-04-2 does not appear to be on the Domestic Substance List.

As explained in the EU Risk Assessment for NP, ch anges relating to nomenclature practice within the United States Environmental Protection Agency (US EPA) and Chemical Abstract Service (CAS) are behind the varied and confusing nomenclature for this compound. NP, CAS RN 25154-52-3 was originally defined by CAS to cove r all nonylphenols. However, sub sequent revisions in nomenclature practice and CAS RN assignments redefined this CAS RN to cover only straight chain NP. Given the method of manufacture for nonylphenols, the production of straight chain NP is essentially impossible; as such, this isomer is considered commercially irreleva nt. However, straight chain NP can be synthesized on a laboratory scale and is available as a research chemical.

In its assessment on NP, the EU assumed that data fro m any of the isomers were representative for NP unless otherwise specified, and NP was used as the gene ric name referring to all of these substances in the assessment (EU, 2001). The European Chemicals Bu reau (ECB) evaluated CAS RN 25154-52-3 based

on data conducted on CAS RN 84852-15-3 due to their analogous structures (ECB, 2003).

In the development of Water Quality Criteria fo r NP, US EPA utilized data from CAS RN 84852-15-3 and 25154-52-3 (US EPA, 2003). In its RM1 document for para -NP US EPA accepted the view of the Alkylphenols & Ethoxylates Research Council (APERC) that CAS RN 84852-15-3 is the most descriptive of the commercially relevant bran ched p-NP, while noting that CAS RN 25154-52-3 (nonylphenol, mixed isomers) a nd CAS RN 104-40-5 (4-nonylphenol) have been also reported by manufacturers to represent this compound (Rodier, 1996).

Environment Canada recognized the complexity in nomenclature and CAS RNs for NP in its Priority Substance List (PSL) Assessment for NP and nonyl phenol ethoxylates, which was conducted under the Canadian Environmental Protection Act (CEPA). Environment Canada acknow ledged the position of US EPA and the alkylphenols industry that CAS RN 84852- 15-3 best represents the commercially relevant nonylphenol (Environment Canada a nd Health Canada, 2001, p. 11). However, in the PSL Assessment,

data from studies with different and/or undefined CAS numbers for NP were taken as representative unless otherwise noted. Following is a table that summarizes CAS numbers for NP and their

nomenclature.

Nonylphenol (Sources: CAS, NIST, APERC) CAS RN Description Other Assessments Structure (Source CAS, NIST)

84852-15-3 Phenol, 4-nonyl-,

branched

Other Names Nonylphenol; 4-Nonylphenol; p-Nonylphenol, branched;

Branched 4- nonylphenol (mixed isomers) EU (2001) ECB (2003) US EPA (2003) US EPA (1996) Unspecified, mixed isomers

### 5 25154-52-3

Other CAS

RNs

84852-15-3

Former

CAS RNs 1300-16-9;

56459-00-4 Phenol, nonyl

Other Names Phenol, nonyl-; x-Nonylphenol; Nonylphenol (mixed isomers); Nonylphenol (mixture) EU (2001) ECB (2003) US EPA (2003) US EPA (1996)

11066-49-2 Isononylphenol

Other Names Phenol, isononyl; Isononylphenol (mixed isomers) EU (2001)

104-40-5 p-Nonylphenol

Other Names 4-Nonylphenol; Phenol, p-nonyl-; 4-n-Nonyl phenol; Phenol, 4-n-nonyl US EPA (1996)

In summary, based on current understanding of the va rious CAS RNs used to describe nonylphenol, there are in principle two forms of the material, i.e., lin ear NP as described by CAS RN 104-40-5 and branched NP, as best described by CAS RN 84852-15-3. However, CAS RNs 25154-52-3 and 11066-49-2 should

be considered as equivalent synonyms for branched NP for the categorization purposes. Further, to be consistent with previous regulatory evalua tions of the material, during the conduct of the categorization assessment Environment Canada should consider all data from any of the isomers as representative for NP.

Nonylphenol Ethoxylates CAS RNs and Nomenclature

Nonylphenol ethoxylates (NPEs) are produced by the based-catalyzed reaction of para -nonylphenol ( p

NP) with ethylene oxide (EO). The branching of th e nonyl group results in additional structural isomers.

Commercially available NPEs generally range from f our moles of ethoxylation (NPE4) up to 80 moles of ethoxylates (NPE80). While some CAS RNs specific to certain levels of ethoxylates do exist, the commercially relevant NPE CAS RNs discussed below are a pplicable to all levels of ethoxylation. By far the most commonly used NPE is manufactured to a target of nine moles of ethoxylation (NPE9). It is important to note that all NPE products are polymers w ith varying levels of ethoxylation distributed in a normal curve around a target ethoxylation level. The figure below provides an illustration of the oligomer

distribution for NPE9.

### 6 0 2 4 6 8 10 12 14

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18

NPE Oligomer NumberPercent of Total

Following is a list of the commercially relevant CAS RNs for NPEs that are known to the Alkylphenols & Ethoxylates Research Council. While these CAS RNs di ffer with respect to their level of description regarding branching and position of the nonyl gr oup on the phenol ring, APERC understands that manufacturers of NPEs use essentially the same starting materials and synthesis process and therefore commercially available NPEs have essentially the same structure and isomeric mix. Thus, the various

CAS RNs should be considered as equivalent synonyms for NPE for the purpose of categorization.

Commercially Relevant CAS RNs for Nonylphenol Ethoxylates (Source: APERC) CAS RN Description 9016-45-9 Poly (oxy-1,2-ethanediyl), alpha -(nonylphenyl)- omega -hydroxy26027-38-3 Poly (oxy-1,2-ethanediyl), alpha -(4-nonylphenyl)- omega -hydroxy - 37205-87-1 Poly (oxy-1,2-ethanediyl), alpha -(isononylphenyl)- omega -hydroxy68412-54-4 Poly (oxy-1,2-ethanediyl), alpha -(nonylphenyl)- omega -hydroxy-, branched 127087-87-0 Poly (oxy-1,2-ethanediyl), alpha -(4-nonylphenyl)- omega -hydroxy-, branched

Octylphenol CAS RNs and Nomenclature

Octylphenol describes a variety of isomeric compounds of the general formula C 6 H 4 (OH)C 8 H 17. The octyl group is a chain of eight carbons, which may be lin ear or branched. OP is produced by the reaction of phenol with octene. Commercial synthesis results in a mixture of various octylphenol isomers rather than a discrete chemical structure. The most important OP is made using the dimer of isobutylene and consists primarily of a single isomer 4-(1,1,3,3-t etramethylbutyl)phenol. The octyl group is positioned predominantly in the para position on the phenol ring, though isomers with the octyl group located at the

ortho position are also formed.

CAS RN 140-66-9 is considered to be the only commercially relevant identifier of OP (4- tert

octylphenol) and is recognized as such in a ri sk assessment conducted by the United Kingdom (UK) on the compound. Noting that octylphenol has been used in many studies with no further description of the isomer, the UK included such data where they were considered to be relevant. In addition, the UK Environment Agency (UKEA) used data from other isom ers of OP where the data could be considered to be representative and justified (UKEA, 2005). Oligomer Distribution of NPE9

### 7 Octylphenol (Sources: CAS, NIST, APERC) CAS RN Description Other

Assessments Structure

140-66-9 4- tert -octylphenol

Other names Phenol, 4-(1,1,3,3- tetramethylbutyl)-; Phenol, p-(1,1,3,3- tetramethylbutyl)-; p-(1,1,3,3- Tetramethylbutyl)phenol;

p- tert -Octylphenol; 4-(1,1,3,3- Tetramethylbutyl)phenol;

4- tert -Octylphenol; p-(1',1',3',3'- Tetramethylbutyl)phenol; p-Terc.oktylfenol;

para

- tert -Octylphenol; Phenol, p-( tert -octyl)- UKEA (2005)

1806-26-4 Phenol, 4-octyl

Other names Phenol, 4-octyl-; Phenol, p-octyl-; p-Octylphenol; 1-(p-Hydroxyphenyl)octane UKEA (2005)

27193-28-8 Phenol, octyl

Other names Phenol, (1,1,3,3-tetramethylbutyl) UKEA (2005)

In summary, based on current understanding of the va rious octylphenol isomers and the CAS numbers used to describe them, there are in principle two forms of the material, i.e., linear OP as described by CAS RN 1806-26-4 and branched OP, as best described by CAS RN 140-66-9. However, for the purposes of CSDSL categorization the other CAS RN (27193-28-8) should be considered as a comparable synonym for branched OP. Further, to be consistent with pr evious regulatory assessments, during the conduct of the categorization assessment Environment Ca nada should consider all data from any of the isomers as representative for OP.

Octylphenol Ethoxylates CAS RNs and Nomenclature

Octylphenol ethoxylates (OPEs) are produced by the based-catalyzed reaction of para -octylphenol ( p -OP) with ethylene oxide (EO). As with the NPEs, these products are polymers with varying levels of ethoxylation distributed in a normal curve around a ta rget ethoxylation level. Commercially available OPEs generally range from 4 moles of ethoxylati on (OPE4) up to 80 moles of ethoxylates (OPE80). Some branching of the octyl group results in additi onal structural isomers, although the isomer shown below is predominant.

### 8 The following three CAS RNs for OPE are known to be commercially relevant. These CAS RNs should

be considered as equivalent synonyms for OP E for the categorization purposes.

Commercially Relevant CAS RNs for Octylphenol Ethoxylates (Source: APERC) CAS RN Description 9002-93-1 Polyethylene glycol octylphenol ether 9036-19-5 Ethoxylated octylphenol 68987-90-6 Poly (oxy-1, 2-ethanediyl), alpha - (octylphenyl)- omega -hydroxy-,

branched

Environment Canada Proposed PBiT Categor izations for NP, OP, and the Ethoxylates

Table 1 (See Appendix) presents Environment Canada ’s currently proposed categorizations for the various CAS RNs that represent NP, NPEs, OP, and O PEs and illustrates discrepancies in decisions for equivalent or similar compounds. The following sec tions will provide a summary of the additional available data that will support c onsistent categorizations for these four classes of alkylphenol compounds as NOT P and NOT B under the CSDSL categorization scheme.

IV. PERSISTENCE ASSESSMENT

The environmental persistence of members of the family of NP, OP, and their ethoxylates has been extensively studied using a variety of test systems, ranging from screening tests (ready and inherent biodegradability) to simulation tests (river die-away, soil degradation). Based on the results of multimedia modeling, the relevant compartments fo r persistence assessment are water and soil. NP, OP, and their ethoxylates are not expected to be present in air due to their extremely low vapor pressures (0.05 to 0.2 Pa for NP and OP, <0.001 Pa for APEs) (EU, 2001; UKEA, 2005; Huntsman 1998a,b and 1999a-d) and Henry's Law coefficients (0.5 to 11 Pa m3/mol) (EU, 2001; UKEA, 2005).

Biodegradation is the dominant mechanism responsib le for removal of NP, OP, and their ethoxylates from aquatic and terrestrial environments. Hydrolysis is not expected to be impor tant, which is consistent with their chemical structures and lack of functional groups susceptible to hydrolytic attack.

Although the amounts of NP, OP, and their ethoxylates in air would be expected to be small, atmospheric photo-oxidation half-lives for NP and OP were calculated using EPISuite v3.12 (US EPA, 2000, AOPWIN v1.66) assuming a unique highly branched structure for each compound. Calculated vapor phase half-lives for NP and OP were 5 and 6 hours, r espectively. These values are below the criterion for persistence in air, and suggest that sunlight mediated degradation in the atmosphere might play a role in

the removal of NP and OP from the environment.

Environment Canada Proposed Persistence Categorization of NP, OP, and Ethoxylates

Environment Canada has categorized several members of the family of NP, OP, and ethoxylates as persistent based on experimental data. The rationa le for the proposed categorizations is as follows:

### 9 EC Proposed Categorization for NP/NPEs and OP/OPEs CAS RNs CAS Number Name Proposed

Classification Evidence Cited

25154-52-3 Phenol, Nonyl P-Exp CITI, 1992

9016-45-9 Poly (oxy-1,2- ethandiyl), alpha

- (nonylphenyl)- omega

hydroxyP-Exp (Preliminary) No reference provided in database

1806-26- 4 Phenol, 4-octyl P-Exp CITI, 1992 9002-93-1 Polyethylene glycol octylphenol ether P-Exp (Preliminary) No reference provided in database 9036-19-5 Ethoxylated octylphenol P-Exp (Preliminary) No reference provided in database

Additional synonyms for NPE (CAS RNs 26027-38- 3, 37205-87-1, 68412-54-4, 127087-87-0) were also categorized as persistent based on a grouping approach (analogy with CAS RN 9016-45-9). Nonylphenol CAS RNs 104-40-5, 11066-49-2, and 84852-15-3 and octylphenol CAS RNs 140-66-9 and 27193-28-8 were categorized as not-P on the basis of QSAR models.

Biodegradation Research

The biodegradation of NP, OP, and their ethoxylat es as related to their fate and lifetime in the

environment is well understood. There is extensive published scientific literature available that supports this understanding, ranging from screening tests, wh ich assess the ready and inherent biodegradability according to standard guidelines (e.g., OECD, EPA, etc) , to simulation tests for surface waters, soils, and wastewater treatment systems, which are designed to examine biodegradation in the laboratory under conditions that closely simulate the actual fate in the environment. Further, microorganisms with the ability to utilize NP, OP, and their ethoxylates as a carbon source for growth have been isolated and described (Maki et al., 1994; John and White, 1998; Soares et al., 2003).

The biodegradability of alkylphenols and their ethoxylates has been studied by researchers for over 40 years. Talmage (1994) reviewed the extensive intern ational literature on the e nvironmental health, fate

and environmental concentrations of alkylphenol ethoxylates (APEs) and summarized the more than 50

studies that examined the fate of NPE4-35 and O PE6-10. The methods used in these studies involved shake flask, die-away, or respirometer tests using various types of microbial inocula. Primary degradation was based on changes observed in the surfactant concentration as measured by various colorimetric tests, changes in physical properties, consumption of oxygen, and spectroscopy. While primitive by today’s methods, these early studies showed that the NPE and OPE surfactants easily underwent primary biodegradation. Although these early studies focused largely on elimination of parent material, recent improvements in analytical methods, use of radiotracers, and refinement and standardization of test methods have improved our understanding of the ultimate biological fate of NP, OP and their ethoxylates in the environment.

Based on the extensive research conducted to date, the mechanism and pathway for biodegradation of

APEs in the environment is well understood. Furt her, knowledge of the pathway is important for assessing the overall fate and lifetime of the family of NP, OP, and ethoxylates in the environment. As illustrated in Figure 1 (See Appendix) the mechanis m of APE degradation proceeds initially by the sequential removal of ethoxylate units. While NP is a possible metabolite of NPE (or OP from OPE), studies have shown that only minor amounts of these alkylphenols are formed under aerobic conditions. In contrast, alkylphenol carboxylates (APECs) are ofte n seen as dominant intermediates. The degradation pathways of APEs are reviewed in Ahel et al. (1994a), DiCorcia et al. (1998) and Maguire (1999).

Studies have also shown that under aerobic conditi ons, the aromatic ring of the alkylphenol and alkylphenol ether carboxylates is opened and the inte rmediate metabolites are further degraded to carbon dioxide (Naylor et al., 2005). Since an understanding of the biodegradability of APECs is important in

### 10 assessing environmental lifetime of NP, OP, and the ethoxy lates, the available data on the biodegradation of APECs will also be presented.

It is important to note that the Canadian DSL incl udes CAS RNs for a variety of NP, OP, and ethoxylated substances (See Table 1; Appendix). Some of these CAS RNs are descriptive of the commercially relevant branched alkyl chain materials, while othe rs are considered descriptive of linear alkyl chain isomers. As will be discussed below, adequate information is available to assess the environmental lifetimes (i.e., persistence) of both the linear and bran ched isomers for each of the substances on the DSL. That said, not all substances (i.e., CAS RNs) have been tested in each of the various laboratory systems. For example, although the ready biodegradability of al l linear alkylphenol isomers has not been examined, there is information available on the biodegradation of these materials in simulation tests (e.g., river die away). In general, the majority of the studies, including both screening and simulation tests, have focused on the commercially relevant branched chain isomers. While there is adequate information to support decisions on the categorization of the linear isomers, it is important to recognize that the available data on the branched isomers is also relevant for assessing the environmental fate of the linear materials. Based on extensive research leading to the development of f undamental principles of microbial degradation, it is clear that the linear isomers would be expected to biodegrade faster than the branched alkyl chain isomers. Further, available simulation test data for NPEs is relevant for assessing the environmental persistence of OPEs, as the initial degradation pr ocesses (i.e., sequential removal of ethoxylate groups) are identical. Thus, the information available for a ll CAS numbers should be considered in the decision, and applied to the entire family of NP, OP, and their ethoxylates.

The following sections will review the available inform ation on the biodegradation of the family of NP, OP, and their ethoxylates. Information will be pr esented on the results of laboratory screening tests,

simulation tests and field studies with the various materials.

Ready Biodegradation Tests

Ready biodegradation tests are stringent methods that use relatively low concentrations of both microbial

inoculum and substance concentration. The results of these tests are commonly used for screening or classification purposes. The test methods measure biode gradation by consumption of oxygen, removal of test substance (i.e., directly measure substance loss), removal of dissolved organic carbon, or by formation of carbon dioxide. According to Envi ronment Canada’s guidance (EC, 2003), substances having >60% biodegradation based on theoretical CO 2 (ThCO 2 ) or oxygen demand (ThOD), or >70% based on dissolved organic carbon (DOC) removal (the di fferent tests have different criteria) within 28 days are said to be readily biodegradable. Given th e stringent test conditions, a substance having <60% or <70% biodegradation is not necessarily persistent, but is said to have slower biodegradation. Inherent biodegradation tests must achieve ≥ 70% degradation to pass.

The Guidance for Categorization of Organic and Inor ganic Substances on the Domestic Substance List, Section 2.3.5.1 and Table 6 (EC, 2003) also provides information on how to extrapolate the results of ready biodegradability tests and estimate environmenta l half-lives. The degradation half-lives from the guidance are summarized below. This guidance will be used to convert results of screening tests for NP, OP, and their ethoxylates for comparison with the persistence criteria.

Guidelines: Extrapolation of Ready Biodegradation Test Results

to Estimate Environmental Half-Lives Ready Test Result % Biodegradation at 28 days Water Half-life (days) Soil Half-life (days) Sediment Half-life (days) >60% (ThCO 2 , ThOD) or >70 % (DOC removal) 5 5 20

<60% or <70%, but ≥ 40% 10 10 40

≥ 20% and <40% with >70% inherent degradability 30 30 120

<20% 10,000 10,000 10,000

### 11 Staples et al. (1999, 2001) reported the results of ready biodegradation tests conducted on NPE9 (CAS RN 9016-45-9), OPE9 (CAS RN 9036-19-5), and several of their biodegradation intermediates, including NPE1.5, OPE1.5, NP (CAS RN 84852-15-3) and OP (C AS RN 140-66-9). Further, to provide a more comprehensive picture on the overall fate of the al kylphenol ethoxylates, the ready biodegradability of important biodegradation intermediates, including NPEC1, NPEC2, OPEC1, and OPEC2 were also investigated. Biodegradation was examined accord ing to standard guidelines, namely OECD method 301B (modified Sturm test) that measures the ultimate biodegradation of the substances by quantifying the formation of carbon dioxide, OECD 301F (manometric respirometry), and a draft ISO method. All of the studies were conducted in accordance with Good Laboratory Practices.

Results of these ready biodegradability tests are su mmarized in Table 2 (See Appendix). Based on CO 2 production observed in the OECD 301B test, NPE9 a nd OPE9 were biodegraded by 70 to 83%, NPE1.5 and OPE1.5 by 59 to 65%, and NP and OP by 48 to 70%. Further, biodegradation of each of the ether carboxylates (NPEC1, NPEC2, OPEC1, and O PEC2) exceeded 60% of the theoretical CO 2 formation by day 28. Results of the OECD 301F test also confir med that NP was extensively degraded, with 62% of the theoretical oxygen demand consumed after 28 da ys. The results also suggested that OP may biodegrade faster than NP. For example, in the OECD 301B test, NP was 48% biodegraded (ThCO 2 ) by day 28, while OP was degraded by 69.9% on day 20.

In addition to quantifying ultimate biodegradation by measuring CO 2 production, primary biodegradation was quantified in the OECD 301B tests. Dissolv ed organic carbon (DOC), suspended organic carbon (SOC), as well as concentrations of APE, AP, and A PEC remaining at days 15 and 35 were measured. Based on DOC removal, all compounds tested were degr aded by 87.1 to 97.6%. Conversion of the test material into cellular biomass or sorbed onto partic ulates was measured as SOC, and comprised 15.1% to

48.4% of the initial carbon. Concentrations of AP, APE1, APE2, and APE3-17 were measured using GC/MS. For each test, concentrations measured on days 15 and 35 were summed and compared to total concentrations on day 0. By day 15, based on the re maining residues, degradation of the various parent materials tested ranged from 91.1 to ~100%. By da y 35, the amount of degradation ranged from 92.3 to

~100% of initial concentrations across all tests.

Hughes et al. (1989) compared the ready biodegradab ility of NPE12 in three standard test systems, including the modified Sturm test (OECD 301B), the Gledhill test (EPA method 835.3120) and the closed bottle test. Ultimate biodegradation in each of the tests as measured by conversion to CO 2 ranged from 30 to 65%. Using the Gledhill test, the authors examined the effect of using both acclimated and unacclimated microbial seed on biodegradation of NPE12 to CO 2 . No significant differences were noted,

as mineralization of the test substance reached 45% ThCO 2 with the unacclimated seed as compared to

42% ThCO 2 in tests with an acclimated inoculum.

Markarian et al. (1989) also compared the biodegrad ability of NPE7 in the closed bottle test and the Gledhill test. The authors reported 60% ThOD and 40% ThCO 2 for NPE7 using unacclimated inocula. While the test procedures used by Hughes et al. ( 1989) and Markarian et al. (1989) were not exactly identical to the OECD 301B used by in Stapl es et al. (1999, 2001), the results from each of these screening tests are relatively consistent. Slight differe nces in the results of the studies are likely attributed

to different sources of microbial inocula.

Some studies have shown that NP and OP do not pass certain screening tests. For example, results of studies conducted using the OECD 301C ready biodegradab ility test as reported by the Japan MITI (CITI, 1992) were selected as the pivot value by Environment Ca nada for the categorization of the persistence of NP (CAS RN 25154-52-3) and OP (CAS RN 1806-26-4). No biodegradation was observed for either test material in the 301C test (CITI, 1992). In additi on, results of studies conducted using OECD 302C that showed no degradation were reported in the EU Ri sk Assessment Report for OP (CAS RN 1806-26-4). It is important to note that discrepancies are often seen between results of OECD 301C or 302C tests (typically used by MITI) and other ready biodegradability tests (e.g., OECD 301B, 301F, analogous US EPA guidelines). The reason for these discrepancies is thought to be due to differences in the microbial inoculum used in the studies. The majority of r eady biodegradability tests conducted according to current

### 12 guidelines (OECD, US EPA) use freshly collected inoculum from municipal wastewater treatment plants. In contrast, studies conducted according to OECD 301C and 302C use a mixed inoculum that has been pre-conditioned for 30 days on a medium containing glucose/peptone as the sole carbon source. Such culturing of the inoculum on easily degraded substrat es is considered to result in substantial loss of microbial diversity originally present in the composite sample, and thus the environmental relevance of the preconditioned inoculum has been criticized. Recent studies by Liu et al. (1997) and Forney et al. (2001) using genetic techniques (DNA molecular probes) have shown that the microbial diversity of the inoculum significantly decreases during this 30 day pr e-conditioning period. Because of the unique and un-natural inoculum used in OECD 301C and 302C tests, the results should not be used for categorization of the persistence of NP and OP. Further, the results of OECD 301C tests are inconsistent with the results of other screening tests conducted using standa rd guidelines and conducted under GLP, as well as

simulation tests with these materials.

In summary, ready and inherent biodegradation studies that adhered to OECD protocols or used very similar protocols have been conducted with branch ed NPE, OPE, NP, and OP. Based on the guidance provided by Environment Canada for the categoriza tion of substances on the DSL (EC, 2003), the results

of these tests were used to derive environmental ha lf-lives. As summarized in Table 2, none of the branched NPE (1.5, 7, 9, 12 mole ethoxylate), OPE (1 .5, 9 mole ethoxylate), NP, and OP materials exceed the criteria for persistence in water, soil or sedimen t. Further, because the linear alkylphenol isomers would be expected to degrade at least as fast if not f aster than the branched isomers, the linear NP and OP substances on the DSL would also be expected to exhibit similar if not shorter half-lives in the environment. This statement is supported by the results of simulation tests as discussed below.

In conclusion, based on the results of stringent screening tests, none of the NPE, OPE, NP or OP substances listed on the DSL (as summarized in Table 1) can be considered as meeting the criteria for persistence.

Simulation Tests with Linear Test Materials

In addition to screening tests, the biodegradability of members of the family of NP, OP, and their ethoxylates have been examined in a variety of simu lation tests for surface waters and soils. Such tests are designed to examine degradation in the laboratory under conditions that closely simulate the fate of the compound in the environment.

The biodegradation of linear NP has been measured in studies using seawater and marine sediment, soil and sludge-amended soils. As will be discussed below, the results support the conclusion that linear NP can be expected to biodegrade at least as fast as the branched chain materials.

Ying and Kookana (2003) reported the r esults of static and aerated die-away tests with linear 4-n-NP in seawater and marine sediment. In seawater, 4-n-NP w as nearly completely biodegraded by 90% within 7 days under constant aeration. In sterile contro l vessels, concentrations of 4-n-NP decreased by approximately 50%. The authors attributed the losses to abiotic processes such as volatilization or sorption to vessel walls. In a follow-up test without aeration, 4-n-NP was biodegraded by >90% during the first 7 days and was eliminated in about 50 days. Elimination was attributed to biodegradation. In marine sediment under oxic conditions, 4-n-NP was degrad ed by 42% in day 1 and 95% by day 21. From these results, the authors estimated half-lives of 5 da ys for seawater and 5.8 days for marine sediment

under oxic conditions.

Ying and Kookana (2003) also reported the results of parallel tests using branched OP. In seawater without aeration, OP was biodegraded by >99% by day 42. Elimination was attributed to biodegradation. In marine sediment, OP was removed by only 16% dur ing the first 21 days (acclimation period), and then rapidly degraded with 91% removal by day 28. Fr om these data, the authors estimated half-lives for branched OP of about 20 days for both seawater and marine sediment under oxic conditions. These results suggest that in seawater and marine sedimen t, 4-n-NP degrades may be more rapidly and more extensively than the branched OP. However, it is im portant to note that both the linear NP and branched OP were easily biodegraded in samples from the marine environment.

### 13 Mortensen and Kure (2003) measured the biodegradation of linear 4-n-NP spiked into a sandy loam soil. The spiked soil was either unplanted or planted with rape . The concentration of the test material at study

initiation was 451 μ g/kg, dry wt. By day 30, concentrations of 4-n-NP in both planted and unplanted soils were reduced by 97.8% and 98.2%, respectively. Assu ming the degradation proceeded in a linear fashion, half-lives for biodegradation of 4-n-NP in soil were approximately 15 days.

Dubroca et al. (2005) measured the degradation of a mixture of linear and branched 14 -C-U-NP in an agricultural soil. By day 8, about 40% of initial radi olabeled test material was bound to the soil. By day 64, about 30% of initial radiolabel was collected as carbon dioxide. The remaining approximately 30% of

initial 14 C-NP was present as extractable material. H PLC analyses showed that NP was present at negligible concentrations by day 32. From these da ta, the authors calculated biodegradation half-lives of 4 days for the mix of linear and branched NP.

To examine differences between the biodegradation ra te for linear and branched NP, Dubroca et al. (2005) measured the degradation of linear 4-n-NP and branched NP separately in fungal cultures. Some of the fungi were isolated from two different sludge amended soils, while others were from the author’s culture collection. Results of the study showed that in general, linear NP degraded two to four times

faster than the branched test material.

Kirchman et al. (1991) studied the biodegradation of linear 4-n-NP. The test chemical was added to soil at concentrations of 10 or 500 mg/ kg and incubated in sealed flasks for three months. Degradation was monitored by analysis for the parent compound and also by CO 2 evolution. Based on the analysis of parent compound remaining in the soil in both treatme nts, greater than 90% of the added nonylphenol remained after 10 days incubation; the levels were reduced to below the detection limit (<0.02 mg/kg) after 20 days incubation. In st udies conducted with 500 mg/kg NP, CO 2 evolution was higher than observed in controls, and indicated that about 61% of the parent compound was converted to CO 2 after 94 days incubation. In studies conducted with 10 mg/kg of test chemical, the CO 2 evolution was similar to controls, suggesting that the lower concentration is below the capability of the method to differentiate mineralization of the test chemical vs. the background soil metabolism. Based on their results, primary biodegradation half-lives of 4-n-NP in soil of 2.5 to 5.0 days were calculated for the 10 mg/kg and 500 mg/kg treatments, respectively.

The laboratory simulation data for linear NP shows that it is extensively biodegraded in seawater, marine sediment, and soil. Half-lives ranged from 5 to 5.5 da ys for linear NP in seawater and marine sediment, and were faster than rates measured for branched OP (h alf-life of 20 days). In soil, 4-n-NP and a mixture of linear and branched NP yielded primary biodegrada tion half-lives of approxima tely 15 days and 4 days, respectively. These confirm that studies with the mo re commonly tested branched NP and OP can be used to conservatively predict the degradability of linear NP and OP. Since the ready test data for branched NP (CAS RN 84852-15-3) and OP (CAS RN 140-66-9) do not meet the Environment Canada’s criteria for persistence, both linear NP (CAS RN 25154-52-3) and linear OP (CAS RN 1806-26-4) also do not meet the criteria for persistence.

Simulation Tests with Branched Test Materials

Die-Away Studies with Freshwater and Freshwater Sediments

Dutka et al. (1998) conducted shake flask tests with NP in freshwater and sediment-containing systems

under either oxic or anoxic conditions. Sediments were taken from a lake near a creek from which

wastewater treatment effluents flow. The sedimen ts contain background concentrations of NP (~84 μ g/g dry weight). One experiment measured the degrad ation of the background NP over time, while a second experiment measured NP concentrations over time in sediment spiked with NP (to a total of about 500

μ g/g NP). Sediment concentrations of NP were m easured over the course of 6 months. NP in unspiked sediment decreased by 89.5% in four weeks, but did not degrade further. The authors speculated that additional NP was being formed in the sediment duri ng the test from degradation of other NPE residues

### 14 present. NP in spiked sediment increased in con centration over the first three months, followed by degradation of 62.5% by the 6 month. It is likel y that the apparent increasing concentrations merely represent non-homogeneity of the spiked NP in the sediments. Similar experiments were conducted

under anoxic conditions. No decrease of NP was noted under anoxic conditions.

Yoshimura (1986) conducted die-away studies of NPE9 using river water and sediments. Studies with sediments (3000 mg sediment /L) used a synthetic test medium. River die-away tests used only river water, but no sediment. Over the course of the 30 day experiments, samples of test media (synthetic media or river water) were collected periodically a nd analyzed for NPE using HPLC. Test vessels were either stirred or static during the test. Primary de gradation with and without sediments was 97 to 99% by 10 days in static vessels and in about 5 days in s tirred vessels. NPE continued to degrade during the remainder of the study (to 30 days). In the studies with sediment present, over 99% of the remaining levels of NPE were in the water phase with the balance adsorbed to the sediment.

Manzano et al. (1998) reported the results of river die-away studies conducted using NPE15 at concentrations ranging from 2.5 to 10 mg/L. Afte r a one to two day lag phase, parent NPE15 was reduced (primary biodegradation) by 85% by day 5 at all concentrations. No further degradation occurred after day 5 through the end of the study at day 30. NPE2 was reported to be the terminal oligomer formed, which in turn degraded to NPEC2, then NPEC1. The authors calculated the apparent extent of mineralization by assuming that the lost EO units (as ethylene glycol) are rapidly mineralized to CO 2

. The extent of mineralization was calculated to range from 64 to 85% across all experiments.

Maki et al. (1996) measured the degradation of NPE9 in tests with standard BOD media inoculated with the filtrate of river water samples. The degradati on of the NPE oligomers was essentially complete since

NPEs were not detectable. The main intermediate degradation product detected was NPEC1 and minor amounts of NPE2. NPE degradation was slower in test systems using river water collected upstream of

sources of NPE.

Sundaram and Szeto (1981) measured the degradation of NP in freshwater-sediment mixtures. Some of the NP added to the systems became adsorbed to the sediment during th e course of the study. However, by day 71, 80% had been biodegraded, while 20% remained incorporated in the sediment.

Yuan et al. (2004) conducted static die-away tests on NP and NPE1 with river sediments. A portion of

the sediments was acclimated to NP. Separate studies were conducted at 30 ° C with unacclimated sediment, acclimated sediment, at various temperatures (20 to 50 ° C), and with or without shaking. For static systems at 30 ° C, the authors reported half-lives across four sediments of 13.6 to 99.0 days for NP and 69.3 to 115.5 days for NPE1. In static systems also at 30 ° C using both unacclimated and acclimated sediment, half-lives for NP were 20.4 and 5.1 days, respectively. The time to disappearance was 70 days for unacclimated sediment and about 27 days for acclim ated sediment. In the same systems with both unacclimated and acclimated sediment, half-lives for NPE1 were 23.1 and 5.7 days, respectively. The time to disappearance was about 85 days for unacclimated sediment and about 55 days for acclimated sediment. Temperature was a significant factor in c ontrolling the degradation rate of both NP and NPE1. For NP, half-lives of 40.8, 4.2, and 3.0 days were reported for 20 ° C, 40 ° C, and 50 ° C. For NPE1, halflives of 57.8, 6.0, and 4.7 days were reported for 20 ° C, 40 ° C, and 50 ° C. An additional important factor

in the conduct of these tests was whether or not the test vessels were shaken or static. Shaking the flasks decreased the half-lives by about a factor of two to five. Absence of shaking likely helped create anoxic conditions in the sediments; NP and NPE are known to degrade slower or not at all under anoxic conditions (Dutka et al., 1998).

Die-Away Studies with Seawater and Marine Sediments

Kvestak and Ahel (1995) reported the results of static die-away tests of NPE10 in estuarine water with varying salinities and that was collected in spring, su mmer, and winter. The authors used the degradation data (total NPE concentrations) collected during the study to determine lag phases prior to the onset of rapid degradation and to estimate pseudo-first order half-lives. Lag phases were <1 day at 20 to 22.5 ° C, 4

### 15 to 13 days at 18 ° C, and 3 to 7 days at 13 ° C, indicating that the duration of the lag phase is somewhat a function of temperature. The duration of the lag ph ase was not affected by salinity. Biodegradation halflives at 20 to 22.5 ° C were 2.5 to 35 days, at 18 ° C were 10 to 35 days, and at 13 ° C were 23 to 69 days. The longest half-lives (35 to 69 days) were all conduc ted at an NPE10 concentration of 1 mg/L (except for a single result at 13 ° C), which led the investigators to suggest that the NPE10 may have been inhibitory to the micro-organisms. During degradation, the oligomer distribution was first reduced but did not shift, then subsequently shifted to lowe r oligomers forming mainly NPE2. No NPE>5 was detected. The analytical method did not analyze for NPEC.

Ekelund et al. (1993) reported the results of shake flask die-away tests of synthesized 14 C-NP in seawater

or a blend of seawater and marine sediment conducted at 11 ° C. With seawater only, the authors reported a 28 day lag phase followed by formation of CO 2 (55% by day 56). With seawater and sediment, no lag phase was observed and formation of CO 2 started immediately reaching 44% by day 56. The overall mass balance of this study was <50%, so it is po ssible that additional mineralization occurred but CO 2 was lost. However, it is more likely that unextractable metabolites were formed or that incorporation into cellular material occurred (Naylor et al., 2005).

Die-away Studies with Soil and Sludge Amended Soils

Marcomini et al. (1988, 1989) reported the results of di e-away studies of NP and NPE1,2 in soil to which anaerobically digested sludge was added. NPE1,2 a nd NP were reduced by 80% in the first 21 days. Further reduction was negligible due to sorption of th e metabolites to soil and their subsequent lack of bioavailability.

Kubiak (2002) conducted laboratory lysimeter studies us ing two soil types with and without plants and monitored the leaching and degradation behavior of 14 C-NPE3. Dissipation half-lives ranged from <1.5 to <6 days. Lower 14 C-NPE and 14 C-NP were formed as transient products that were quickly degraded to

14 CO 2 (the actual rates were not provided in the text). Up to 20% of the applied radiolabel consisted of

bound residue.

Topp and Starratt (2000) examined the mineralization of ring-U- 14 C-labeled NP in several soils as a function of temperature. 14 C-NP fate was assessed in microcosms incubated at either 30 ° C or a range of temperatures from 4 to 30 ° C. In all soils, an initial rapid phase of degradation occurred (30% loss of applied radioactivity in first 10 days) followed by a slower phase. The authors calculated mineralization half-lives ranging from 4.5 to 16.7 days based on th e more rapid initial phase. Degradation was a function of temperature (tested at 5 to 30 ° C) with increasing lag times and slower overall degradation at lower temperatures (half-lives not reported). Most re maining activity was not extractable (50% of initial activity). The authors also showed that the moisture c ontent of the soil affected the rate of degradation. Saturated soil and nearly dry soil had significantly lo wer capability to mineralize the NP (10 to 25% of that achieved with moist but not saturated soil). An additional test was conducted in which soils were amended with municipal sewage sludge from an aerobic treatment plant to assess the effect of the sludge on the bioavailability of NP for mineralization. The authors reported that the sludge-amended soil degraded NP rapidly, but was controlled by oxy gen availability. Heavily amended soil may limit degradation of the NP due to the high oxygen demand of the sludge.

Kuchler (1996) conducted mineralizati on and sorption studies of NPE9 in soil lysimeters. Neither NPE nor NP migrated beyond the first 10 cm depth of the soil and no NPE or NP was detected in soil or leachate after three to four weeks. Removal of all applied NPE was 99.8%.

Hughes et al. (1998) measured the fate of NPE9 used as a commercial adjuvant. Theoretical CO 2 production from NPE9 reached 50% in 50 days and 57% in 64 days. HPLC analyses showed that the detection of the aromatic ring decreased by 69 to 70% by day 21, by 82 to 83% by day 42, and disappeared completely by day 63. The authors conc luded that NPE9 degrades ultimately and completely

in soil.

### 16 Mortensen and Kure (2003) reported the results of die- away tests of NP in soils to which sewage sludge and other organic wastes were amended. Amended soils were either planted with rape or not planted. Soil amended with anaerobic digester sludge that was planted with rape had less residual NP at 30 days than amended soil without plants (13 vs. 26%). Similar results were found with soil amended with activated sludge (8.3% in planted amended soil vs. 18% without plants). NP in soil amended with activated sludge or compost was reduced by 84% and 64%, respectively, by day 30.

Dettenmaier et al. (2004) conducted similar tests ex amining the fate of NPE9, NPE4, and NP in soil amended with biosolids and which were either unplante d or planted with crested wheat grass. The extent of degradation was not different between planted and unplanted microcosms. Mineralization ranged from 7.2 to 11% for NP, 17 to 24% for NPE4, and 21 to 28% for NPE9 by day 150.

Jacobsen et al. (2004) conducted lysimeters studies using a sandy loam soil into which anaerobically digested sewage sludge (that contained NP) was incorpor ated to a depth of 15 cm. After incorporation,

the NP concentration was 0.56 μ g/g, dry weight. Lysimeters were placed outdoors to receive natural lighting, temperature, and precipitation. Lysimeters we re also irrigated with a constant volume of water. Samples of leachate and soil from three layers were collected periodically during the 110 day experiment and analyzed for NP. The authors reported rapid die-aw ay of NP in the first 10 days (55% reduction). A

slower but consistent reduction o ccurred throughout the remainder of the experiment (to 110 days). The half-life of the degradation of NP over da ys 10 to 110 was calculated to be 37 days.

Field Studies

Heinis et al. (1999) reported the results of a field study in which NP was added every two days over a 20 day period to an artificial enclosure of a littoral zone in a small lake. Samples of water, sediment, and plant tissue were periodically measured for NP cont ent over the course of 440 days. Dissipation halflives (includes degradation and sorption) averaged 0. 74 day in surface water, 66 days in sediments, and 10 days in plants. The authors did not distingui sh between amounts of water-borne NP that was biodegraded and that adsorbed to sediments, macrophyt es or soil. However, the authors did calculate an approximate mass balance of all applied NP for two of the enclosures. Initially all NP was in the water column. By day two after application, approximately 10% was sorbed to enclosure walls (8.3 to 8.7%), macrophytes (1.2 to 1.4%), and sediment (0.7%). The rate of dissipation from the water column increased during the 20 day dosing period so proportiona lly the amounts sorbed to the walls, plants, and sediment increased. However, the absolute amounts continued to decrease. Following the dosing period of 20 days, the amount of applied NP that was sorbed to the sediment never exceeded about 1%. It was

assumed that the sorbed NP to the en closure walls desorbed back into th e water over time. NP sorbed to the plants either degraded or d esorbed to the water. The fact that only relatively minor amounts of applied NP were present in the sediment at any time suggests that biodegradation was the dominant removal process in the enclosures. Thus the dissipati on half-life of NP in the water of 0.74 days is effectively a biodegradation half-life.

Jonkers et al. (2003) reported on bi odegradation of NPE and their degradation intermediates in field studies of estuarine sediment. Samples were collected in two estuaries, the first was stratified with a short retention time (1 to 3 days) and the second with a rete ntion time of two to three months. Sources of NPE were wastewater treatment plants upstream of the estua ries. Despite the relatively short retention time in the first estuary, NPE steadily decreased forming mainly NPEC and negligible NP. Due to stratification of salinity and the relatively short retention time of water flowing in the estuary, the authors could not conclude that biodegradation occurred. However, in the second estuary, NPE steadily decreased with distance from the source with NPEC>3, NPE, and NP decreasing faster than the formation and subsequent disappearance of NPEC1,2. No CAPEC were detected. The authors conclude that these data show evidence of biodegradation in the field and support laboratory results as well (Jonkers et al., 2001).

Kuchler et al. (1996) examined the fate of NPE in so ils in field trials. During a one year period, 10 soil plots were treated with two different sewage sludg es that contained residues of NPE and NP. After application, the sewage sludge was incorporated into th e top 5 cm of the soil. Soil samples were collected

### 17 from various depths (0-10 cm, 10-20 cm and 20-30 cm ) and analyzed for the presence of NPE and NP. NPE concentrations rapidly decreased, with no compound being detected after 20 days. No leaching of NPE or NP was seen from the top (0-10 cm layer) indicating that removal was by biodegradation. Although NP initially increased in concentration during the first 10 days, which indicates that it was possibly formed from the degradation of nonylphenol ethoxylate, but no nonylphenol was detected after 20 days, indicating that it was degraded.

Gross et al. (2004) traced the fate of APE1-3, AP, and combined APEC1-3 and CAPEC1-3 in a river and wetlands. The substances originated from four wastew ater treatment plants located on the river or on creeks that discharged into the wetlands. The river flowed into the wetlands and the outflow of the wetlands returned to the river. Water samples were taken immediately after the treatment plant outfall

and 10 km downstream. The authors calculated that median concentrations of AP, APE1-3, and APEC1- 3 combined with carboxylated APEC1-3 decreased by 92%, 95%, and 85%, respectively. The median concentration of all analyses combined decreased by 86%. Between wetland inflow and outflow, median concentrations of APs declined by 75%, however, combined APEC1-3+CAPEC1-3 decreased only by 8%.

APEs were not detected in either inflow or outflow . The median concentration of all analytes combined decreased by 11%. In the samples taken immediately downstream of the discharge, APEC1-3+CAPEC1- 3 comprised 83% of all analytes, while 10 km downstream they comprised 99% of all analytes. Median

concentrations of CAPEC1-3 were approximately two to seven times that of the APEC1-3. At the outflow to the wetlands, nearly all of the subs tances are as APEC1-3+CAPEC1-3, which are acidic substances. The authors suggested that both the ri ver and the wetland were highly effective at removing AP and APE and that the river effectively remove d APEC+CAPEC, but that the wetland treatment was poor at attenuating APEC and CAPEC.

Huntsman et al. (2005) examined the fate of NPE9 (commercial detergent product) discharged into an onsite wastewater disposal (septic) system. NPE9 based detergents were metered daily into the plumbing at a single-family household. The ethoxylate-containing wastewater was discharged into the highly anoxic environment of a 4500 L septic tank prior to distribution into the oxic subsurface via 100 meters of leach line. Soil pore water and ground water samples were co llected and analyzed for NPE, NPEC, and NP. The NPE9 and the degradation intermediates that were measured were reduced by 99.99% on a molar

basis. An 18% reduction in molar concentration with in the septic tank was observed. This was followed by a further 96.7% reduction of molar concentration w ithin the leach lines. As the pore water migrated through the vadose zone, an additi onal 99.69% reduction in molar con centration was measured between the bottom of the leach lines (leach line effluent) and the lowest vadose zone monitoring location. Only trace amounts of degradation residuals were detected in soil pore water. The results indicate that degradation of the surfactant occurs within the anoxi c portion of the disposal system followed by rapid biodegradation in the oxic unsaturated zone. These results show that NPE rapidly and completely degrades in on-site wastewater disposal (septic) systems.

Persistence - Summary and Conclusions

The 1992 CITI studies conducted for Japan MITI, which were cited by Environment Canada to support a proposed categorization of Persistent (P-Exp) fo r Nonylphenol (CAS RN 25154-52-3) and Octylphenol (CAS RN 1806-26-4) (CITI, 1992), are inconsistent with the weight of the evidence provided by numerous laboratory and field studies on these and rela ted compounds. Discrepancies, thought to be due to differences in microbial inoculum, are often seen between results of OECD 301C or 302C tests (typically used by MITI) and other ready biodegradability tests (e.g., OECD 301B, 301F, analogous US EPA guidelines).

Furthermore, biodegradation - the dominant remova l process of alkylphenols and their ethoxylate from water, sediment, and soil - has been extensively st udied and several dozen studies have been conducted that focus on both the primary and ultimate degradation these compounds. Three main types of biodegradation studies have been performed. The fi rst are stringent screening tests that measured the ready biodegradability of a test substance. Ready bi odegradation tests have been conducted for APE, AP, and APEC. The second type are simulation tests in wh ich biodegradation is measured in laboratory tests using procedures and bacterial inocula that simula te potential degradation in specific environmental

### 18 compartments or within wastewater treatment plan ts. Simulation tests for biodegradation of APE and their degradation intermediates are available for all relevant environmental compartments including freshwater, freshwater sediment, seawater, marine sedim ent, and soil. The third line of evidence that biodegradation is the key removal process for APE a nd their biodegradation intermediates is based on

results of field studies.

The ready biodegradability of NPE9, OPE9, NPE1.5 , OPE1.5, NP, and OP has been measured using standard OECD 28-day protocols under GLP. Th e Environment Canada Guidance (EC, 2003) gives

instructions on how to estimate environmental half-liv es based on the test results from ready tests. Using the experimental results for NPE (CAS RN 9016- 45-9), OPE (CAS RN 9036-19-5), NP (CAS RN 84852- 15-3), and OP (CAS RN 140-66-9) and the guidance, half-lives in water, soil, and sediment were determined. The results are presented in Table 2 (See Appendix) and show that none of the test substances meet the criteria for persistence.

Half-lives for biodegradation within specific envir onmental compartments were also taken from the

simulation studies that were reviewed above. Half-liv es were either calculated by the authors or were

calculated here if there were sufficient data to do so. The compiled biodegradation half-lives from studies that used the commercially relevant branched materi als are shown in Table 3 (See Appendix). Half-lives for NP in river water ranged from 2.5 to 40.8 days. River die-away half-lives for NPE ranged from 18.6 to 57.8 days. In seawater and marine sediments, die-aw ay half-lives were 5.9 and 26.3 days, respectively. NPE10 biodegradation under varying conditions of temperature and sa linity ranged from 2.5 to 35 days for 12 experiments and 69 days for one experiment . The latter test was conducted at the lower temperature (13 ° C and 38%, respectively) and highest salinity used and so some bias to slower degradation was noted. Biodegradation half-lives for OP of 20 days were reported in both seawater and marine sediment. In soil, biodegradation half-lives for NP ranged from 4.5 to 37 days. Degradation halflives for NPE ranged from 1.5 to 18.2 days in soil.

The laboratory simulation studies showing biodegrada tion half-lives averaging about 20 days across all media and all compounds are further supported by the results of the field studies that showed that alkylphenols and their ethoxylates are extensively bi odegraded in surface waters, sediments, and soil. Collectively, the data from the laboratory and field support the results of the ready biodegradation tests that showed that all NPE, OPE, NP, OP, NPEC, and OPEC are not persistent compounds.

In conclusion, the fate of alkylphenol ethoxylates a nd their degradation intermediates in the environment is well understood. Their physical properties and partiti oning characteristics are well defined. The fate of APE and their degradation intermediates has been studied in all compartments where they may be expected to be found. The database examining the fate of APE and their degradation intermediates in the environment includes stringent screening tests, laboratory simulation tests and field studies, which show conclusively that neither the linear research chemi cals nor the branched commercial APE products - or any of their degradation intermediates (lower mole A PE, AP and APEC) - can be categorized as persistent.

V. BIOACCUMULATION ASSESSMENT

Environment Canada has proposed various categoriza tions for the bioconcentration potential for members of the family of NP, OP, and their ethoxylates (See Table 1, Appendix). These decisions were based on a combination of the limited review of experimental data, QSAR modeling, and expert judgment. The purpose of this section is to provide additional information to inform the decision on categorization of NP, OP, and their ethoxylates.

According to the data preference outlined in the Guidance Manual for categorization of Organic and Inorganic Substances on Canada’s Domestic Substa nces List (EC, 2003), when bioaccumulation (BAF) or bioconcentration (BCF) data are available for di fferent classes of species (e.g., algae, water fleas, and fish), the BAF or BCF data for the fish species will be used to represent the bioaccumulation behavior of the chemical substances for the categorization process. The rationale for the preference of fish data over those for lower trophic levels is that the potential fo r chemical bioaccumulation in the food chain is best

### 19 measured by the BAF or BCF in higher trophic level orga nisms. Consequently, in view of the extensive information available, the following discussion will fo cus on the BAF and BCF potentials in fish only.

Nonylphenol

Environment Canada has categorized two NP substances as not meeting the criteria for B based on experimental data while another two CAS RNs for NP are categorized as meeting the criteria based on QSAR models. The rationale for the proposed categorizations is as follows:

Proposed Categorization for Nonylphenol CAS Numbers CAS RN Name Proposed Cla ssification Evidence Cited

104-40-5 p -Nonylphenol NOT B Exp. (Giesy et al., 2000) * 25154-52-3 Phenol, Nonyl NOT B Exp. (CITI, 1992) 11066-49-2 Isononylphenol B QSAR Model 84852-15-3 Phenol, 4-nonyl-,

branched B QSAR Model

* Note: Incorrect CAS RN assignment in Geisy et al. , 2000 is discussed in text below.

Gobas and Arnot (2003) reviewed some of the ava ilable data for bioaccumulation potential for NP. Several studies were rejected due to insufficien t duration of exposure periods (McLeese et al., 1980; McLeese et al., 1981; Ekelund et al., 1990). Further, BCF data for NP (CAS RN 25154-52-3) were initially rejected based on the assumption that the resu lts were based on nominal exposure concentrations (CITI, 1992). This decision has since been rever sed, since it has been clarified that the exposure

concentrations were measured.

As summarized in Tables 4a and 4b (See Appendix), a number of studies with fish are available to inform the decision on the BAF and BCF potential for the vari ous NP isomers. The subject has also been reviewed in several publications (Staples et al., 1998; Servos, 1999; EU, 2001). Note that some studies have specified the CAS RN of the test chemical, while other have only indicated whether the test material

was linear or branched NP. While several studies are judged “valid” or “reliable” for assessing BAF and BCF potential, the other studies, which provide results from shorter exposure periods or field investigations, are useful in contributing to the overall weight of evidence.

Further, since the uptake, partitioning, and metabolism behavior of linear and branched NP isomers would be expected to be essentially identical in fish, the information available for all CAS numbers should be considered in the decision, and applied to all members of the NP family for the purpose of categorization. This conclusion is supported by the available data for the BCF potentials of branched and linear NP, as

discussed below.

Laboratory studies

The bioconcentration of NP (technical grade, branched) in juvenile Atlantic salmon ( Salmo salar ) was studied by McLeese et al. (1981) with static expos ure studies conducted over a period of 4 days. Given the short exposure period and the sta tic exposure conditions, the study s hould be judged use with caution. The uptake rate constant was measured to be 45 mL-g -1 -day -1 and the excretion rate constant was 0.16 day -1 , giving a steady-state bioconcentration factor of around 280 mL/g wet weight. The excretion halflife was estimated to be about four days.

Ekelund et al. (1990) studied the bioaccumulation of 14 C-labeled NP in three-spined stickleback ( Gasterosteus aculeatus L.). The 14 C-labelled p-NP was synthesized from uniformly labeled phenol and unlabelled nonene. Analysis of the synthetic material indicated the isomer distribution was consistent with commercial NP. The animals were exposed to 4.8-4.9 micrograms/L 14 C-NP in a flow-through system. Exposure was for 16 days followed by an e limination period of 32 days. The authors found that steady state bioaccumulation had been reached by the end of the exposure period, and that NP was then

### 20 rapidly eliminated from the fish. Bioconcentration fact ors for the fish were in the range of 1,200 to 1,300 mg/L on a fresh weight basis. Subsequent kinetic an alysis of these data by Dow revealed an uptake rate constant (K 1 ) of 1370 mL-g -1 -day -1 and elimination rate constant (k 2 ) of 1.03 day -1 , for a kinetic BCF value of 1333 mL/g and an elimination half-life of approximately 0.7 day (Dow , 2005). This kinetic BCF value agrees well with the 1200-1300 mL/g estimate offere d for NP in stickleback by the authors. The

authors noted that the BCF values are based on total 14 C measurements in fish tissue, thus, the presence of metabolites in the organisms may have led to an ove restimate of the bioconcentration potential. Although Gobas and Arnot had previously rejected this study dur ing their review of the available information, the fact that steady state conditions were achieved, but th at BCF values were based on total fish radioactivity, supports classifying this study as use with caution.

The bioconcentration of NP (84852-15-3, co mmercial product) in the fathead minnow ( Pimephales promelas ) was investigated by Ward and Boeri (1991). The study was conducted according to US EPA test guidelines in accordance with Good Laboratory Prac tices, and is considered valid for the purposes of categorization. Fathead minnows (0.5-1 g) were exposed to separate nominal concentrations of 5 micrograms/L and 25 micrograms/L NP in an intermittent flow-through system for 20 days. The exposure period was then followed by a 7-day depuration period. The system was analyzed for NP concentrations, dissolved oxygen content, temperat ure, and pH. Measured levels of NP were 4.9 micrograms/L and 22.7 micrograms/L in the two test systems, respectively. The concentration of NP in fish tissues increased from background concentrations to steady state concentrations during the first 3-10 days of exposure, a finding that was consistent with previous bioconcentration studies with NP. Uptake and depuration of NP appeared to be independent of the concentration of the test substance in water. Exposure of fathead minnows to 4.9 micrograms/L NP in water for 20 days resulted in a BCF of 271 L/kg fresh weight with an uptake rate constant of 133 mL-g -1 -day -1 and a depuration rate constant of 0.49 day -1

. Exposure to 22.7 micrograms/L NP for 20 days resulted in a BCF of 344 L/kg fresh weight with an uptake rate constant of 193 mL-g -1 -day -1 and a depuration rate constant of 0.56 day -1 . In summary, flowthrough exposures of NP with fathead minnows conducted in accordance with US EPA guidelines produced steady-state BCF values of approximately 300 mL/g, with uptake rate constants of ~150 mL-g -1

- d -1 and depuration rate constants of 0.5 day -1 ; the depuration rate constant s are consistent with an

elimination half-life for NP from fath ead minnows of approximately 1.5 days.

CITI (1992) reported bioconcentration factors for carp ( Cyprinus carpio ) exposed to measured concentrations of linear NP (25154-52-3) in separate exposures at 100 micrograms/L and 10 micrograms/L in a flow through system. According to the EU risk assessment, the test is considered valid

for the determination of the bioconcentration factor (EU, 2003). The carp were exposed to linear NP for eight weeks. At the 100 micrograms/L exposure con centration, bioconcentration factors of 250-330 were measured over the eight week period. At the 10 mi crograms/L exposure level, bioconcentration factors ranged from 90 to 220. These results are similar to those observed in an US EPA-guideline study (Ward and Boeri, 1991) with fathead minnows (a member of the carp family) and the results collectively indicate that NP has a low potential for bioaccumulation in fish.

Brooke (1993) determined bioconcentration factors for fathead minnow ( Pimephales promelas ) and bluegills ( Lepomis macrochirus ) over 28 days with exposure to 5 sep arate concentrations of NP (84852- 15-3). Results of the study are considered valid fo r the purposes of categorization. For fathead minnows exposed to concentrations of 9.3, 19.2, 38.1, 77.5 and 193 micrograms/L, the mean BCF (+ standard deviation) value was 586 ± 273 mL/g wet weight after 14 days exposure and 741 ± 206 mL/g wet weight after 28 days exposure. The BCF with fathead minnows was found to be independent of NP concentration at 28 days but not at 14 days; reduced growth of the fish was seen at the two highest concentrations tested. For bluegill exposed to NP concentrations of 5.6, 12.4, 27.6, 59 and 126 micrograms/L, the mean BCF (+ standard deviation) value was 262 ± 70 mL/g wet weight after 14 days and 220 mL/g wet weight after 28 days. The BCF for this species was found to be independent of exposure concentration at the three lowest concentra tions, but the BCF at the two higher concentrations was found to be lower, particularly at 14 days, than that obtained at the lower concentrations.

### 21 Lewis and Lech (1996) studied the uptak e, disposition, and persistence of 14 C-NP from water in the rainbow trout ( Oncorhynchus mykiss ). The bioconcentration of 14 C-NP in the trout was determined by exposing juvenile fish weighing 40-60 g unde r static conditions to 18 micrograms/L 14 C-nonylphenol (uniformly ring labeled) for 5-24 hours. The calcula ted bioconcentration factors were 24 mL/g for the whole fish after 5 hours exposure and 98 for the viscera after 24 hours exposure. Given the limited exposure period, the study should be used with caution. To investigate the disposition of 14 C-NP in the trout, juvenile fish weighing 40-60 g were expo sed under static conditions to 36 micrograms/L 14 C-NP for

14 hours. The test chemical was detected in the following tissues in descending order of concentration: bile, liver, kidney, fat, gill, heart, and muscle. The half-life of 14 C-nonylphenol in specific tissues was determined by exposing juvenile fish weighing 40 to 60g under static conditions to 18 micrograms/L 14 CNP for 8 hours. The half-life was calculated as 19.8 hours in fat, 18.6 hours in muscle, and 5.9 hours in liver; these dissipation half-life values for NP in fish ar e consistent with previous research. To investigate

the metabolism of 14 C-NP in the fish, bile from exposed fish was analyzed and shown to contain three glucuronide metabolites, indicating phase II metabolism of NP in rainbow trout.

Giesy et al. (2000) determined bioconcentration factors for adult fathead minnow ( Pimephales promelas ) following 42 days exposure to five separate concentra tions of NP (commercial product, branched). NP concentrations in water were measured at the beginning, middle, and end of the experiments. For fathead minnows exposed to concentrations of 0.05, 0.16, 0. 4, 1.6, and 3.4 micrograms/L, the BCF (on a wet weight basis) values ranged from 203 to 268 after 42 da ys, indicating that BCF was independent of the exposure concentration. Results of this study have been considered valid by Environment Canada for the purposes of categorization. However, Environmen t Canada has assigned the results to CAS number 104- 40-5 which would indicate the linear NP isomer. This assignment is incorrect, given the test material was

obtained from one of the manufacturers of commercial NP (Schenectady International). As indicated in Table 4(a) (See Appendix), the experimental results of Giesy et al. (2000) have been assigned to CAS RN

84852-15-3.

The bioconcentration of branched NP in the killifish ( Oryzias latipes ; medaka) was investigated by Tsuda et al. (2001). The fish (0.16-0.24 g wt) were exposed to measured NP concentrations of 3.6 + 0.9 micrograms/L NP in a flow-through system for 7 days. The exposure period was then followed by a 1- day depuration period. BCF values in whole fish (wet weight) reached a plateau after 48 hours of exposure, suggesting that steady state conditions were established; hence the study has been judged use with caution. Exposure of the minnows to 3.6 micrograms/L NP in water for 7 days resulted in a median

BCF of 167 + 23 mL/g-wet weight. Following transfer to clean water the excretion rate constant and biological half-life were determined to be 0.07 h -1 and 9.9 h, respectively.

Smith and Hill (2004) have recently examined the upt ake and metabolism of technical NP in the roach ( Rutilus rutilus ). Branched 14 C-NP was synthesized using procedures consistent with the manufacture of the commercial product. Sexually mature roach were exposed to 4.9 + 1.1 micrograms NP/L in a flow through system. A 4-day exposure period was chosen fo r the study since prior work had shown that NP concentrations in the tissues of trout reached steady st ate after 4 days, findings which are consistent with previously cited work. Bioconcentration factors for NP in various fish tissues were determined; however,

a BCF value for whole fish was not calculated. The c oncentration of NP residues was the highest in bile and liver, with apparent BCF tissue values of 34,121 and 605, respectively. In the other tissues, apparent BCF tissue values ranged between 13 and 250. A single major metabolite of NP was present in liver and bile, which was identified as the glucuronide conjugate of 4-(hydroxyl-nonyl)-phenol.

Field studies

In addition to laboratory studies, the bioaccumulation of NP has also been ex amined in several field experiments. While laboratory tests involving wate r exposures are considered as providing information on bioconcentration, field studies are considered as providing information on bioaccumulation potential since exposures may arise from both food and water. Thus, field studies may provide more relevant indications of bioaccumulation potential for classification and risk assessment purposes.

### 22 Ahel et al. (1993) studied the bioaccumulation potential of NP in freshwater organisms in the Glatt River

and one of its tributaries in Switzerland. The av erage concentration of NP in the river was 3.9 micrograms/L. NP concentrations in the tissues of fish collected from the river were as follows: Squalius cephalus , muscle 0.18 mg/kg dry weight, gut 0.46-1.2 mg/ kg dry weight, liver 1.0-1.4 mg/kg dry weight, gills 0.98-1.4 mg/kg dry weight; Barbus barbus L., muscle 0.38 mg/kg dry weight, gut 0.05 mg/kg dry weight, liver 0.98 mg/kg dry weight, gills <0.03 mg/ kg dry weight, heart 0.30 mg/kg dry weight, roe 0.09 mg/kg dry weight; Oncorhynchus mykiss , muscle 0.15 mg/kg dry weight, gut 1.6 mg/kg dry weight. Based upon the average concentration of NP in water, the BAF for fish was calculated in the range of 13 to 408 L/kg dry weight on an individual organ basis. Because NP concentrations in the fish were lower than those detected in algae, the authors conclude d that negligible bioaccumulation of NP was occurring through the food chain.

To convert the field BAF values reported by Ahel et al. (1993) to units comparable with the other BCF data, the reported dry weight con centrations were recalculated on a fresh (or wet) weight basis by assuming that fish muscle was 85% water and 15% dry matter (Staples et al., 1998). As reported in Table 4(b) (See Appendix) the resulting non-lipid based fresh weight BAF values for NP for the three species ranged from 6 to 15 L/kg.

Liber et al. (1999) exposed bluegill sunfish to NP in a series of 18 aquatic mesocosms (littoral enclosures) in northeastern Minnesota. The test chemical, high purity NP (96.4%, branched isomers, Schenectady International), was applied to the enclosures every tw o days over a 20-day period at dose levels of 3, 30, 100 and 300 micrograms/L. NP concentrations were m easured in the water column and in fish, plants, and sediments. Based on the results, the calculated bi oaccumulation factor for NP in bluegill sunfish was

87 + 124 L/kg wet weight.

Tsuda et al. (2001) examined the bioaccumulation potential of NP in ayu fish ( Plecoglossus altivelis ) from three rivers flowing into Lake Biwa (Japan). NP was detected in all samples of river water, with median concentrations of 0.14 to 3.08 micrograms/L. NP concentrations in whole fish collected from the rivers ranged from 10 to 110 mg/kg wet weight. Based on NP concentrations in the river water and fish, a BAF for NP in the ayu fish of 21 + 15 L/kg was calculated.

Hu et al. (2005) have recently studied the trophodynami c behavior of NP and NP ethoxylates in a marine aquatic food web from Bohai Bay, North China. NP and NPEO concentrations were measured in 14 marine species. Co-analysis of DDT and DDT metabolites in of all the samples allowed direct comparison of the bioaccumulation behavior of DDT w ith that of NP and NPEO. The lipid equivalent concentrations of several DDT metabolites (DDE and DDMU) increased with increasing trophic level as determined by stable isotope ratios. Trophic ma gnification factors for DDE and DDMU were 3.26 and 3.7, respectively. In contrast, lipid equivalent concen trations of NP and all of the NPEO did not exhibit a statistically significant correlation with trophic levels in the food web. Trophic magnification factors for NP and NPEO ranged from 0.45 to 1.22. These results show that trophic magnification of NP and NPEO is not expected.

Bioaccumulation Conclusions - Nonylphenol

Based on a comprehensive search of the literature, it is a pparent that considerable information is available on the BAF and BCF potentials for NP in fish. R esults are available for studies conducted according to established guidelines, as well as research projects which contribute to the overall weight of evidence. Valid laboratory studies conducted with both linear a nd branched NP isomers indicate that BCF values range from 90 to 741 L/kg wet weight, although the major ity of the values fell in the range of 200 to 350 L/kg for the various fish species. BCF values fo r studies judged ‘use with caution’ ranged from 75 to 1300 L/kg wet weight. The higher BCF values are likely due to the fact that they were derived from total radioactivity measurements. None of these values exceed the criteria for categorization as B. Further, the similarities in structure and log K

ow values as well as comparable BCF values for the various test chemicals support the conclusion that the uptake, pa rtitioning, and metabolism behavior of linear and branched NP isomers are essentially identical in fish . Consequently, the information available for all

### 23 CAS RNs, linear and branched, should be considered in the decision, and the categorization conclusion applied to all members of the NP family.

Results of several field studies are also availa ble for assessing the BAF for NP, including mesocosm

studies where littoral enclosures were dosed with the test chemical. It is important to note that the results of field studies, where NP concentrations are measu red in both the water column and in fish tissues, would be expected to be representative of the envi ronmental behavior of the commercially relevant,

branched isomers. BAF values calculated from fiel d studies ranged from 6 to 87 L/kg wet weight. None of the BAF values exceed the criteria for categorization as B. Further, the fact that field determined BAF values are lower than laboratory measured BCF values indicates that trophic magnification is not expected.

In summary, an examination of fish uptake and elimination constants indicate that all NP isomers demonstrate uptake into fish that is significantly attenuated by rapid elimination/metabolism processes in fish, resulting in generally low to moderate labor atory BCF values and no significant bioaccumulation in field studies. Overall, the weight of evidence suppor ts the conclusion that the four NP CAS RNs on the

Canadian Domestic Substance List (104-40-5, 11066-49-2, 25154-52-3 and 84852-165-3) do not meet the criteria to be categorized as B.

Octylphenol (OP)

Environment Canada has categorized two OP subs tances as meeting the criteria for B based on QSAR

models while another CAS number for OP is cat egorized as not meeting the criteria based on experimental data. The rationale for th e proposed categorizations is as follows:

Proposed Categorizations for OP CAS Numbers CAS Number Description Proposed Categorization Evidence Cited

140-66-9 4- tert -octylphenol B QSAR 1806-26-4 Phenol, 4-octyl NOT B Exp. (CITI, 1992) 27193-28-8 Phenol, octyl B QSAR

Gobas and Arnot (2003) previously reviewed some of the available data for bioaccumulation potential for OP. BCF data for OP (CITI, 1992) was initially rej ected based on the assumption that the results were based on nominal exposure concentrations. This decisi on has since been reversed, and the study has been accepted as valid.

As summarized in Tables 4(a) and 4(b) (See Appendi x), a number of additional studies are available to inform the decision on the BAF and BCF potential for the various OP isomers. The subject has also been reviewed in the UK Environment Agency’s Environmental Risk Assessment on 4- tert -octylphenol (UKEA, 2005). Note that some studies have specified the CAS RN of the test chemical, while others

have indicated the chemical structure. While so me are judged as “valid” or “reliable” for assessing the BAF and BCF potential, the other studies (e.g., ‘use w ith caution’) are useful for contributing to the overall weight of evidence for assessing BAF potential for OP.

Further, since the uptake, partitioning, and metabolism behavior of linear and branched OP isomers would be expected to be essentially identical in fish, th e information available for all CAS RNs should be evaluated in the decision, and then applied to the cate gorization decision for all members of the OP family. This approach is supported by the available data for BCF potentials of branched and linear NP, as previously discussed, and considering the similarity in structures the NP data further support the

conclusions for OP.

Laboratory studies

### 24 CITI (1992) reported bioconcentration factors for carp ( Cyprinus carpio ) exposed to measured concentrations of OP at 100 micrograms/L and 10 micrograms/L in a flow-through system. This study is considered valid by Environment Canada for the pur poses of categorization. However, based on the CAS number used to index the study, there could be some confusion over the nature of the test substance. While the study information is accessible by searching on CAS RN 1806-26-4, the name of the test chemical (p-(1,1,3,3-tetramethylbutyl)phenol), the ch emical structure shown, and the CAS RN indicated in the section with the BCF study results (140-66-9) is descriptive of the commercially relevant product. During the CITI (1992) study, carp were exposed to OP in a flow-through mode for eight weeks. At the 10 micrograms/L exposure concentration, bioconcen tration factors of 12-135 L/kg were measured over the eight week period, whereas at the 100 micrograms/L exposure, bioconcentration factors ranged from 113-469 L/kg. These results indicate that OP has a low to moderate bioconcentration potential.

Ferreira-Leach and Hill (2000) examined th e bioconcentration and metabolism of 4- tert -octylphenol (4- (1,1,3,3-tetramethylbutylphenol)) in the roach ( Rutilus rutilus ). Roach fry, aged 7 days post hatch (DPH), were exposed to 5.8 micrograms/L of 14 C-t-OP in a semi-static system, and were analyzed after 5, 12, and 19 days. Steady state conditions were established af ter 12 days and BCF values for whole fish ranged from 1061-1134 L/kg over days 12-19. However, because these values are based on total 14 C measurements, it is likely that the presence of me tabolites in the organisms may have led to an overestimate of the bioconcentration potential. Con sequently, the BCF values have been judged ‘use with caution’. Radioactivity analysis of 7 DPH fry exposed for 5 days to 14 C-t-OP revealed that in young fish, the majority of the t-OP residues were present as th e parent compound. In contrast, when 26 DPH fish were exposed for 5 days to 14 C-t-OP, only 22% of the total radioactivity in the fish was parent compound. Since the majority of the test chemical was found to be converted to the glucuronide conjugate, these results suggest that roach fry can rapidly convert 4- t-OP into polar metabolites and this metabolic ability

increases with maturation of the ro ach larvae to the fry stage.

Ferreira-Leach and Hill (2001) continued their investig ation of the biotransformation, bioconcentration,

and tissue distribution of 4- t- OP in juvenile rainbow trout ( Oncorhynchus mykiss ) using radiolabeled test chemical. Bioconcentration factors were measured in the study, with a value for whole fish of 470 L/kg after 10 days exposure to 4 micrograms OP/L in a flow-through exposure system. Although steady state conditions were achieved after 4 days, given the limite d exposure and the fact that BCF values were based on total radioactivity, the study should be ‘used with caution’. The results suggested that exposure to

waterborne 4- t- OP results in rapid conjugation and elimination of the chemical via the liver/bile route, but that the parent compound can accumulate in a variety of other fish tissues. Some individual tissues had higher BCF values; for example, BCF of 800-1,200 L/ kg were measured in fat, intestine, liver, and pyloric caeca, but BCF were below 300 L/kg in other ti ssues. These values are for soluble residues, and so likely include contributions from metabolites as well as from the parent compound.

The bioconcentration of 4-t-OP in the killifish ( Oryzias latipes ; medaka) was investigated by Tsuda et al. (2001). The fish (0.16-0.24 g wt) were expo sed to measured OP concentrations of 4.7 + 0.8 micrograms/L in a flow-through system for 7 days. The exposure period was then followed by a 1-day depuration period. BCF values in whole fish (wet weight) reached a plateau after 48 hours of exposure, suggesting that steady state conditions were establis hed. Exposure of the minnows to 4.7 micrograms/L OP in water for 7 days resulted in a median BCF of 261 + 62 L/kg (wet weight). Following transfer to

clean water the excretion rate constant and biol ogical half-life were determined to be 0.09 h -1 and 7.7 h, respectively.

The metabolism of t-OP was examined by Pederson and Hill (2002) in the adult rudd ( Scardinius erythrophthalmus ). The fish were exposed to 4.7 micrograms/L 14 C-t-OP in a flow-through system for 10 days, followed by a 10 day elimination period, and BCF values were measured for various tissues. The researchers noted that OP was extensively metaboli zed to at least 10 different metabolites, including aromatic and aliphatic hydroxylation, glucuronidation, and glucosidati on. While whole fish BCF values for OP were not calculated, the significant degree of me tabolism noted for OP in the rudd is similar to that observed previously by other researchers on OP with other fish species.

Field studies

### 25 In addition to laboratory studies, the bioaccumulation of OP has also been examin ed in the field. As previously discussed, field studies are considered as providing information on bioaccumulation potential since exposures may arise from both food and water, and this may provide more relevant indications of bioaccumulation potential for risk assessment purposes.

Tsuda et al. (2001) examined the bioaccumulation potential of OP in ayu fish ( Plecoglosus altivelis ) from three rivers flowing into Lake Biwa (Japan). OP was detected in only some of the water samples, with concentrations ranging from below the detection limit to 0.07 micrograms/L. OP concentrations in whole fish collected from the rivers ranged from below the detection limit to 5.7 mg/kg wet weight. Based on OP concentrations in the river water and fish, a BAF for ayu fish of 297 + 194 L/kg was calculated.

Bioaccumulation Conclusions - Octylphenol

Based on a comprehensive search of the literature, it is apparent that additional information is available on the BAF and BCF potentials for OP in fish. R esults are available for studies conducted according to established guidelines, and for research projects which c ontribute to the overall weight of evidence. The BCF values measured in the valid study (CITI, 1992) ranged from 12 to 469 L/kg, while values from studies judged ‘use with caution’ ranged from 261-1134 L/ kg. The higher BCF values noted in the latter studies are likely due to the fact that they were de rived from total radioactivity measurements. The BAF value determined from a field investigation was 297. None of these values exceed the criteria for categorization as B.

The majority of the laboratory studies have used 4 -(1,1,3,3-tetramethylbutylphenol) as the test chemical,

which is consistent with CAS RN 140-66-9. However, it is reasonable to conclude that the uptake, partitioning, and metabolism behavior of linear OP (CAS RN 1806-26-4 and 27193-28-8) would be consistent with the behavior of the branched isomer, given the structural similarity and consistency in log

K

ow values. This conclusion is also supported by the la rger body of evidence available for the structurally related NP isomers, and specifically that BCF va lues were shown to be comparable for linear and branched materials. Consequently, th e BCF and BAF information available for tert -OP (branched) should be applied to all members of the OP family.

Overall, the weight of evidence supports the conc lusion that the three OP isomers on the Canadian Domestic Substance List (CAS RNs 140-66-9, 1806-26-4 and 27193-28-8) do not meet the criteria to be categorized as B.

Ethoxylates of NP and OP

Environment Canada has categorized the ethoxylates of NP and OP as not meeting the criteria for bioaccumulation in fish (See Table 1; Appendix). This decision is consistent with the physical properties, namely higher water solubility and lower K

ow values than the more hydrophobic phenols (NP and OP). Further, there is information reported in the litera ture on BAF values for the low mole NP ethoxylates (NP1EO, NP2EO, etc) which indicates that the ethoxyla tes have low potential to bioaccumulate in fish. Further, the potential to bioaccumulate would be expected to decrease as the degree of ethoxylation increases and water solubility increases. Given the st ructural similarities between the NP and OP ethoxylates, this conclusion is suppor ted for all members of the family.

Ahel et al. (1993) evaluated the bioaccumulation potential of NP1EO and NP2EO in freshwater organisms in the Glatt River and one of its tributar ies in Switzerland. The average concentrations of NP1EO and NP2EO in the river were 24 micrograms/L and 9.4 micrograms/L, respectively. NP1EO and NP2EO concentrations in the tissues of fish ( Oncorhynchus mykiss

, Squalius cephalus, Barbus barbus ) collected from the river ranged from 0.06 to 7.0 mg/ kg dry weight and < 0.03 to 3.0 mg/kg dry weight, respectively. Based upon the average concentrations of NP1EO and NP2EO in water, the BAF values for fish were calculated in the range of 3 to 300 L/kg dry weight and 3 to 326 L/kg dry weight, respectively.

### 26 To convert the field BAF values reported by Ahel et al. (1993) to units comparable with the other BCF data, the reported dry weight con centrations were recalculated on a fresh (or wet) weight basis by assuming that fish muscle was 85% water and 15% dry matter (Staples et al. 1998). As reported in Table 4(b) (See Appendix) the resulting non-lipid based, fresh weight BAF values for NP1EO and NP2EO for the three species ranged from 0.8 to 37 L/kg.

Bioaccumulation Conclusions - Ethoxylat es of Nonylphenol and Octylphenol

In conclusion, none of the NP or OP ethoxylates would be expected to bioaccumulate; hence they do not

meet the criteria to be categorized as B. This conclu sion is consistent with the current categorization decision proposed by Environment Canada.

VI. CONCLUSIONS

Considerable experimental information is available in the published literature to evaluate the persistence and bioaccumulation potential for the family of NP, OP, and their ethoxylates as addressed in this report. To be consistent with previous governmental assessments (EC-HC, 2001; US EPA, 2004; Rodier, 1996; EU, 2001; ECB, 2003; UKEA, 2005) of these materials Environment Canada shoul d consider data from any of the isomers as representative of the group during the conduct of the CS DSL categorization

assessment.

 Based on the availability of ready biodegradability data sufficient to support evaluation of all substances on the DSL, estimated environmental half- lives for water and soil are in the range of 5 to 30 days, which are clearly below the criteria that Environment Canada has defined for categorization of substances on the DSL (Water, Soil t1⁄2 < 182 days ). This conclusion is supported by additional

results of numerous simulation tests conducted with representative members of the family of NP, OP, and the ethoxylates.

 Experimental BCF values for fish are also suffici ent to support evaluation of all substances on the DSL; BCF values range from 12 to 741 L/kg wet weight for valid studies, which are clearly below the

criteria that Environment Canada has defined fo r categorization of substances on the DSL (BCF > 5,000).

Based on a comprehensive review of the available experimental data, all of the NP or OP isomers and the NP and OP ethoxylates on the Domestic Substance List do not meet the criteria for P or B as defined by Environment Canada. Although it is acknowledged that certain members of the family meet the criteria for inherent toxicity (iT) to environmental organisms, the fact that criteria for P and B are not met

indicates that none of the members of the family should be categorized as PiT or BiT. These categorizations are also consistent with pers istence and bioaccumulation conclusions in other governmental assessments of these compounds (EC-HC, 2001; US EPA, 2004; Rodier, 1996; EU, 2001 ECB, 2003; UKEA, 2005).

Recommended Revised Categorizations for NP/NPE and OP/OPE

CAS RN Decision Evidence/Basis for Decision Nonylphenol

104-40-5 Not P

Not B Exp. (Ying and Kookana, 2003) Exp-Group; Analogy with 25154-52-3 and 84852-15-3

11066-49-2 Not P

Not B Exp-Group; Analogy with 84852-15-3 Exp-Group; Analogy with 84852-15-3

25154-52-3 Not P

Not B Exp-Group; Analogy with 84852-15-3 Exp. (CITI, 1992)

84852-15-3 Not P

Not B Exp. (Staples et al., 1999, 2001 and others) Exp. (Wart and Boeri, 1991; Brooke, 1993; Giesy et al., 2000)
Nonylphenol Ethoxylates

9016-45-9 Not P

Not B Exp. (Staples et al., 1999, 2001 and others) Exp-Group (Ahel et al., 1993, Phys/chem.. properties)

### 27 26027-38-3 Not P

Not B Exp-Group; Synonym/Analogy with 9016-45-9 Exp-Group Synonym/Analogy with 9016-45-9

37205-87-1 Not P

Not B Exp-Group; Synonym/Analogy with 9016-45-9 Exp-Group Synonym/Analogy with 9016-45-9

68412-54-4 Not P

Not B Exp-Group; Synonym/Analogy with 9016-45-9 Exp-Group Synonym/Analogy with 9016-45-9

127087-87-0 Not P

Not B Exp-Group; Synonym/Analogy with 9016-45-9 Exp-Group Synonym/Analogy with 9016-45-9 Octylphenol

140-66-9 Not P

Not B Exp. (Staples et al., 1999, 2001 and others) Exp. (CITI, 1992)

1806-26-4 Not P

Not B Exp-Group; Analogy with 140-66-9 Exp-Group; Analogy with 140-66-9

27193-28-8 Not P

Not B Exp-Group; Analogy with 140-66-9 Exp-Group; Analogy with 140-66-9 Octylphenol Ethoxylates

9002-93-1 Not P

Not B Exp-Group; Synonym/Analogy with 9036-19-3 Exp-Group; Analogy with analogy with NPE; Phys/chem. properties

9036-19-3 Not P

Not B Exp (Staples et al., 1999, 2001) Exp-Group; Analogy with NPE; Phys/chem. properties

68987-90-6 Not P

Not B Exp-Group; Synonym/Analogy with 9036-19-3 Exp-Group; Analogy with NPE; Phys/chem. properties

### 28 VII. REFERENCES

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Alkylphenols & Ethoxylates Research Council (APERC). www.aperc.org . Personal communication with

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Bhatt, B.D., Prasad, J.V., and Ali, S. (1992). Separation and characterization of isomers of p -nonylphenols by capillary GC/GC-MS/GC-FTIR techniques. J. Chromatographic Sci., 30, 203-210.

Brooke, L.T. (1993). Acute and chronic toxicity of nonyphenol to ten species of aquatic organisms. EPA report 68 C-0034. U.S. Environmental Protection Agency, Duluth, MN.

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Dettenmaier, E.M., Chard, J.K., and Doucette, W.J. (2004). Fate of alkylphenol-polyethoxylates in soil/biosolids systems planted with crested wheatgrass. Society of Environmental Toxicology and Chemistry, 25th Annual Meeting, Portland, OR.

DiCorcia, A., Costantino, A., Crescenzi, C., et al. ( 1998). Characterization of recalcitrant intermediates from biotransformation of the branched alkyl side ch ain of nonylphenol ethoxylate surfactants. Environ. Sci. Technol., 32, 2401-2409.

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Dubroca, J., Brault, A., Kollmann, A., Touton, I., Jo livalt, C., Kerhoas, L., and Mougin, C. (2005). Biotransformation of nonylphenol surfactants in so ils amended with contaminated sewage sludges. Environmental Chemistry: Green Chemistry and Pollutants in Ecosystems, pp. 305-315.

Dutka, B.J., Liu, D., Jurkovic, A., McInnis, R., Lee, H-B., Onuska, F., and Rao, S.S. (1998). Observations from a six month study on the effect of biodegradation processes in sediment on the toxicity potential of targeted chemicals. Environ. Toxicol. Water Qual., 13, 313-322.

Ekelund, R., Bergman, A., Granmo, A., and Berggren , M. (1990). Bioaccumulation of 4-nonylphenol in marine animals - a re-evaluation. Environ. Pollut., 64, 107-120.

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### 29 Environment Canada and Health Canada (EC-HC). (2001). Priority Substances List Assessment Report: Nonylphenol and Its Ethoxylates. www.ec.gc.ca/ccebl/eng/final/index_e.html

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Huntsman Corporation. (1998a). T echnical Bulletin for Nonylphenol (NP).

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Huntsman Corporation. (1999a). Technical Bulletin for Surfonic N-10 (NPE1.5).

### 30 Huntsman Corporation. (1999b). Technical Bulletin for Surfonic N-40 (NPE4).

Huntsman Corporation. (1999c). Technical Bulletin for Surfonic N-95 (NPE9.5).

Huntsman Corporation. (1999d). Technical Bulletin for Surfonic N-300 (NPE30).

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John, D.M., and White, G.F. (1998). Mechanism fo r biotransformation of nonylphenol polyethoxylates to Xenoestrogens in Pseudomonas putida . J. Bacteriol., 180(17), 4332-4338.

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### 31 Manzano, M.A., Perales, J.A., Sales, D., et al. (1998) . Effect of concentration on the biodegradation of a nonylphenol polyethoxylate in river water. Bull Environ. Contam. Toxicol., 61, 489-496.

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### 33 APPENDIX - TABLES AND FIGURESPage 34Table 1. Environment Canada Proposed Categorization Decisions (April 20, 2005)CAS RN Sort/GroupMeets ECCriteria?Color Code ConfidenceCategoryApproachPersistentBioaccumulativeInherentlyToxicNonylphenol104-40-5 Organics No Red naNot P -QSAR Not B - Exp.iT - Exp.11066-49-2 Organics Yes Blue MediumNot P -QSARB - QSARiT - QSAR25154-52-3 Organics Yes Purple HighP - Exp.Not B - Exp.iT - Exp.84852-15-3UVCB YesPurpleUVCBNot P -QSARB - QSARiT - QSARNonylphenol Ethoxylates9016-45-9OriginalPolymersNo White High YesP (Exp.-preliminary)*Not BNot iT (Exp.-preliminary*26027-38-3OriginalPolymersNo White Medium YesP (Group)Not BNot iT(Group)37205-87-1OriginalPolymersNo White Medium YesP (Group)Not BNot iT(Group)68412-54-4UVCBPolymersNo White YesP (Group)Not BNot iT(Group)127087-87-0UVCBPolymersNo White YesP (Group)Not BNot iT (Exp.-preliminary*Octylphenol140-66-9Organics Yes Purple HighNot P -QSARB - QSARiT - Exp.1806-26-4 Organics No White naP - Exp.Not B - Exp. Not iT - Exp.27193-28-8 Organics Yes Yellow MediumNot P -QSARB - QSARiT - QSAROctylphenol Ethoxylates9002-93-1OriginalPolymersNo White HighP (Exp.-preliminary*Not BiT - Exp.9036-19-5OriginalPolymersNo White HighP (Exp.-preliminary*Not BiT - Exp.68987-90-6UVCBPolymersNo White YesNot P(Group) Not B iT (Group)Page 35Table 2. Summary of the Ready Biodegradability of APE and AP SubstancesCompound CAS RN Test Method Result Estimated AquaticHalf-Life (days)ReferenceNP 84852-15-3 OECD301B (GLP) 48.0% ThCO210Staples et al., 1999, 2001NP 84852-15-3 OECD301F (GLP) 62.0% ThOD5Staples et al., 1999, 2001NP 25154-52-3 OECD 301C 0% ThODNR1CITI, 1992NPE1.5 9016-45-9 OECD301B (GLP) 58.7% ThCO210Staples et al., 1999, 2001NPE7 9016-45-9 Closed bottle testGledhill test260% ThOD40% ThCO2510Markarian, et al., 1989NPE9 9016-45-9 OECD301B (GLP) 79.5% ThCO25Staples et al., 1999, 2001NPE9 9016-45-9 ISO 14593 69.5% ThCO25Staples et al., 1999, 2001NPE12 9016-45-9 Closed bottle testOECD 301BGledhill test2Gledhill test (a)330% ThOD65% ThCO245% ThCO242% ThCO23051010Hughes et al., 1989NPEC1 None OECD301B (GLP) 62.5 ThCO25Staples et al., 1999, 2001NPEC2 None OECD301B (GLP) 65.1 ThCO25Staples et al., 1999, 2001OP 140-66-9 OECD301B (GLP) 69.9% ThCO25Staples et al., 1999, 2001OP 1806-26-4 OECD 301C 0% ThODNR1CITI, 1992OP 1806-26-4 OECD 302C 0% ThODNR1CITI, 1992OPE1.5 9036-19-5 OECD301B (GLP) 64.6% ThCO25Staples et al., 1999, 2001OPE9 9036-19-5 OECD301B (GLP) 82.5% ThCO25Staples et al., 1999, 2001OPEC1 None OECD301B (GLP) 68.5 ThCO25Staples et al., 1999, 2001OPEC2 None OECD301B (GLP) 80% ThCO25Staples et al., 1999, 2001Page 36 Table 3. Summary of Biodegradati on Half-lives from Ready Tests, Laboratory Simulation Data, and Field Studies (All Media) for Studies that Used Commercially Relevant (Branched) Test Materials Compound Basis of decision Media Half-life (days) Reference NP Ready test Water

Soil

Sediment 5 to 10

5 to 10

20 to 40 NP River die-away Water 40.8 NP River die-away Water 2.5 to 35 NP Seawater die-away Water 26.3 NP Sediment die-away Sediment 26.3 NP Seawater die-away Sediment 5.9 NP Soil die-away Soil 10.4 NP Sludge-amended soil Soil 7.3 to 9.5 Staples et al., (1999, 2001)

NP Loamy sand die-away Soil 37 Jacobsen et al., (2004) NP Soil die-away Soil 4.5 to 16.7 Topp and Starratt (2000) NPE1.5 Ready test Water

Soil

Sediment 5

5

20 Staples et al., (1999, 2001)

NPE1.5 River die-away Water 57.8 NPE3 Soil die-away Soil 1.5 to 6 Kubiak (2002) NPE7 Ready test Water

Soil

Sediment 5 to 10

5 to 10

20 to 40 Markarian (1989)

NPE9 Ready test Water

Soil

Sediment 5

5

20 NPE9 Ready test (ISO) Water

Soil

Sediment 5

5

20 NPE9 River die-away Water 18.6 NPE9 Soil die-away Soil 18.2 Staples et al., (1999, 2001)

NPE10 Estuarine die-away, varying temperature and salinity 2.5 to 69 Kvestak and Ahel (1995)

NPE12 Ready test Water

Soil

Sediment 5 to 30

5 to 30

20 to 120 Hughes et et., (1989)Page 37 Table 4(a). Summary of Experimental Determination of Bioconcentration/Bioaccumulation Potential for NP, OP, and their Ethoxylates in Fish - Lab Studies

Test

Chemical CAS

RN Species Experimental Design BCF (mL/g

wet wt) Rate

Constants

and Halflives Comment Reference

NP Technical,

branched Not specified Atlantic

salmon

Salmo salar 4-d test

Static 170-310 ug/L 75-235

280 (k 1 /k 2 ) 45 mL-g -1

- d
1 (uptake)

0.16 d -1
(depuration)

4 d (depuration

t 1⁄2 ) Use with

caution

BCFs were

measured(nonequilibrium)

and

calculated (from k 1 /k 2 ; equilibrium) McLeese et al. ,1981

14 C-NP Analysis

showed

consistent

with

commercial

NP Not specified Threespine

stickleback

Gasterosteus

aculeatus 16-d exposure

32-d elimination Flow-through 4.8-4.9 ug/L 1200-

1300

1333 (k 1 /k 2 ) Dowcalculated:

K 1 = 1370 mL/g/day

k 2 = 1.03 day -1

Half-life =

16 hr

Time to90% steadystate = 2.2 days Use with

caution depuration complete in 5 days

BCF based

on 14 C-NP (>80% in fish, analyzed via TLC) Ekelund et al., 1990

NP

Branched 84852-

15-3 Fathead

minnow Pimephales promelas

Juvenile US EPA Guideline, GLP 20-d exposure

7-d elimination Flow-through [2] exposure concs (4.9 and 22.7 ug/L) 271

344 133 mL-g -1

- d -1 (uptake)

0.49 d -1
(depuration)

1.4 d (depuration

t 1⁄2 )

193 mL-g -1

- d -1 (uptake)

0.56 d -1
(depuration)

1.2 d (depuration

t 1⁄2 ) Valid (GLP) Steady state

reached during first 3-10 days Depuration was rapid Ward and Boeri,

1991a

NP

Linear 25154-

52-3 Carp Cyprinus carpio MITI Guideline

56-d test [2] exposure concs (10 and 100 ug/L) 90-220

250-

330 Valid CITI, 1992

NP

Branched 84852-

15-3 Bluegill

sunfish Lepomis

macrochirus 28-d test [5] exposure concs (5.6, 12.4, 27.6, 59, 126 ug/L) 220 Valid Rapid depuration Brooke,

1993

NP

Branched 84852-

15-3 Fathead

minnow Pimephales promelas 28-d test [5] exposure concs (9.3, 19.2, 38.1, 77.5, 193 ug/L) 741 Valid Rapid depuration Brooke,

1993Page 38 Table 4(a). Lab Studies (cont’d)

Test

Chemical CAS

RN Species Experimental Design BCF (mL/g wet wt) Rate

Constants

and Halflives Comment Reference

14 C-NP Not specified Rainbow

trout Oncorhyncus mykiss

Juvenile 5-24 h

exposures Static 18 ug/L NP 24 (carcass)

98 (vicera) Use with

caution

BCF based

on total 14 C

Three glucuronide

metabolites

detected Lewis and Lech, 1996

NP

Commercial product,

branched 84852-

15-3 (See text) Fathead

minnow Pimephales promelas

Adult 42-d exposure Flow-through [5] exposure concs (0.05, 0.16, 0.4, 1.6, 3.4 ug/L 203-268 Valid Giesy et al.,

2000

NP

Branched Not specified Killifish;

medaka Oryzias latipes 7-d exposure

1-d elimination Flow-through 3.6 ug/L 167 1.68 d -1
(depuration)

9.9 h (depuration

t 1⁄2 ) Use with

caution steady state

reached

after 48 h Tsuda et al., 2001

14 C-NP Branched,

consistent

with

commercial product Not specified Roach

Rutilus

rutilus

Adult 4-d exposure Flow-through 4.9 ug/L 13-250 (various

soft tissues) Use with

caution glucuronide

metabolite

identified Smith and Hill, 2004

OP (p-1,1,3,3) 140-66-9-

1806-26-

4 (See text) Carp Cyprinus carpio MITI Guideline

56-d test [2] exposure concs (10 and 100 ug/L) 12-135

113-469 Valid CITI, 1992

14 C-OP (p-1,1,3,3) Not specified Roach

Rutilus

rutilus Fry 19-d exposure

Semi-static 5.8 ug/L 1061-

1134 Use with

caution steady state

at 12 d

BCF based

on total 14 C,likely overestimated

metabolism

differs

between

larva and fry; fry rapidly

metabolize

OP to glucuronide conjugates FerreiraLeach and Hill, 2000Page 39 Table 4(a). Lab Studies (cont’d)

Test

Chemical CAS

RN Species Experimental Design BCF (mL/g

wet wt) Rate

Constants

and Halflives Comment Reference

14 C-OP (p-1,1,3,3) Not specified Rainbow trout Oncorhyncus mykiss

Juvenile 10-d exposure Flow-through 4 ug/L 470 Use with

caution steady state

reached after 4 days

OP

metabolized

to glucuronide conjugate FerreiraLeach and Hill (2001)

OP (4- tert ) Not specified Killifish; medaka Oryzias latipes 7-d exposure

1-d elimination Flow-through 4.7 ug/L 261 2.16 d -1
(depuration)

7.7 h (depuration

t 1⁄2 ) Use with

Caution Steady state

reached

after 48 h Tsuda et al., 2001

14 C-OP (p-1,1,3,3) Not specified Rudd

Scardinius erythrophthalmus

Adult 10-d exposure

10-d

elimination Flow-through 4.7 ug/L 24-169 (various tissues) Use with

caution

OP was extensively

metabolized (10 metabolites) rapid

elimination Pederson and Hill,

2002Page 40Table 4(b). Summary of Experimental Determination ofBioconcentration/Bioaccumulation Potential for NP, OP,and their Ethoxylates in Fish - Field StudiesChemical Species Experimental Design Observations Field BAFL/kg –wet wtReferenceNP Rainbow troutOncorhyncus mykisBarbelBarbus barbusChub(Leuciscus cephalusField studyGlatt River and tributaries(Switzerland)[NP] = 3.9 ug/L in water6157Ahel et al., 1993NP1EO Rainbow troutOncorhyncus mykisBarbelBarbus barbusChubLeuciscus cephalusField studyGlatt River and tributaries(Switzerland)[NP1EO] = 24 ug/L in water3191Ahel et al., 1993NP2EO Rainbow troutOncorhyncus mykisBarbelBarbus barbusChubLeuciscus cephalusField studyGlatt River and tributaries(Switzerland)[NP2EO] = 9.4 ug/L in water0.8372Ahel et al., 1993NP Bluegill sunfishLepomis macrochirusField littoral enclosuresCommercial NP appliedevery 2 days over 20 dayperiod[4] dosages (3, 30, 100, 300ug/L)Maximum NP levels in the water column2 h after application averaged 5 +4, 23 +11, 76 +21, and 243 +41 ug/L,respectively87 Liber et al., 1999NP Ayu fishPlecoglossus altivelisField studyRivers flowing into LakeBiwa (Japan)[NP] river = 0.14-3.08 ug/L[NP] fish = 10-110 mg/kg (wet weight)21 Tsuda et al., 2001OP Ayu fishPlecoglossus altivelisField studyRivers flowing into LakeBiwa (Japan)[OP] river = ND-0.07 ug/L[OP] fish = ND-5.7 mg/kg (wet weight)297 Tsuda et al., 2001Page 41Figure I: Pathway for Biodegradation of NPEsRRRO -(CH -2CH - O -)H2n-1O -(CH -2CH - O -)H2n-1O - CH -2COOHRO - CH -2CH - O - CH - COOH22NPnEONP2ECNP1ECRO -(CH -2CH - O -)H22NP2EORO - CH -2CH - OH2NP1EOROHNPNPEO(n-1)MAJORMAJORMAJORMAJORMINORCO2Ring-opened fragments,highly soluble carboxylicacids, alcohols, incorporate ascell biomass+(Anaerobic)