aperc_np and npes

7/29/2019 Aperc_np and Npes

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ASSESSMENT OF THE PERSISTENCE

AND BIOACCUMULATION POTENTIAL FORNONYLPHENOL, 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

1The Dow Chemical Company, Midland, MI

2Assessment Technologies, Inc., Spotsylvania, VA

3 RegNet Environmental Services, Washington, DC


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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


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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 substanceson 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Identifying additional studies either from the published literature (or unpublished industry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

Nonylphenol104-40-5 Organics

11066-49-2 Organics

25154-52-3 Organics

84852-15-3 UVCB

Nonylphenol Ethoxylates

9016-45-9 Original Polymers


37205-87-1 Original Polymers68412-54-4 UVCB Polymers

127087-87-0 UVCB Polymers

Octylphenol

140-66-9 Organics

1806-26-4 Organics

27193-28-8 Organics

Octylphenol Ethoxylates



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


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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 Councils APE Panel, and its member companies have been actively engaged in

environmental fate and effects research on alkylphenols and their ethoxylates. Consequently, the industrymembers 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 thepara position of phenol. Commercial nonenedoes 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 thepara isomer predominates, the product is most accurately described aspara-nonylphenol,

branched (p-NP or PNP). The following table lists the results ofhigh-resolution GC analyses ofp-NP

and identifies 22 branchedpara-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 alpha 3 8 9%


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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 RN90481-04-2 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 synthesizedon 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 84852-15-3and 25154-52-3 (US EPA, 2003). In its RM1 document forpara-NP US EPA accepted the view of the

Alkylphenols & Ethoxylates Research Council (APERC) that CAS RN 84852-15-3 is the mostdescriptive 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 PrioritySubstance List (PSL) Assessment for NP and nonylphenol ethoxylates, which was conducted under the

Canadian Environmental Protection Act (CEPA). Environment Canada acknowledged the position of USEPA 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-,

branchedOther Names

Nonylphenol;

4 N l h l

EU (2001)

ECB (2003)US EPA (2003)

US EPA (1996)

Unspecified, mixed isomers


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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-NonylphenolOther 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 synonymsfor branched NPfor 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 asrepresentative for NP.

Nonylphenol Ethoxylates CAS RNs and Nomenclature

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

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


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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

PercentofTotal

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 descriptionregarding 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 Description9016-45-9 Poly (oxy-1,2-ethanediyl), alpha-(nonylphenyl)-omega-hydroxy-

26027-38-3 Poly (oxy-1,2-ethanediyl), alpha-(4-nonylphenyl)-omega-hydroxy -

37205-87-1 Poly (oxy-1,2-ethanediyl), alpha-(isononylphenyl)-omega-hydroxy-

68412-54-4 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

Oligomer Distribution of NPE9


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Octylphenol

(Sources: CAS, NIST, APERC)

CAS RN Description Other

Assessments

Structure

140-66-9 4-tert-octylphenol

Other namesPhenol, 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-


4-tert-Octylphenol;

p-(1',1',3',3'-


p-Terc.oktylfenol;

para-tert-Octylphenol;

Phenol, p-(tert-octyl)-

UKEA (2005)

1806-26-4 Phenol, 4-octyl

Other namesPhenol, 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 acomparablesynonym 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


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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 Canadas currently proposed categorizations for the

various CAS RNs that represent NP, NPEs, OP, and OPEs and illustrates discrepancies in decisions forequivalent 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 beenextensively 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,


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EC Proposed Categorization for NP/NPEs and OP/OPEs CAS RNs

CAS Number Name Proposed

Classification

Evidence Cited

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

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

(nonylphenyl)-omega-

hydroxy-

P-Exp (Preliminary) No referenceprovided 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 theenvironment 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 todays methods, these early studies showed that the NPE and OPE surfactants easily

underwent primary biodegradation. Although these early studies focused largely on elimination of parentmaterial, 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

APE i h i i ll d d F h k l d f h h i i f


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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 chainisomers. 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. Basedon 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 orclassification 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 byformation of carbon dioxide. According to Environment Canadas 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


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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 alsoinvestigated. 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 byday 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% to48.4% 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 theGledhill 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.


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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 guidanceprovided 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 forpersistence.

Simulation Tests with Linear Test Materials

In addition to screening tests, the biodegradability of members of the family of NP, OP, and theirethoxylates 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 ofthe 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 7days 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


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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 g/kg, dry wt. By day 30, concentrations of 4-n-NP in both planted and unplanted soils

were reduced by 97.8% and 98.2%, respectively. 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 authors

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 ofparent compound remaining in the soil in both treatments, greater than 90% of the added nonylphenolremained after 10 days incubation; the levels were reduced to below the detection limit (


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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 withsediments (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 atconcentrations 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 withthe 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 minoramounts 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 30C with unacclimated

sediment, acclimated sediment, at various temperatures (20 to 50C), and with or without shaking. For

static systems at 30C, 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 30C using both unacclimated and acclimatedsediment, 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. Thetime 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 20C, 40C, and 50C. For NPE1, half-

lives of 57.8, 6.0, and 4.7 days were reported for 20C, 40C, and 50C. An additional important factor

in the conduct of these tests was whether or not the test vessels were shaken or static. Shaking the flasks


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to 13 days at 18C, and 3 to 7 days at 13C, 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 half-

lives at 20 to 22.5C were 2.5 to 35 days, at 18C were 10 to 35 days, and at 13C 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 13C), 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 butdid 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 11C. 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


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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 g/g, dry weight. Lysimeters were placed outdoors to receive naturallighting, 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 20day 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 half-

lives (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 anapproximate 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 columnincreased 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


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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 on-site 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 molarbasis. 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 migratedthrough 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


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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 thesimulation 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 (13C and 38%, respectively) and highest salinity used and so some bias to slowerdegradation 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 half-

lives 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 thatalkylphenols 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


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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 Model84852-15-3 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 fieldinvestigations, 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.

Th t k t t t d t b 45 L -1 d -1 d th ti t t t 0 16


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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 5micrograms/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 anuptake rate constant of 193 mL-g-1-day-1 and a depuration rate constant of 0.56 day-1. In summary, flow-

through 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 measuredconcentrations 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


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Lewis and Lech (1996) studied the uptake, disposition, and persistence of14C-NP from water in the

rainbow trout (Oncorhynchus mykiss). The bioconcentration of14C-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 of14C-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 of14C-nonylphenol in specific tissues was

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

NP 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 of14C-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 RN84852-15-3.

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 usewith 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


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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


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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 infield 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 Agencys 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.

F th i th t k titi i d t b li b h i f li d b h d OP i ld


25/42

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 of14

C-t-OP in a semi-static system, and were analyzed after 5, 12, and19 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 stateconditions 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 towaterborne 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


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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 availableon 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 fortert-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

T h fi ld BAF l d b Ah l l (1993) i bl i h h h BCF


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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 notmeet 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 persistenceand 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 t < 182 days). This conclusion is supported by additionalresults 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 thecriteria 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/OPECAS RN Decision Evidence/Basis for Decision

Nonylphenol

104 40 5 Not P Exp (Ying and Kookana 2003)

26027 38 3 N t P E G S /A l ith 9016 45 9


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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 PNot B

Exp-Group; Synonym/Analogy with 9016-45-9Exp-Group Synonym/Analogy with 9016-45-9

68412-54-4 Not P

Not B


Exp-Group Synonym/Analogy with 9016-45-9

127087-87-0 Not PNot B

Exp-Group; Synonym/Analogy with 9016-45-9Exp-Group Synonym/Analogy with 9016-45-9

Octylphenol

140-66-9 Not PNot 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


27193-28-8 Not P

Not B



Octylphenol Ethoxylates9002-93-1 Not P

Not B


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; Analogy with NPE; Phys/chem. properties

VII REFERENCES


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