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. Woodburn1 

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 Substances List (DSL) with respect to persistence, 

bioaccumulation, and inherent toxicity (PBiT) characteristics. Applicable guidance for the evaluation of 

substances is given in EC (2003). Environment Canada has provisionally categorized selected substances 

on the DSL for PBiT and is soliciting further information to improve the technical basis of its initial 

assessment. 

This report is intended to inform the decision on the categorization of persistence and bioaccumulation 

potential for the family of nonylphenols (NP), octylphenols (OP), and their ethoxylates by: 

• Identifying additional studies either from the published literature (or unpublished industry 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 respect to persistence, bioaccumulation, and inherent 

toxicity (PBiT) characteristics by September 14, 2006. Environment Canada (EC, 2003) has published 

technical guidelines describing the systematic approach that 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 supporting 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 alkylphenols and their ethoxylates. Consequently, the industry 

members can contribute considerable information and expertise relevant to the environmental assessment 

of these substances. 

The objective of this report is to provide a technical 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 experimental 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 behavior, particularly the persistence and bioconcentration 

potential of the various NP, OP, and their ethoxylates, has been extensively summarized in several review 

articles (Staples et al., 1998; Servos, 1999; Talmage, 1994). In addition, various regulatory authorities 

have recently published their decisions on the environmental properties for many of these substances in 

risk assessment reports (Environment Canada and Health 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 complex 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 C9H18 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 nomenclature 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, changes 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 cover all nonylphenols. However, subsequent 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 irrelevant. 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 from any of the isomers were representative for NP 

unless otherwise specified, and NP was used as the generic name referring to all of these substances in the 

assessment (EU, 2001). The European Chemicals Bureau (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 for 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 branched p-NP, while noting that CAS RN 25154-52-3 

(nonylphenol, mixed isomers) and 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 nonylphenol ethoxylates, which was conducted under the 

Canadian Environmental Protection Act (CEPA). Environment Canada acknowledged the position of US 

EPA and the alkylphenols industry that CAS RN 84852-15-3 best represents the commercially relevant 

nonylphenol (Environment Canada and 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 various CAS RNs used to describe nonylphenol, there 

are in principle two forms of the material, i.e., linear 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 evaluations 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 (pNP) with ethylene oxide (EO). The branching of the nonyl group results in additional structural isomers. 

Commercially available NPEs generally range from four 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 applicable 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 with 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 Number

Percent 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 differ with respect to their level of description 

regarding branching and position of the nonyl group 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 C6H4(OH)C8H17. The 

octyl group is a chain of eight carbons, which may be linear 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-tetramethylbutyl)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-tertoctylphenol) and is recognized as such in a risk 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 isomers 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 various 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 previous regulatory assessments, during the 

conduct of the categorization assessment Environment Canada 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 target ethoxylation level. Commercially available 

OPEs generally range from 4 moles of ethoxylation (OPE4) up to 80 moles of ethoxylates (OPE80). 

Some branching of the octyl group results in additional 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 OPE 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 Categorizations 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 OPEs and illustrates discrepancies in decisions for 

equivalent or similar compounds. The following sections will provide a summary of the additional 

available data that will support consistent 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 for 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 responsible for removal of NP, OP, and their ethoxylates 

from aquatic and terrestrial environments. Hydrolysis is not expected to be important, 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, respectively. 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 rationale 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)-omegahydroxyP-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 ethoxylates 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, which 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 international literature on the environmental health, fate 

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

studies that examined the fate of NPE4-35 and OPE6-10. 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. Further, 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 mechanism 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 often 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 conditions, the aromatic ring of the alkylphenol and 

alkylphenol ether carboxylates is opened and the intermediate 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 ethoxylates, the available data on the biodegradation 

of APECs will also be presented. 

It is important to note that the Canadian DSL includes 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 others 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 branched 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 all 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 fundamental 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 processes (i.e., sequential removal of ethoxylate groups) 

are identical. Thus, the information available for all 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 information on the biodegradation of the family of NP, 

OP, and their ethoxylates. Information will be presented 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 biodegradation 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 Environment Canada’s guidance (EC, 2003), substances 

having >60% biodegradation based on theoretical CO2 (ThCO2) or oxygen demand (ThOD), or >70% 

based on dissolved organic carbon (DOC) removal (the different tests have different criteria) within 28 

days are said to be readily biodegradable. Given the 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 Inorganic 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 environmental 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% (ThCO2, 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 (CAS RN 140-66-9). Further, to provide a more 

comprehensive picture on the overall fate of the alkylphenol ethoxylates, the ready biodegradability of 

important biodegradation intermediates, including NPEC1, NPEC2, OPEC1, and OPEC2 were also 

investigated. Biodegradation was examined according 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 summarized in Table 2 (See Appendix). Based on CO2

production observed in the OECD 301B test, NPE9 and 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 OPEC2) exceeded 60% of the theoretical CO2 formation by 

day 28. Results of the OECD 301F test also confirmed that NP was extensively degraded, with 62% of 

the theoretical oxygen demand consumed after 28 days. The results also suggested that OP may 

biodegrade faster than NP. For example, in the OECD 301B test, NP was 48% biodegraded (ThCO2) by 

day 28, while OP was degraded by 69.9% on day 20. 

In addition to quantifying ultimate biodegradation by measuring CO2 production, primary biodegradation 

was quantified in the OECD 301B tests. Dissolved organic carbon (DOC), suspended organic carbon 

(SOC), as well as concentrations of APE, AP, and APEC remaining at days 15 and 35 were measured. 

Based on DOC removal, all compounds tested were degraded by 87.1 to 97.6%. Conversion of the test 

material into cellular biomass or sorbed onto particulates 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 remaining residues, degradation of the various parent 

materials tested ranged from 91.1 to ~100%. By day 35, the amount of degradation ranged from 92.3 to 

~100% of initial concentrations across all tests. 

Hughes et al. (1989) compared the ready biodegradability 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 CO2 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 CO2. No significant differences were noted, 

as mineralization of the test substance reached 45% ThCO2 with the unacclimated seed as compared to 

42% ThCO2 in tests with an acclimated inoculum. 

Markarian et al. (1989) also compared the biodegradability of NPE7 in the closed bottle test and the 

Gledhill test. The authors reported 60% ThOD and 40% ThCO2 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 Staples et al. (1999, 2001), the results from each of these 

screening tests are relatively consistent. Slight differences 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 biodegradability test as reported by the Japan MITI (CITI, 

1992) were selected as the pivot value by Environment Canada 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 addition, results of studies conducted using OECD 302C that 

showed no degradation were reported in the EU Risk 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 ready 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 substrates 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 pre-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 standard 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 branched NPE, OPE, NP, and OP. Based on the guidance 

provided by Environment Canada for the categorization of substances on the DSL (EC, 2003), the results 

of these tests were used to derive environmental half-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 sediment. Further, because the linear alkylphenol isomers 

would be expected to degrade at least as fast if not faster 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 simulation 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 results of static and aerated die-away tests with linear 4-n-NP in 

seawater and marine sediment. In seawater, 4-n-NP was nearly completely biodegraded by 90% within 7 

days under constant aeration. In sterile control 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 degraded by 42% in day 1 and 95% by day 21. From 

these results, the authors estimated half-lives of 5 days 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% during the first 21 days (acclimation period), and then 

rapidly degraded with 91% removal by day 28. From 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 sediment, 4-n-NP degrades may be more rapidly and more 

extensively than the branched OP. However, it is important 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 μg/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. Assuming 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 radiolabeled 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 14C-NP was present as extractable material. HPLC analyses showed that NP was present at 

negligible concentrations by day 32. From these data, the authors calculated biodegradation half-lives of 

4 days for the mix of linear and branched NP. 

To examine differences between the biodegradation rate 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 CO2 evolution. Based on the analysis of 

parent compound remaining in the soil in both treatments, 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 studies conducted with 500 mg/kg NP, CO2 evolution was higher than 

observed in controls, and indicated that about 61% of the parent compound was converted to CO2 after 94 

days incubation. In studies conducted with 10 mg/kg of test chemical, the CO2 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 days for linear NP in seawater and marine sediment, 

and were faster than rates measured for branched OP (half-life of 20 days). In soil, 4-n-NP and a mixture 

of linear and branched NP yielded primary biodegradation half-lives of approximately 15 days and 4 days, 

respectively. These confirm that studies with the more 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 sediments contain background concentrations of NP (~84 μg/g 

dry weight). One experiment measured the degradation 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 

μg/g NP). Sediment concentrations of NP were measured 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 during the test from degradation of other NPE residues 
 14 

present. NP in spiked sediment increased in concentration over the first three months, followed by 

degradation of 62.5% by the 6 month. It is likely 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 and analyzed for NPE using HPLC. Test vessels were 

either stirred or static during the test. Primary degradation with and without sediments was 97 to 99% by 

10 days in static vessels and in about 5 days in stirred 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. After 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 CO2. 

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 degradation 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 the 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 acclimated 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 controlling 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, summer, 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 phase 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 conducted 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 lower 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 14C-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 CO2 (55% by day 56). With seawater and sediment, no lag 

phase was observed and formation of CO2 started immediately reaching 44% by day 56. The overall 

mass balance of this study was <50%, so it is possible that additional mineralization occurred but CO2

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 die-away studies of NP and NPE1,2 in soil to which 

anaerobically digested sludge was added. NPE1,2 and NP were reduced by 80% in the first 21 days. 

Further reduction was negligible due to sorption of the metabolites to soil and their subsequent lack of 

bioavailability. 

Kubiak (2002) conducted laboratory lysimeter studies using two soil types with and without plants and 

monitored the leaching and degradation behavior of 14C-NPE3. Dissipation half-lives ranged from <1.5 

to <6 days. Lower 14C-NPE and 14C-NP were formed as transient products that were quickly degraded to 

14CO2 (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-14C-labeled NP in several soils as a 

function of temperature. 14C-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 the 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 remaining activity was not extractable (50% of initial 

activity). The authors also showed that the moisture content of the soil affected the rate of degradation. 

Saturated soil and nearly dry soil had significantly lower 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 oxygen availability. Heavily amended soil may limit 

degradation of the NP due to the high oxygen demand of the sludge. 

Kuchler (1996) conducted mineralization 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 CO2

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 concluded 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 examining the fate of NPE9, NPE4, and NP in soil 

amended with biosolids and which were either unplanted 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 incorporated to a depth of 15 cm. After incorporation, 

the NP concentration was 0.56 μg/g, dry weight. Lysimeters were placed outdoors to receive natural 

lighting, temperature, and precipitation. Lysimeters were 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-away of NP in the first 10 days (55% reduction). A 

slower but consistent reduction occurred throughout the remainder of the experiment (to 110 days). The 

half-life of the degradation of NP over days 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 content 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 distinguish between amounts of water-borne NP that was 

biodegraded and that adsorbed to sediments, macrophytes 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 proportionally 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 enclosure walls desorbed back into the water over time. NP sorbed to 

the plants either degraded or desorbed 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 dissipation half-life of NP in the water of 0.74 days is 

effectively a biodegradation half-life. 

Jonkers et al. (2003) reported on biodegradation 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 retention time of two to three months. Sources of NPE 

were wastewater treatment plants upstream of the estuaries. 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 soils in field trials. During a one year period, 10 soil 

plots were treated with two different sewage sludges that contained residues of NPE and NP. After 

application, the sewage sludge was incorporated into the 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 wastewater 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 substances are as APEC1-3+CAPEC1-3, which are acidic 

substances. The authors suggested that both the river and the wetland were highly effective at removing 

AP and APE and that the river effectively removed 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 collected 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 within the septic tank was observed. This was followed 

by a further 96.7% reduction of molar concentration within the leach lines. As the pore water migrated 

through the vadose zone, an additional 99.69% reduction in molar concentration 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 anoxic 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) for 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 related 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 removal process of alkylphenols and their ethoxylate from 

water, sediment, and soil - has been extensively studied 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 first are stringent screening tests that measured the 

ready biodegradability of a test substance. Ready biodegradation tests have been conducted for APE, AP, 

and APEC. The second type are simulation tests in which biodegradation is measured in laboratory tests 

using procedures and bacterial inocula that simulate potential degradation in specific environmental 
 18 

compartments or within wastewater treatment plants. Simulation tests for biodegradation of APE and 

their degradation intermediates are available for all relevant environmental compartments including 

freshwater, freshwater sediment, seawater, marine sediment, and soil. The third line of evidence that 

biodegradation is the key removal process for APE and 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. The Environment Canada Guidance (EC, 2003) gives 

instructions on how to estimate environmental half-lives 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 environmental compartments were also taken from the 

simulation studies that were reviewed above. Half-lives 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 materials 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-away half-lives were 5.9 and 26.3 days, respectively. 

NPE10 biodegradation under varying conditions of temperature and salinity 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 biodegradation 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 biodegraded 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 and their degradation intermediates in the environment 

is well understood. Their physical properties and partitioning 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 chemicals nor the branched commercial APE products - or 

any of their degradation intermediates (lower mole APE, AP and APEC) - can be categorized as persistent. 

V. BIOACCUMULATION ASSESSMENT 

Environment Canada has proposed various categorizations 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 Substances List (EC, 2003), when bioaccumulation (BAF) 

or bioconcentration (BCF) data are available for different 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 for chemical bioaccumulation in the food chain is best 
 19 

measured by the BAF or BCF in higher trophic level organisms. Consequently, in view of the extensive 

information available, the following discussion will focus 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 Classification 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 available data for bioaccumulation potential for NP. 

Several studies were rejected due to insufficient 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 results were based on nominal exposure concentrations 

(CITI, 1992). This decision has since been reversed, 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 various 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 exposure studies conducted over a period of 4 days. Given 

the short exposure period and the static exposure conditions, the study should 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 14C-labeled NP in three-spined stickleback 

(Gasterosteus aculeatus L.). The 14C-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 14C-NP in a flow-through 

system. Exposure was for 16 days followed by an elimination 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 factors for the fish were in the range of 1,200 to 1,300 

mg/L on a fresh weight basis. Subsequent kinetic analysis of these data by Dow revealed an uptake rate 

constant (K1) of 1370 mL-g-1-day-1 and elimination rate constant (k2) 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 offered for NP in stickleback by the authors. The 

authors noted that the BCF values are based on total 14C measurements in fish tissue, thus, the presence of 

metabolites in the organisms may have led to an overestimate of the bioconcentration potential. Although 

Gobas and Arnot had previously rejected this study during their review of the available information, the 

fact that steady state conditions were achieved, but that BCF values were based on total fish radioactivity, 

supports classifying this study as use with caution. 

The bioconcentration of NP (84852-15-3, commercial 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 Practices, 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, temperature, 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 constants are consistent with an 

elimination half-life for NP from fathead 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 concentration, bioconcentration factors of 250-330 were 

measured over the eight week period. At the 10 micrograms/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 separate concentrations of NP (84852-

15-3). Results of the study are considered valid for 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 concentrations, 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 uptake, disposition, and persistence of 14C-NP from water in the 

rainbow trout (Oncorhynchus mykiss). The bioconcentration of 14C-NP in the trout was determined by 

exposing juvenile fish weighing 40-60 g under static conditions to 18 micrograms/L 14C-nonylphenol 

(uniformly ring labeled) for 5-24 hours. The calculated 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 14C-NP in the 

trout, juvenile fish weighing 40-60 g were exposed under static conditions to 36 micrograms/L 14C-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 14C-nonylphenol in specific tissues was 

determined by exposing juvenile fish weighing 40 to 60g under static conditions to 18 micrograms/L 14CNP 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 are consistent with previous research. To investigate 

the metabolism of 14C-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 concentrations 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 days, 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, Environment 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 uptake and metabolism of technical NP in the roach 

(Rutilus rutilus). Branched 14C-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 for the study since prior work had shown that NP 

concentrations in the tissues of trout reached steady state 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 concentration 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 examined in several field 

experiments. While laboratory tests involving water 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 average 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 concluded 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 concentrations 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 two days over a 20-day period at dose levels of 3, 30, 

100 and 300 micrograms/L. NP concentrations were measured in the water column and in fish, plants, 

and sediments. Based on the results, the calculated bioaccumulation 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 trophodynamic 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 with 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 magnification factors for DDE and DDMU were 3.26 and 

3.7, respectively. In contrast, lipid equivalent concentrations 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 apparent that considerable information is available 

on the BAF and BCF potentials for NP in fish. Results 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 and branched NP isomers indicate that BCF values 

range from 90 to 741 L/kg wet weight, although the majority of the values fell in the range of 200 to 350 

L/kg for the various fish species. BCF values for 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 Kow values as well as comparable BCF values for the various test 

chemicals support the conclusion that the uptake, partitioning, 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 available 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 measured in both the water column and in fish tissues, 

would be expected to be representative of the environmental behavior of the commercially relevant, 

branched isomers. BAF values calculated from field 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 laboratory BCF values and no significant bioaccumulation in 

field studies. Overall, the weight of evidence supports 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 substances as meeting the criteria for B based on QSAR 

models while another CAS number for OP is categorized as not meeting the criteria based on 

experimental data. The rationale for the 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 rejected based on the assumption that the results were 

based on nominal exposure concentrations. This decision has since been reversed, and the study has been 

accepted as valid. 

As summarized in Tables 4(a) and 4(b) (See Appendix), 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 some are judged as “valid” or “reliable” for assessing the 

BAF and BCF potential, the other studies (e.g., ‘use with 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, the information available for all CAS RNs should be 

evaluated in the decision, and then applied to the categorization 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 purposes 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 chemical 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, bioconcentration 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 the 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 14C-t-OP in a semi-static system, and were analyzed after 5, 12, and 

19 days. Steady state conditions were established after 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 14C 

measurements, it is likely that the presence of metabolites in the organisms may have led to an 

overestimate of the bioconcentration potential. Consequently, the BCF values have been judged ‘use with 

caution’. Radioactivity analysis of 7 DPH fry exposed for 5 days to 14C-t-OP revealed that in young fish, 

the majority of the t-OP residues were present as the parent compound. In contrast, when 26 DPH fish 

were exposed for 5 days to 14C-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 roach larvae to the fry stage. 

Ferreira-Leach and Hill (2001) continued their investigation 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 limited 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 tissues. 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 exposed 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 established. 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 biological 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 14C-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 metabolized to at least 10 different metabolites, including 

aromatic and aliphatic hydroxylation, glucuronidation, and glucosidation. While whole fish BCF values 

for OP were not calculated, the significant degree of metabolism 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 examined 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. Results are available for studies conducted according to 

established guidelines, and for research projects which contribute 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 derived 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 

Kow values. This conclusion is also supported by the larger body of evidence available for the structurally 

related NP isomers, and specifically that BCF values were shown to be comparable for linear and 

branched materials. Consequently, the 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 conclusion 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 Kow values than the more hydrophobic phenols (NP and OP). 

Further, there is information reported in the literature on BAF values for the low mole NP ethoxylates 

(NP1EO, NP2EO, etc) which indicates that the ethoxylates 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 structural similarities between the NP and OP 

ethoxylates, this conclusion is supported 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 tributaries 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 concentrations 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 - Ethoxylates 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 conclusion 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 should 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 sufficient 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 for 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 persistence 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 

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 33 

APPENDIX - TABLES AND FIGURES 
Page 34 

Table 1. Environment Canada Proposed Categorization Decisions (April 20, 2005) 

CAS RN Sort/Group 

Meets EC 

Criteria? Color Code Confidence 

Category 

Approach Persistent Bioaccumulative 

Nonylphenol104-40-5 Organics No Red na 

Not P - 

QSAR Not B - Exp. 

11066-49-2 Organics Yes Blue Medium 

Not P - 

QSAR B - QSAR 

25154-52-3 Organics Yes Purple High P - Exp. Not B - Exp. 84852-15-3 UVCB Yes 

PurpleUVCB Not P - QSAR B - QSAR 

Nonylphenol Ethoxylates9016-45-9 

Original 

Polymers No White High Yes 

P (Exp.-

preliminary)* Not B 

26027-38-3 

Original 

Polymers No White Medium Yes P (Group) Not B 

37205-87-1 

Original 

Polymers No White Medium Yes P (Group) Not B 

68412-54-4 

UVCB 

Polymers No White Yes P (Group) Not B 

127087-87-0 

UVCB 

Polymers No White Yes P (Group) Not B 

Octylphenol140-66-9 Organics Yes Purple High 

Not P - 

QSAR B - QSAR 

1806-26-4 Organics No White na P - Exp. Not B - Exp. 27193-28-8 Organics Yes Yellow Medium 

Not P - 

QSAR B - QSAR 

Octylphenol Ethoxylates9002-93-1 

Original 

Polymers No White High 

P (Exp.-

preliminary* Not B 

9036-19-5 

Original 

Polymers No White High 

P (Exp.-

preliminary* Not B 

68987-90-6 

UVCB 

Polymers No White Yes 

Not P 

(Group) Not B 
Page 35 

Table 2. Summary of the Ready Biodegradability of APE and AP Substances 

Compound CAS RN Test Method Result Estimated Aquatic 

Half-Life (days) 

Ref

NP 84852-15-3 OECD301B (GLP) 48.0% ThCO2 10 Staples et aNP 84852-15-3 OECD301F (GLP) 62.0% ThOD 5 Staples et aNP 25154-52-3 OECD 301C 0% ThOD NR1 CITI, 1992 NPE1.5 9016-45-9 OECD301B (GLP) 58.7% ThCO2 10 Staples et aNPE7 9016-45-9 Closed bottle test 

Gledhill test2 

60% ThOD 

40% ThCO2 

5 

10 

Markarian, 

NPE9 9016-45-9 OECD301B (GLP) 79.5% ThCO2 5 Staples et aNPE9 9016-45-9 ISO 14593 69.5% ThCO2 5 Staples et aNPE12 9016-45-9 Closed bottle test 

OECD 301B 

Gledhill test2 

Gledhill test (a)3 

30% ThOD 

65% ThCO2 45% ThCO2 42% ThCO2 

30 

5 

10 10 

Hughes et a

NPEC1 None OECD301B (GLP) 62.5 ThCO2 5 Staples et aNPEC2 None OECD301B (GLP) 65.1 ThCO2 5 Staples et aOP 140-66-9 OECD301B (GLP) 69.9% ThCO2 5 Staples et aOP 1806-26-4 OECD 301C 0% ThOD NR1 CITI, 1992 OP 1806-26-4 OECD 302C 0% ThOD NR1 CITI, 1992 OPE1.5 9036-19-5 OECD301B (GLP) 64.6% ThCO2 5 Staples et aOPE9 9036-19-5 OECD301B (GLP) 82.5% ThCO2 5 Staples et aOPEC1 None OECD301B (GLP) 68.5 ThCO2 5 Staples et aOPEC2 None OECD301B (GLP) 80% ThCO2 5 Staples et a
 Page 36 

Table 3. Summary of Biodegradation 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 

(k1/k2) 

45 mL-g-1-d1

 (uptake) 

0.16 d-1

(depuration) 

4 d 

(depuration 

t1⁄2 ) 

Use with 

caution 

BCFs were 

measured 

(nonequilibrium) 

and 

calculated 

(from k1/k2; 

equilibrium) 

McLeese et 

al. ,1981 

14C-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 

(k1/k2) 

Dowcalculated: 

K1 = 1370 

mL/g/day 

k2 = 1.03 

day-1

Half-life = 

16 hr 

Time to 

90% steadystate = 2.2 

days 

Use with 

caution 

depuration 

complete in 

5 days 

BCF based 

on 14C-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 

t1⁄2 ) 

193 mL-g-1-

d-1 (uptake) 

0.56 d-1

(depuration) 

1.2 d 

(depuration 

t1⁄2 ) 

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, 

1993 
 Page 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 

14C-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 14C 

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 

t1⁄2 ) 

Use with 

caution 

steady state 

reached 

after 48 h 

Tsuda et 

al., 2001 

14C-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 

14C-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 14C, 

likely overestimated 

metabolism 

differs 

between 

larva and 

fry; fry 

rapidly 

metabolize 

OP to 

glucuronide 

conjugates 

FerreiraLeach and 

Hill, 2000 
 Page 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 

14C-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 

t1⁄2 ) 

Use with 

Caution 

Steady state 

reached 

after 48 h 

Tsuda et 

al., 2001 

14C-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, 

2002 
Page 40 

Table 4(b). Summary of Experimental Determination of Bioconcentration/Bioaccumulation Potential fo

and their Ethoxylates in Fish - Field Studies 

Chemical Species Experimental Design Observations Field BAF 

L/kg –wet wt 

R

NP Rainbow trout 

Oncorhyncus mykis Barbel Barbus barbus Chub( Leuciscus cephalus 

Field study Glatt River and tributaries (Switzerland) 

[NP] = 3.9 ug/L in water 6 

15 

7 

A

NP1EO Rainbow trout 

Oncorhyncus mykis Barbel Barbus barbus Chub Leuciscus cephalus 

Field study Glatt River and tributaries (Switzerland) 

[NP1EO] = 24 ug/L in water 3 

19 

1 

A

NP2EO Rainbow trout 

Oncorhyncus mykis Barbel Barbus barbus Chub Leuciscus cephalus 

Field study Glatt River and tributaries (Switzerland) 

[NP2EO] = 9.4 ug/L in water 0.8 

37 

2 

A

NP Bluegill sunfish 

Lepomis macrochirus 

Field littoral enclosures Commercial NP applied every 2 days over 20 day period [4] dosages (3, 30, 100, 300 ug/L) 

Maximum NP levels in the water column 2 h after application averaged 5 + 4, 23 +

11, 76 + 21, and 243 + 41 ug/L, respectively 

87 L

NP Ayu fish 

Plecoglossus altivelis 

Field study Rivers flowing into Lake Biwa (Japan) 

[NP] river = 0.14-3.08 ug/L [NP] fish = 10-110 mg/kg (wet weight) 

21 T

OP Ayu fish 

Plecoglossus altivelis

Field study Rivers flowing into Lake Biwa (Japan) 

[OP] river = ND-0.07 ug/L [OP] fish = ND-5.7 mg/kg (wet weight) 

297 T
Page 41 

Figure I: Pathway for Biodegradation of NPEs

R

R

R

O - (CH -2 CH - O - 2 )Hn-1

O - (CH -2 CH - O - 2 )Hn-1

O - CH -2 COOH

R O - CH -2 CH - O - CH - COOH 22

NPnEO

NP2ECNP1EC

R O - (CH -2 CH - O - 2 ) H2

NP2EO

R O - CH -2 CH - OH 2

NP1EO

R OH

NP

NPEO(n-1)

MAJOR MAJOR

MAJOR MAJOR

MINOR

CO2

Ring-opened fragments, highly soluble carboxylic acids, alcohols, incorporate as cell biomass

+

(Anaerobic)