DRAFT DO NOT CITE OR QUOTE

Page 1

United States Office of Chemical 

Environmental Protection Safety and Pollution 

Agency Prevention

TSCA Workplan Chemical Risk Assessment 

HHCB

1,3,4,6,7,8-Hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-γ-2-

benzopyran

CASRN: 1222-05-5

H3C CH3

CH3

CH3

H3C

O

CH3

December 2012

NOTICE

This document is an External Review Draft. This information is distributed solely for the 

purpose of pre-dissemination peer review under applicable information quality guidelines. It 

has not been formally disseminated by EPA. It does not represent and should not be construed 

to represent any Agency determination or policy. It is being circulated for review of its 

technical accuracy and science policy implications.
DRAFT DO NOT CITE OR QUOTE

Page 2

Table of Contents

AUTHORS / CONTRIBUTORS / ACKNOWLEDGEMENTS / REVIEWERS....................................................5

GLOSSARY OF TERMS AND ABBREVIATIONS .......................................................................................6

EXECUTIVE SUMMARY .......................................................................................................................8

CHAPTER 1: BACKGROUND AND SCOPE ...........................................................................................10

INTRODUCTION...................................................................................................................................10

A. SCOPE OF ASSESSMENT FOR HHCB .......................................................................................................12

B. CONCEPTUAL MODEL FOR ENVIRONMENTAL ASSESSMENT OF HHCB..........................................................12

CHAPTER 2: SOURCES AND ENVIRONMENTAL FATE..........................................................................15

INTRODUCTION...................................................................................................................................15

A. PHYSICAL AND CHEMICAL PROPERTIES OF HHCB.....................................................................................15

B. ENVIRONMENTAL FATE........................................................................................................................17

CONCLUSIONS OF ENVIRONMENTAL FATE....................................................................................................21

C. PRODUCTION AND USES.......................................................................................................................21

CONCLUSIONS ON PRODUCTION VOLUME AND USES.....................................................................................26

CHAPTER 3: ENVIRONMENTAL ASSESSMENT...................................................................................27

A. ENVIRONMENTAL EXPOSURE ASSESSMENT..............................................................................................27

B. ENVIRONMENTAL HAZARD ASSESSMENT.................................................................................................32

C. ENVIRONMENTAL RISK CHARACTERIZATION.............................................................................................39

D. KEY SOURCES OF UNCERTAINTY AND DATA LIMITATIONS ..........................................................................41

CONCLUSIONS OF ENVIRONMENTAL ASSESSMENT.........................................................................................42

REFERENCES ....................................................................................................................................43

APPENDIX A: HUMAN HEALTH TOXICITY STUDIES, BIOMONITORING DATA, AND RISK ASSESSMENTS61

A-1. HUMAN HAZARD CHARACTERIZATION.................................................................................................61

A-2. HUMAN BIOMONITORING.................................................................................................................67

A-3. SUMMARY OF 2008 EU HUMAN HEALTH RISK ASSESSMENT..................................................................70

A-4. KEY SOURCES OF UNCERTAINTY AND DATA LIMITATIONS ON HUMAN HEALTH...........................................74

A-5. CONCLUSIONS OF HUMAN HEALTH ASSESSMENT ..................................................................................75

APPENDIX B: CHEMICAL SYNTHESIS OF HHCB...................................................................................77

APPENDIX C: HHCB DIASTEREOISOMERS (#1 TO #6) AND RELATED STRUCTURAL ANALOGS (#7 TO #15)

........................................................................................................................................................78

APPENDIX D: IUR DATA FOR HHCB...................................................................................................81

APPENDIX E: MODELED RELEASE ESTIMATES ACCORDING TO STAGE OF PRODUCTION AND USE.......83

APPENDIX F: ENVIRONMENTAL FATE AND BIOACCUMULATION STUDY SUMMARIES ........................92

APPENDIX G: ENVIRONMENTAL MONITORING DATA ANALYSIS........................................................95
DRAFT DO NOT CITE OR QUOTE

Page 3

List of Tables

Table 2-1. Physical-Chemical Properties of HHCB......................................................................................16

Table 2-2. HHCB Degradation Half-Lives and Half-Disappearance Times for Environmental Media. .......19

Table 2-3. BCFs and BAFs (L/kg ww) of HHCB in Various Aquatic Species.................................................20

Table 2-4. BCFs (L/kg ww) of HHCB in Benthic and Terrestrial Species. ....................................................21

Table 2-5. Major US Manufacturers or Importers of HHCB.......................................................................22

Table 2-6. US Production/Import Volume of HHCB...................................................................................23

Table 2-7. Estimated Distribution of Fragrance Oils By Use. .....................................................................24

Table 2-8. Cosmetic Product Types and Upper Levels of Fragrance Incorporation...................................25

Table 3-3. Aquatic Toxicity Data for HHCB—Acute Toxicity. .....................................................................32

Table 3-4. Aquatic Toxicity Data for HHCB—Chronic Toxicity. ..................................................................34

Table 3-5. Sediment Toxicity Data for HHCB..............................................................................................36

Table 3-6. Soil Toxicity Data for HHCB. ......................................................................................................37

Table 3-7. COCs for Environmental Toxicity...............................................................................................39

Table 3-8. Calculated Environmental Risks of HHCB..................................................................................40

Table A-1. Summary of Human Health Hazard Information......................................................................65

Table A-2. Human Biomonitoring Data for HHCB. .....................................................................................68

Table B-1. HHCB Diastereoisomers and Related Structural Analogs. ........................................................78

Table D-1. IUR National HHCB Information.a .............................................................................................81

Table D-2. HHCB IUR Process and Industry Information............................................................................81

Table D-3. HHCB IUR Consumer Information.............................................................................................82

Table D-4. IUR Company Site Information.................................................................................................82

Table E-1. Summary of Estimated Environmental Releases. .....................................................................84

Table E-2. Geographic Distribution for Facilities under NAICS 32562 Toilet Preparation Manufacturing.87

Table E-3. Geographic Distribution for Facilities under NAICS 325611 Soap and Other Detergent 

Manufacturing........................................................................................................................88

Table E-4. Geographic Distribution for Facilities under NAICS 325612 Polish and Other Sanitation Good 

Manufacturing........................................................................................................................89

Table G-1. Measured Concentrations of HHCB in Wastewater. ................................................................97

Table G-2. Measured Concentrations of HHCB in Surface Water............................................................100

Table G-3. Measured Sediment Concentrations at Locations in the US..................................................102

Table G-4. Measured Concentrations of HHCB in Biosolids.....................................................................104

Table G-5. Measured Concentrations of HHCB in Soil. ............................................................................107

Table G-6. Measured Concentrations of HHCB in Wildlife Biota.............................................................109

Table G-7. Summary of Box Plots for USGS HHCB Data...........................................................................112

Table G-8. USGS Medium and Site Codes. ...............................................................................................113
DRAFT DO NOT CITE OR QUOTE

Page 4

List of Figures 

Figure 1-1. Conceptual Model for HHCB Assessment................................................................................13

Figure 2-1. Chemical Structure of HHCB (HSDB, 2007)..............................................................................15

Figure 2-2. Chemical Synthesis of HHCB. ...................................................................................................77

Figure G-1. Monitoring Data Summary from USGS NWIS for HHCB Concentrations in Water, Filtered. 114

Figure G-2. Monitoring Data Summary from USGS NWIS for HHCB Concentrations in Water, Unfiltered.

..............................................................................................................................................115

Figure G-3. Monitoring Data Summary from USGS NWIS for HHCB Concentrations in Solids................115
DRAFT DO NOT CITE OR QUOTE

 Page 5

Authors / Contributors / Acknowledgements / Reviewers

This report was developed by the Office of Pollution Prevention and Toxics (OPPT), Office of 

Chemical Safety and Pollution Prevention (OCSPP). The 2012 workplan risk assessments were 

prepared based on existing data and any additional information received up to June 15, 2012.

EPA Assessment Team

Lead: Susan A. Laessig, OPPT/RAD

Team: Katherine Anitole, OPPT/RAD

Fred Arnold, OPPT/EETD

Lynne Blake-Hedges, OPPT/EETD

Bob Boethling, OPPT/EETD

Ellie Clark, OPPT/CCD

Majd El-Zoobi, OPPT/EETD

Jonghee Koh, visiting scientist KEPA

Chuantung Lin, OPPT/EETD

Andrea Pfahles-Hutchens, OPPT/RAD

Elizabeth Sommer, OPPT/EETD

Eva Wong, OPPT/EETD

Acknowledgements

Portions of this document were developed with support from:

Contractor name (contract number) Task

Abt Associates (EP-W-08-010) Production and Uses

Eastern Research Group (EP-W-10-014) Releases

Versar (EP-W-10-005) Monitoring

SRC, Inc. (EP-W-09-027) Ecotoxicity

Internal Peer Reviewers

H. Kay Austin, OPPT/EETD

Stan Barone Jr., OPPT/RAD

Todd Stedeford, OPPT/RAD

External Peer Review - A peer review panel is being arranged for this influential workplan 

assessment product based upon need and following Agency peer review guidance. The format 

will be a teleconference of an ad hoc panel meeting consisting of independent experts.
DRAFT DO NOT CITE OR QUOTE

 Page 6

Glossary of Terms and Abbreviations

AHTN 6-Acetyl-1,1,2,4,4,7-hexamethyltetraline

BB Benzyl benzoate

BAF Bioaccumulation factor

BCF Bioconcentration factor

CASRN Chemical Abstract Service Registry Number

CBI Confidential Business Information

ChV Chronic value

COC Concentration of concern

DEP Diethyl phthalate

dw Dry weight

E2 Estradiol

EC European Commission

ECHA European Chemicals Agency

EPA Environmental Protection Agency

EPCRA Emergency Planning and Community Right-to-Know

EPI Estimation Programs Interface 

ESD Emission Scenario Document

EU European Union

EUSES European Union System for the Evaluation of Substances

FIFRA Federal Insecticide, Fungicide and Rodenticide Act

GC/MS Gas chromatography/mass spectrometry

GD Gestation day

GLP Good Laboratory Practice

HERA Human and Environmental Risk Assessment Project

HHCB 1,3,4,6,7,8-Hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-γ-2-benzopyran

HPLC High performance liquid chromatography

HPV High production volume

IFRA International Fragrance Association

IPM Isopropyl myristate

IUR Inventory Update Reporting Rule

k1 Uptake rate constant

k2 Elimination rate constant

Kd Sorption coefficient

kg Kilogram(s)

KOC Organic carbon-normalized sorption coefficient

KOW Octanol:water partition coefficient

lbs Pound

LDR Larval development ratio

LOAEC Lowest-observed-adverse-effect concentration

LOAEL Lowest-observed-adverse-effect level

LOEC Lowest-observed-effect concentration
DRAFT DO NOT CITE OR QUOTE

 Page 7

Log KOC Logarithmic organic carbon partition coefficient

Log KOW Logarithmic octanol:water partition coefficient

LRL Laboratory reporting level

LT-MDL Long-term method detection level

lw Lipid weight

MOS Margin of safety

NAICS North American Industry Classification System

ND Not detected

NHANES National Health and Nutrition Examination Survey

NOAEC No-observed-adverse-effect concentration

NOAEL No-observed-adverse-effect level

NOEC No-observed-effect concentration

NWIS National Water Information System

NWQL National Water Quality Laboratory

OCSPP Office of Chemical Safety and Pollution Prevention

OECD Organisation for Economic Co-operation and Development

OPPT Office of Pollution Prevention and Toxics 

OSPAR Oslo Paris

PNEC Predicted no-effect concentration

REACH Registration, evaluation, authorisation and restriction of chemicals

RHOearthworm Bulk density of earthworm

RAR Risk assessment report

RQ Risk quotient

SCAS Semi-continuous activated sludge

SIAM SIDS Initial Assessment Meeting

SIAR SIDS Initial Assessment Report

SIDS Screening information data set

SSO Sanitary sewer overflow

TG Test Guideline

TLC Thin-layer chromatography

TSCA Toxic Substances Control Act

UF Uncertainty factor

UFtotal Total uncertainty factor

US United States

USGS United States Geological Survey

wt% Weight percent

ww Wet weight

WWTP Wastewater treatment plant

yr(s) Year
DRAFT DO NOT CITE OR QUOTE

 Page 8

Executive Summary

1,3,4,6,7,8-Hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-γ-2-benzopyran (HHCB) was assessed 

as part of the United States (US) Environmental Protection Agency (EPA) Existing Chemicals 

Management Program. HHCB was selected for this assessment based on a moderate hazard 

rating for potential developmental toxicity, a high ranking for potential exposure, presence in 

human samples, and estimated high releases to the environment. HHCB exhibits moderate 

environmental persistence and moderate bioaccumulation potential. 

This assessment evaluated the environmental risk to the aquatic and terrestrial environments

from the use of HHCB as a fragrance ingredient in consumer and commercial products. 

Scenarios of interest included discharge of wastewater to surface water or application of 

biosolids to land. The endpoints of interest were acute and chronic toxicity to aquatic species, 

chronic toxicity to sediment organisms, and chronic toxicity to terrestrial invertebrates and 

plants. A detailed review of the available human health data, including a risk assessment 

performed by the European Union, indicate negligible risks of concern. A 2-year cancer study is 

not available for HHCB. As an alternative, EPA evaluated the weight of the evidence from 

mutagenicity, subchronic, and endocrine studies, to support the conclusion that cancer data are 

not a critical data need. Therefore, human health risk was not assessed further. 

HHCB is one of the most widely used polycyclic musk fragrance ingredients in a range of 

consumer products including perfumes, cosmetics, shampoos, lotions, detergents, fabric 

softeners, and household cleaners. HHCB is not produced in the US and is imported by four 

importers. The imported volume of HHCB in the US has been reported in the range of 1 to 

10 million pounds (lbs) per year from 1990 to 2006. Within this range, the volume of use in the 

US has increased steadily since 2000, suggesting that polycyclic musks are not being replaced 

with other synthetic musks, as may be occurring in Europe.

In the US, imported HHCB is compounded into fragrance oils, which are blended into end-use 

products and sold for both commercial and consumer use. An estimated 90 to 94 percent of 

the HHCB use-volume is released to municipal wastewater. HHCB is moderately persistent in 

soil and sediment, has low mobility in soil, suboptimal removal in wastewater treatment plant 

(WWTP) (with removal primarily through sorption to sludge), and is moderately 

bioaccumulative in some species.

Monitoring data for HHCB in wastewater, surface water, sediment, and soil were used to

characterize potential exposures in this assessment. The concentrations of HHCB measured in 

the environment are likely to represent multiple sources and uses. The exposure assessment

analyzed US environmental monitoring data from the literature and from the US Geological 

Survey (USGS). More than 6800 sampling measurements were available for effluent, surface 

water and sediment from the USGS National Water Information System (NWIS) database. 

These values represent a range of dates and locations across the US. In addition, soil 

monitoring data was available from 2 published studies. 
DRAFT DO NOT CITE OR QUOTE

 Page 9

Ecotoxicity studies for HHCB have been conducted in fish, aquatic invertebrates, aquatic plants, 

sediment invertebrates, soil invertebrates, and terrestrial plants. HHCB has high acute and

chronic toxicity for aquatic organisms with effects on adult and embryo survival, behavior, 

respiration, and larval development. This assessment considered only direct exposure to 

aquatic species, sediment-dwelling species, terrestrial species, and plants via contact with 

contaminated water, sediment, or soil. Based on half-lives in water, sediment, soil, and 

continuous wastewater inputs, exposure to organisms was considered continuous. Standard 

uncertainty factors were used to calculate a concentration of concern (COC) for each endpoint.

Risk for each environmental scenario was calculated using the risk quotient (RQ), the ratio of 

measured environmental concentration to the toxicological COC for each environmental 

scenario. An RQ < 1 was considered a negligible risk of concern, whereas an RQ > 1 was 

considered as a potential risk of concern. To incorporate the range of measured environmental 

concentrations, the RQ was calculated using the mean, maximum, and/or 95th percentile of 

measured environmental concentrations for surface water and sediment. Available soil 

concentrations from two published studies were used to calculate an RQ for the maximum 

measured soil concentration. 

The acute and chronic risk for aquatic species did not exceed the COCs at maximum measured 

surface water concentrations. The chronic COC was exceeded (i.e., RQ > 1) for sedimentdwelling species at the upper range of measured environmental concentrations for less than 

1.5% of the more than 600 sediment measurements from the USGS. Though the data were 

limited, chronic risk for terrestrial invertebrates and plants did not exceed the COC at maximum 

measured soil concentrations. At the highest measured soil concentration, the COC was not 

exceeded for terrestrial invertebrates and plants. The available soil monitoring data were 

limited, but EPA determined the potential risk of concern associated with application of 

biosolids was negligible; if concentrations are marginally higher in some locations than shown 

in the available studies, the risk may exceed the COC for soil invertebrates or plants. The 

calculated RQs indicate negligible risks of concern.

The use of measured environmental concentrations, detailed fate characterization, and 

availability of robust ecotoxicity data for multiple species are strengths of this assessment. 

Monitoring data reported in the literature and in the USGS NWIS data for surface water and 

sediment represent different sampling times and locations across the US. The HHCB 

concentrations are assumed to represent reasonable estimates of environmental exposures. 
DRAFT DO NOT CITE OR QUOTE

Page 10

Chapter 1: Background and Scope

INTRODUCTION

The chemical 1,3,4,6,7,8-Hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-γ-2-benzopyran (HHCB; 

CASRN 1222-05-5) was identified for assessment as part of the United States (US) 

Environmental Protection Agency’s (EPA)’s Existing Chemicals Management Program 

(http://www.epa.gov/oppt/existingchemicals/pubs/workplans.html). 

HHCB is a synthetic polycyclic musk used as an ingredient in a wide range of consumer products 

including perfumes, cosmetics, shampoos, lotions, detergents, fabric softeners, air fresheners, 

and cleaning agents. HHCB has moderate hazard potential for developmental toxicity1

, high 

potential exposure, and has been detected in humans. EPA estimated potentially high releases 

of HHCB to the environment. HHCB presents moderate environmental persistence and 

moderate bioaccumulation potential. 

HHCB has not been the subject of specific regulatory activity in the US, Canada, or the European 

Union (EU). HHCB is listed on the TSCA Inventory, is a High Production Volume (HPV) chemical 

substance, and is reported for Inventory Update Reporting (IUR), but is not otherwise subject to 

specific TSCA regulations. It is not listed on the Emergency Planning and Community Right-toKnow (EPCRA) Section 313 List of Toxic Chemicals. HHCB is approved as an inert ingredient for 

fragrance and non-food use in pesticide products under the Federal Insecticide, Fungicide and 

Rodenticide Act (FIFRA).

2 In Canada, HHCB is listed on the Domestic Substances List and is 

characterized as “inherently toxic to aquatics,” but did not meet Canada’s full categorization 

criteria and was not recommended for further assessment. In Europe, HHCB is registered under 

the Registration, Evaluation, Authorisation and Restriction of Chemicals (REACH) regulation.

Previous assessments in Europe, Sweden, and Australia have evaluated the environmental risk 

from HHCB (Balk and Ford, 1999a; 1999b; Balk et al., 2004; HERA, 2004; OSPAR, 2004; EC, 2008; 

SWECO, 2008; Langdon et al., 2010). HHCB was assessed for human and environmental risk in 

Europe in a 2008 EU Risk Assessment Report (RAR) (EC, 2008), which concluded: “There is at 

present no need for further information and/or testing and no need for risk reduction measures 

beyond those which are being applied already.” No potential risk concerns were identified for 

surface water and sediment in Sweden (SWECO, 2008) or for the marine environment (OSPAR, 

2004). Potential risk concerns were identified for certain highly contaminated sediments in 

 1 Human health risks from HHCB uses were not characterized in this assessment. After a detailed review of the 

available data, there is a low concern for human health hazards, including developmental toxicity, based on the 

weight of the evidence from repeated-dose toxicity studies (see Appendix A).

2 HHCB is commonly diluted in diethyl phthalate (DEP), benzyl benzoate (BB), or isopropyl myristate (IPM) before 

being added to fragrance formulations Rimkus (1999). However, EPA recently proposed a ban for DEP as an inert 

ingredient in pesticide products (US EPA, 2012c).
DRAFT DO NOT CITE OR QUOTE

Page 11

Europe (HERA, 2004) and for HHCB in surface runoff from biosolids-amended land in Australia 

(Langdon et al., 2010). In addition, a screening-level risk assessment for California recently 

identified a risk concern for HHCB in surface water and recommended monitoring in inland 

waterways (Anderson et al., 2012). 

EPA reviewed previous assessments and literature published since 2005, the date of the last 

literature review indicated in the EU RAR (EC, 2008). This review of HHCB chemistry, uses, 

sources, fate, exposure, and hazard of HHCB to both humans and the environment indicated:

• HHCB has an aggregate annual import volume of between 1 and 10 million lbs. HHCB 

was imported to the US by three companies in 2006. A fourth company listed for 

reporting on the IUR also is assumed to import.

• HHCB is continuously released to the environment, mainly in municipal wastewater and 

sludge.

• HHCB is not readily biodegradable and binds strongly to sediment, soil, and sludge.

• HHCB is moderately persistent and moderately bioaccumulative in the environment 

even though HHCB is biotransformed to less toxic, less bioaccumulative, polar 

metabolites.

• HHCB has high acute and chronic toxicity for aquatic organisms. Chronic effects of 

HHCB in aquatic organisms include effects on adult and embryo survival, behavior, 

respiration, and larval development. Mechanistic studies show that HHCB may

interfere with steroid hormone receptors, metabolism, and membrane transporters in 

fish and invertebrates, but not in mammals. 

• Acute and chronic toxicity of HHCB has been studied in benthic and terrestrial 

invertebrates and in crop plants.

• Monitoring data for the US are available from the US Geological Survey (USGS) National 

Water Information System (NWIS) database and the published scientific literature.

• After a detailed review of the available data, there is a low concern for human health 

hazards, based on minimal developmental toxicity at relatively high oral exposures and 

no developmental toxicity at levels several times greater than detected in human 

breast milk (see Appendix A for a summary of human health data).

• Human dermal and inhalation exposures are expected, and HHCB is present in human 

blood and breast milk (see Appendix A for a summary of human biomonitoring data).

• The human health hazards and risks of HHCB have been previously assessed by other 

reliable entities. The conclusions of the 2008 EU RAR are summarized in Appendix A. 

• A 2-year cancer study is not available for HHCB. As an alternative, EPA evaluated the 

weight of the evidence from mutagenicity, subchronic, and endocrine studies, to 

support the conclusion that cancer data are not a critical data need.
DRAFT DO NOT CITE OR QUOTE

Page 12

A. SCOPE OF ASSESSMENT FOR HHCB

Based on its moderate persistence and bioaccumulation, constant and high release to the 

environment, and acute/chronic toxicity to aquatic organisms, this assessment evaluated

potential risk to aquatic and terrestrial environments from the use of HHCB as a fragrance 

ingredient in consumer and commercial products. Data are unavailable on the amount of HHCB 

used in separate product categories or specific products. The use of products containing HHCB, 

such as personal care products and cleaning products, results in “down-the-drain” release into 

the wastewater stream. HHCB removal during treatment of municipal wastewater influent is 

approximately 50 to 85 %; therefore, HHCB can be present at high levels in aqueous 

wastewater effluent as well as in biosolids (WERF, 2007). 

Due to the availability of a sufficient quantity of monitoring data for the US, EPA analyzed

measured levels of HHCB and data from ecotoxicological studies of acceptable quality to assess 

the risks of HHCB to the environment.

Three key questions were addressed in this assessment:

1. How do levels of HHCB measured in surface water in the US compare to levels that produce

adverse effects in aquatic invertebrates, fish, or plants?

2. How do levels of HHCB measured in sediment in the US compare to levels that produce

adverse effects in sediment-dwelling invertebrates?

3. How do levels of HHCB measured in soil in the US compare to levels that produce adverse 

effects in soil invertebrates and plants?

The assessment did not address environmental risks from specific uses because the sources of 

HHCB to the environment are widespread and diffuse. In addition, ecological effects on 

mammals or birds were not assessed and population- or community-level ecological endpoints 

were beyond the scope of this assessment.

B. CONCEPTUAL MODEL FOR ENVIRONMENTAL ASSESSMENT OF HHCB

The scenarios evaluated in this assessment are 1) release of HHCB in wastewater to surface 

water or 2) release of HHCB in biosolids to land. The ecological endpoints evaluated for HHCB 

are acute and chronic toxicity to aquatic species, chronic toxicity to sediment organisms, and 

chronic toxicity to terrestrial invertebrates and plants. 

The exposure scenarios for EPA’s HHCB assessment are depicted in the Conceptual Model

(Figure 1-1) and described below. 
DRAFT DO NOT CITE OR QUOTE

Page 13

Figure 1-1. Conceptual Model for HHCB Assessment.

*Includes all fragranced products such as soaps, detergents, fabric softeners, shampoos, cosmetics, and 

cleaners. 

**EPA estimates that between 23,000 and 75,000 sanitary sewer overflows (SSOs) occur each year in the US, 

resulting in releases of untreated wastewater between 3 billion and 10 billion gallons. 

After commercial or residential use of consumer products containing HHCB, the majority of

HHCB enters the wastewater stream. Other minor releases to the environment also were 

estimated (see Appendix E: Release Estimates). Due to the ubiquity of fragranced products, 

consumer uses are a continuing source of HHCB to wastewater. When HHCB enters

wastewater treatment plants (WWTPs), the chemical is expected to partition to organic-rich 

particles during wastewater treatment based on the physical-chemical properties. However, 

HHCB removal during treatment in WWTPs is suboptimal (see Chapter 2, Fate in Wastewater) 

and HHCB is present in both wastewater effluent and solids (WERF, 2007). Biosolids can be 

subsequently land-applied and result in HHCB loadings to soil.

During the wastewater treatment process, influent is separated into two components: 

wastewater effluent and solids, known as either sewage sludge or biosolids. Wastewater 

effluent is released to surface water at outfall sites which are located in freshwater, estuarine, 

and marine environments. Under this scenario, HHCB is released in the effluent and is diluted 

in streams, lakes, estuaries, and oceans where aquatic species may be exposed. In addition, 

thousands of sanitary sewer overflows (SSOs) occur each year in the US, resulting in releases of 

billions of gallons of untreated wastewater directly to surface water. When HHCB has not been 

removed during the treatment process, adsorption to suspended solids in the surface water 

and partitioning to sediment is expected. Under these conditions, HHCB is subsequently

available for uptake by fish, aquatic invertebrates, aquatic plants, and benthic organisms,

leading to potential exposure and toxicity.

HHCB also is found in the solid component of the waste stream (sludge or biosolids). This

organic carbon-rich material is disposed of by landfill or incineration, or may be utilized for land 
DRAFT DO NOT CITE OR QUOTE

Page 14

application to enhance physical soil properties as well as plant yield (Correa et al., 2005).

WWTPs need to dispose of a continuous supply of biosolids and about half of all sludge as 

biosolids is applied to agricultural land each year.

3 Once biosolids containing HHCB are applied

to land, HHCB has a long half-life in soil and may be available for uptake by plants and soil 

invertebrates, leading to potential exposure and toxicity.

This environmental assessment evaluated HHCB exposures for the scenarios described above

based on available monitoring data for the US and assessed the potential ecological hazards for 

aquatic, benthic, and soil species by calculating concentration of concerns (COCs). The

potential risks to aquatic, benthic, and soil species were characterized by calculating risk 

quotients (RQs) for each route of environmental exposure (i.e., water, sediment, and soil)

based on measured concentrations. Because exposure to HHCB is likely to occur in all types of 

aquatic environments, no distinction in risk was made between freshwater and marine species.

 3 About 7.2 dry US tons of biosolids were beneficially used or disposed in 50 states in 2004. Of that total, 

approximately 55 percent was applied to soils for agronomic, silvicultural, and/or land restoration purposes, or 

was likely stored for such uses. The remaining 45 percent was disposed of in municipal solid waste landfills, 

surface disposal units, and/or incineration facilities. (NEBRA, 2007).
DRAFT DO NOT CITE OR QUOTE

Page 15

Chapter 2: Sources and Environmental Fate

INTRODUCTION

To assess the environmental risks of HHCB, a firm understanding of the physical and chemical 

properties of HHCB, sources of HHCB related to production and uses, and fate of HHCB in the 

environment is important. These topics are introduced in Chapter 2 and discussed in relation to 

the environmental risk characterization in Chapter 3.

A. PHYSICAL AND CHEMICAL PROPERTIES OF HHCB

HHCB is an almost colorless viscous liquid with a musk-like odor. The polycyclic chemical 

structure for HHCB is shown in Figure 2-1.

H3C CH3

CH3

CH3

H3C

O

CH3

Figure 2-1. Chemical Structure of HHCB (HSDB, 2007).

Synonyms for HHCB include:

• 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethyl-cyclopenta-gamma-2-benzopyran

• 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethyl-cyclopenta-[γ]-2-benzopyran

• 6-oxa-1,1,2,3,3,8-hexamethyl-2,3,5,6,7,8-hexahydro-1H-benz[f]indene

• Galaxolide

• Galaxolide 50

• Galaxolide 50BB

• Galaxolide 50IPM

• Galaxolide White

• Abbalide

• Pearlide

The manufacturing process for HHCB is described in Appendix B. Technical HHCB consists of a 

mixture of isomers that are unspecified. There are four diastereoisomers (i.e., an isomer that 

differs in the spatial arrangement of atoms in the molecule, but is not a mirror image) of HHCB

that are known to exist: two (-)/4S isomers (4S, 7R & 4S, 7S) have the characteristic odor, and 

the other two (+)/4R isomers (4R, 7R & 4R, 7S) have little to no odor (DrugLead, 2009; 
DRAFT DO NOT CITE OR QUOTE

Page 16

http://www.druglead.com/cds/hhcb.html). Additionally, there are many other structural 

isomers or analogs related to HHCB; a complete list is provided in Appendix C. The HHCB 

commercial product is diluted (65 percent wet weight [ww]) in diethyl phthalate (DEP)4

, benzyl 

benzoate (BB), or isopropyl myristate (IPM) prior to compounding and formulation into 

products (HERA, 2004b).

The physical and chemical properties of HHCB are shown in Table 2-1.

Table 2-1. Physical-Chemical Properties of HHCB

Molecular formula C18H26O

Molecular weight 258.44

Physical form Colorless liquid; highly viscous liquid at 20 °C and 1,013 hPa with a musk 

odor

Melting point -10 to 0 °C; -20 °C

Boiling point 320 °F at 760 mmHg (converted from 160 °C at 4 hPa)

Vapor pressure 0.0727 Pa (5.47 × 10-4 mmHg) @ 25 °C (measured; Organisation for 

Economic Co-operation and Development [OECD] 104)

Logarithmic octanol:water 

partition coefficient (log KOW) 5.3 ("slow stirring" method); 5.9 @ 25 °C (measured; OECD 117)

Water solubility 1.65 mg/L at pH 7; 1.99 mg/L at pH 5; 1.69 mg/L at pH 9 (at 25 °C);

2.3 mg/L at 20 °C (OECD 105)

Flash point 144 °C

Source: (HSDB, 2007).

 

4 EPA recently proposed a ban for DEP as an inert ingredient in pesticide products (US EPA, 2012c).
DRAFT DO NOT CITE OR QUOTE

Page 17

B. ENVIRONMENTAL FATE

The environmental fate of a compound is important to understanding the impact of potential 

exposures to specific environmental media (e.g., water, sediment, soil, and plants) and target 

organisms of concern. HHCB enters domestic wastewater treatment and is subjected to 

removal by a variety of processes. HHCB and any degradation products generated in 

wastewater treatment may be discharged to receiving waters, disposed of in biosolids, or 

volatilized via air stripping during wastewater treatment. This section summarizes current 

knowledge of the persistence, bioaccumulation, and bioconcentration of HHCB in wastewater 

treatment and the environment including biological and abiotic reactions and environmental 

distribution (See Appendix D for detailed study summaries).

ENVIRONMENTAL PERSISTENCE

Fate in Wastewater Treatment 

RadiolabeledHHCB (25 μg/L) was added to activated sludge collected from three different 

WWTPs and the disappearance and formation of metabolites were monitored over time 

(Langworthy et al., 2000; as cited in EC, 2008); HHCB disappeared with the subsequent 

appearance of more polar metabolites. The half-life for HHCB in activated sludge was 21 hours 

(first-order rate constant = 0.033 hours-1), and 85 percent of the HHCB initially present 

disappeared in 150 hours (6.25 days). No significant volatilization was detected. A more recent 

study reported that the first-order rates for loss of HHCB in activated sludge were 0.010 and 

0.021 hours-1 with 5 and 25 μg/L starting concentrations, respectively. Results showed 

disappearance at 50 percent, with disappearance times of 69 and 33 hours, respectively 

(Federle et al., 2002; as cited in EC, 2008). In a series of die-away tests on HHCB with a starting 

concentration of 17.4 μg/L in activated sludge, the overall half-lives derived for parent 

compound disappearance were 10 to 15 hours (see Table 2-2) (Schaefer, 2005; as cited in EC, 

2008). For the above studies, the geometric mean for activated sludge half-disappearance time 

was 22.5 hours. This corresponds to “moderate-to-slow” biodegradation in activated sludge; 

see guidance in the Estimation Programs Interface (EPI) Suite v4.10, 

http://www.epa.gov/opptintr/exposure/pubs/episuite.htm and 

http://www.epa.gov/oppt/exposure/pubs/halflife.htm. 

HHCB is expected to partition strongly to solid phases based on its high measured log KOW of 5.9 

(see Rimkus, 1999 for a summary of values for musks) and estimated soil/sediment organic 

carbon partition coefficient (KOC; log basis) of 4.1-4.3 (KOCWINTM program v2.00; in EPI SuiteTM 

v4.10 ; download from http://www.epa.gov/opptintr/exposure/pubs/episuite.htm) (US EPA, 

2011). These predictions are supported by numerous measured sorption coefficients for a 

variety of matrices including soil, sediment, suspended particulate matter in water, and primary 

and secondary sludge (EC, 2008). Values for both Kd (sorption coefficient) and KOC (organic 

carbon-normalized sorption coefficient) are generally in the range of 3 to 4 on a logarithmic 

scale. This means that HHCB will have low mobility in soil, bind strongly to benthic and 

suspended sediment, and be substantially removed by sorption to sludge in WWTPs. The 

Human and Environmental Risk Assessment Project (HERA, 2004a) summarized both theoretical 
DRAFT DO NOT CITE OR QUOTE

Page 18

and measured percent removal values for European and US activated sludge-based treatment 

systems. Observed removal was generally >50 percent (usually >80 percent), and removal 

percentage correlated well with removal of total suspended solids (Simonich et al., 2002). 

Waste (settled) activated sludge in WWTPs is generally sent to anaerobic digesters to reduce 

sludge volume and organic load before disposal of the concentrated material. However, EPA 

did not locate data on degradation of HHCB under anaerobic conditions in sludge digesters. 

Fate in Water

HHCB was not readily biodegradable (0 percent in 28 days) in an OECD 301B test (Jenkins, 1991; 

as cited in EC, 2008) and a sealed vessel headspace test (OECD 310), the latter with adapted 

inoculum (King, 1994; as cited in EC, 2008). HHCB appears not to undergo direct photolysis in 

water at a significant rate (Buerge et al., 2003).

In batch experiments (Langworthy et al., 2000; Federle et al., 2002; as cited in EC, 2008),

effluent diluted in river water was spiked with radiolabeled HHCB at concentrations of 1.0 or 

0.5 μg/L. The observed degradation half-life of HHCB was 33 to 43 hours, depending on the 

starting concentration. After 12.5 days, about 40 percent of the radioactivity was lost by 

volatilization. In a more recent presentation reporting on the same or extended experiments, 

with 0.5 μg/L of HHCB in river water, the half-life for disappearance of the parent substance 

was 43 hours. In a series of die-away tests, an overall half-life of 100 hours was derived for 

disappearance of HHCB in river water (Schaefer et al., 2005; as reported in EC, 2008). A 

significant fraction (up to 16 percent) was volatilized after 28 days. Volatilization half-lives were

on the order of days to weeks. 

Overall, in the modified river die-away tests of Langworthy et al. (2000), Federle et al. (2002), 

and Schaefer (2005) (all as cited in EU, 2008), which were designed to simulate river water 

conditions immediately below discharge of treated sewage from activated sludge treatment 

plants, half-lives for HHCB disappearance of 2 to 4 days were observed (see Table 2-2). 

Degradation is expected to be slower in streams that do not receive effluents from WWTPs or 

that are less impacted by such effluents.

Fate in Soil and Sediment

The fate of HHCB was studied using microcosms containing oak forest and agricultural soils, 

river sediment, and agricultural soil that routinely applies sludge from a domestic WWTP

(Envirogen, 1998; as cited in EC, 2008). For the four different soil types, an average of 14

percent of the HHCB initially present remained after one year. Half-lives calculated from the 

measured degradation rate constants were 105, 95, 239, and 79 days for sludge-amended soil, 

forest soil, agricultural soil, and river sediment, respectively; the average half-life for the four 

soils was 128 days. 

DiFrancesco et al. (2004) studied the dissipation of fragrance materials in sludge-amended soils 

in a one year die-away experiment with four different soils, with and without spiking of the test 

materials. The four soils included sandy agricultural soil; silty agricultural soil; clayey soil; and 

highly weathered oxide-rich soil. The initial HHCB concentrations in spiked soil were 6 and 13
DRAFT DO NOT CITE OR QUOTE

Page 19

mg/kg soil, whereas the levels in unspiked soil were 0.1 to 0.27 mg/kg. In all soils, the 

concentrations rapidly decreased. After one month, the concentrations of HHCB in the four 

soils were 30 to 90 percent of the initial concentration, and after 90 days, they ranged from 8 to 

60 percent of the initial concentration. During the three months when the soil was frozen,

concentrations of HHCB were stable. After one year, the residues ranged from below 10 to 14

percent of the initial concentration. The rate of dissipation was higher in the soils with lower 

content of organic material. Loss processes could have included volatilization and leaching as

well as biotransformation. The EC (2008) calculated half-disappearance times based on these 

data and reported values of 141 and 144 days for the spiked and unspiked sludge-amended 

soils, respectively. 

Table 2-2. HHCB Degradation Half-Lives and Half-Disappearance Times for Environmental Media.

Source

Degradation Half-Life or Half-Disappearance Timea

Water Sediment Soil Sludge Sludge-Amended 

Soil

Langworthy et al.

(2000) (as cited in 

EC, 2008)

33 hours 21 hours

Federle et al.

(2002) (as cited in 

EC, 2008)

43 hours 33-69 hours

Schaefer (2005) (as 

cited in EC, 2008)

100 hoursb 10-15 hours

Envirogen (1998)

(as cited in EC, 

2008)

79 days 95 days (forest 

soil)

105 days

Envirogen (1998)

(as cited in EC, 

2008)

239 days 

(agricultural 

soil)

DiFrancesco et al.

(2004)

141-144 days

a

Half-disappearance time differs from degradation half-life in that it includes all mechanisms contributing to 

disappearance (e.g., volatilization in addition to degradation). b

Water with 10 mg/L suspended solids. 

Observed soil and sediment half-lives consistently exceeded 60 days (Table 2-2). Field 

measurements on sludge-amended soil indicated that HHCB disappeared almost completely 

from soil within one year. The half-life based on unfrozen conditions in sludge-amended soil 

studies was around 140 to 144 days (DiFrancesco et al., 2004). The residues in soil after one 

year ranged from below 10 to 14 percent of the initial concentrations. In the EU RAR (EC, 
DRAFT DO NOT CITE OR QUOTE

Page 20

2008), a half-life of 105 days in the sludge-amended soil was deemed most relevant for 

modeling the fate of HHCB in soil using the European Union System for the Evaluation of 

Substances (EUSES) model, while 79 days was noted for the sediment (Envirogen, 1998; as cited 

in EC, 2008). EPA agrees that these values are reasonable for modeling and assessment 

purposes.

Fate in Air

Aschmann et al. (2001) studied the atmospheric oxidation (photooxidation) of HHCB and 

reported a rate constant for hydroxyl radical oxidation of k = 2.6 ± 0.6 × 10-11 cm3 molecule-1

second-1

. This leads to a calculated atmospheric half-life of 3.4 hours for HHCB (based on a 

12-hour daylight period), and suggests that long-range atmospheric transport is unlikely to be 

an important process for HHCB.

BIOACCUMULATION AND BIOCONCENTRATION

Based on available data, HHCB is considered a moderate concern for bioaccumulation. 

Bioaccumulation data exist for several aquatic species; most BCF and BAF values are in the 

range of 20 to 260. These values are consistent with low concern. However, BCF values of 

1,584 for bluegills and 2,692 for Lumbriculus indicate moderate bioaccumulation potential. 

These studies, together with monitoring data in biota, suggest that HHCB is not subject to 

biomagnification. Table 2-3 summarizes published bioaccumulation data for aquatic species.

Table 2-3. BCFs and BAFs (L/kg ww) of HHCB in Various Aquatic Species.

Species Common Name BCF BAF

Scardinius erythrophthalmus Rudd 20

Tinca tinca Tench 510

Carassius carassius Crucian carp 580

Anguilla anguilla Eel 862 290

Dreissena polymorpha Zebra mussel 620

Lepomis macrochirus Bluegill sunfish 1,584

Danio rerio Zebrafish 620

Source: Dietrich and Hitzfield (2004).

Jager (1998) (as cited in EC, 2008) reported a calculated BCF for earthworms (Lumbricus 

terrestris) (Table 2-4). The steady-state concentration of a substance in earthworms is assumed 

to be proportional to the soil pore water concentration. Bioconcentration can be described as a 

hydrophobic partitioning between the pore water and the phases inside the organism, and a 

BCF can be calculated using an equation given by Jager (1998) (as cited in EC, 2008):

BCFworm = (0.84 + 0.012 × Kow)/RHOearthworm
DRAFT DO NOT CITE OR QUOTE

Page 21

Table 2-4. BCFs (L/kg ww) of HHCB in Benthic and Terrestrial Species.

Species Common Name BCF Reference

Chironomus riparius Midge 85-138 Artola-Garicano et al. (2003)

Lumbriculus 

variegatus

Blackworm 2,692 Artola-Garicano et al. (2003)

Lumbricus terrestris Earthworm 2,395a Jager (1998)

a

Calculated.

CONCLUSIONS OF ENVIRONMENTAL FATE

HHCB is hydrophobic and partitions strongly to solid phases. Therefore, HHCB is removed in 

WWTPs predominantly due to sorption to sludge and biodegradation, but the degree of 

removal is considered suboptimal. HHCB has low persistence in water and is moderately 

persistent in soil and sediment. The lack of measured half-lives for anaerobic conditions 

(benthic sediments and sludge digesters) was considered a significant data gap. Several 

measured BAFs and BCFs for HHCB suggest bioaccumulation potential in the low range, but 

HHCB is considered moderately bioaccumulative for some species.

C. PRODUCTION AND USES

This section discusses the US production volumes and uses specifically for HHCB and more 

generally for fragrances. HHCB is a High Production Volume chemical that is used in a vast 

number of cleaning and personal care products. Data on production volume and uses are

amenable for determining release estimates to the environment. These data are also useful for 

understanding potential exposure routes and pathways through which HHCB may enter the 

environment through industrial and consumer applications.

Production

HHCB and AHTN (6-acetyl-1,1,2,4,4,7-hexamethyltetraline; CASRN 1506-02-1; also CASRN 

21145-77-7) are the most widely used of the polycyclic musks. Together, HHCB and AHTN 

represent 90 percent of the US polycyclic musk market (HERA, 2004a), with HHCB being the 

most widely consumed of the polycyclic musks (EC, 2008). Four US firms reported nonconfidential business information (CBI) IUR data for HHCB in 2006 (US EPA, 2012d)

5 . A list of 

 5 These four producers may be an underestimate because a production site is only required to submit a Form U 

(the IUR reporting instrument) if it produces or imports more than 25,000 pounds of a chemical during the 

reporting year.
DRAFT DO NOT CITE OR QUOTE

Page 22

the US importers of HHCB is provided in Table 2-5. Of the four companies listed, three reported 

importing HHCB, and one did not indicate if it is a manufacturer or importer. EPA assumes that 

this company is an importer because HHCB is not manufactured in the US (IFRA, 2012a). HHCB 

is imported in liquid forms at a maximum concentration of ≥90 percent (US EPA, 2012d).

Table 2-5. Major US Manufacturers or Importers of HHCB.

Chemical Company Reported IUR 

Data in 2006

Manufacturer or 

Importer

HHCB, Galaxolide, 

Musk GX, Abbalide Berje, Inc. Yes Importer

International Flavors & Fragrances, Inc. Yes Importer

Quest International Flavors and Fragrances Yes Importer

S C Johnson & Son, Inc. Yes No dataa

a

‘No data’ indicates that data are not available on whether the company is a manufacturer or importer.

Source: US EPA (2012d). 

Between 1996 and 2000, the US consumption of synthetic musk fragrances increased by 

25 percent, from about 5,200 to 6,500 tons (10.4 to 13 million lbs), while the consumption of 

fragrance chemicals only grew by 15 percent (Somogyi and Kishi, 2001; as cited in Peck et al., 

2006). Global musk production increased by 12.5 percent between 1987 and 1996 and during 

this time period, production shifted from nitro musks, such as musk xylene, to polycyclic musks 

(Rimkus, 1999; as cited in Luckenbach and Epel, 2005). This shift, reflected in decreasing 

production rates of nitro musks and increasing production rates of polycyclic musks (Hutter et 

al., 2005), is expected to continue for several reasons. In June 2009, the European Chemicals 

Agency (ECHA) recommended xylene musk for authorization under REACH (ECHA, 2009a). 

Additionally, in June 2009, the International Fragrance Association (IFRA) voluntarily banned 

musk xylene through the IFRA Standards, part of the fragrance industry’s global self-regulatory 

program contained in the IFRA Code of Practice. The Code of Practice is mandatory for IFRA 

members, and membership accounts for approximately 90 percent of the global volume of 

fragrance materials (IFRA, 2011). Finally, in February 2011, ECHA added musk xylene to its 

authorization list under REACH, which means that musk xylene can be used only in cases where 

an authorization has been granted for a specific use (European Union, 2011). 

Data from IFRA have shown an increase in the volume of HHCB (point estimate opposed to 

range estimate) used in the US from the yrs (yrs) 2000 to 2011. IFRA estimated US HHCB use 

volume to be approximately 1,275 tons (2.8 million lbs) per yr in 2000 and slightly under 1,400 

metric tons (3.1 million lbs) per yr in 2004 (IFRA, 2012a). More recently, IFRA provided 

additional use volume estimates in the US for 2008 and 2011 of approximately 1,600 and 

1,700 metric tons (3.52 and 3.74 million lbs) per yr, respectively (IFRA, 2012b).

6 According to 

 6 IFRA’s letter to EPA dated March 30, 2012 indicated that the 2008 use volume of HHCB in the US was 

approximately 1,300 tons per year. This estimate was revised by IFRA in an email dated June 29, 2012 to 1,600 

tons per year in 2008.
DRAFT DO NOT CITE OR QUOTE

Page 23

IFRA, the increase in HHCB use in the US is not a reflection of increased use of this particular 

musk over others, but a reflection of increased market demand for fragranced consumer 

products, which has expanded the market for HHCB and other musk chemicals (IFRA, 2012c).

It is unclear whether the North American market for synthetic musks will eventually experience 

the same shift, as in the EU, from nitro musks to polycyclic musks (Gatermann et al., 1999; as 

cited in Peck et al., 2006). At present, it appears as though the US production volume (including 

both manufacture and import) trends for HHCB and musk xylene in the US are consistent for 

both chemicals. For example, the annual (non-CBI) production volume of HHCB has ranged 

between 1 and 10 million lbs since 1990 (Table 2-6). In addition, with the exception of 2002, 

the non-CBI IUR production volume for musk xylene has been consistently reported to be 

<500,000 lbs (US EPA, 2012d; 2012e). Therefore, and in contrast to the EU, the steady range of 

production levels for both HHCB and musk xylene suggests that there has not been a significant 

shift away from nitro musks to polycyclic musks in the US.

Table 2-6. US Production/Import Volume of HHCB.

Chemical

Production/Import Volume (in Thousands, K, or Millions, M, of lbs)

Year

1986 1990 1994 1998 2002 2006

HHCB, Galaxolide, 

Musk GX, Abbalide

500K-1M 1M-10M 1M-10M 1M-10M 1M-10M 1M-10M

Musk Xylene 500K-1M 10K-500K 10K-500K 10K-500K 10K-500K <500K

Source: US EPA (2012d); US EPA (2012e).

Uses

Musks are considered important compounds to the fragrance industry because of their unique 

odor properties, ability to improve the fixation of fragrance compounds, and ability to bind 

fragrances to fabrics (Sommer, 2004). The function of HHCB in fragrance formulations is as 

both a fragrance and a fragrance enhancer (ECHA, 2009b). 

HHCB is used as a fragrance ingredient in cleaning and personal care products because it is 

alkali-stable and does not discolor in light (Ash et al., 2009; OSPAR, 2004; IFF, 2012; Fahlbusch

et al., 2012). HHCB and other musks provide a unique, dominant scent in products. Because 

HHCB binds fragrances to fabric and skin, the scent is balanced and longer lasting (OSPAR, 

2004). HHCB is often used in laundry detergent fragrances because it is one of the few 

chemicals that can leave a small residual fragrance on cloth after washing and can cover up 

odors from the detergent itself as well as from dirt in the wash solution (Schreiber, 2004). 

Synthetic musks, including HHCB, also may be used to mask chemical odors and can be found in 

products labeled “unscented,” but do not seem to be added to products labeled “fragrance 

free” (Potera, 2007).
DRAFT DO NOT CITE OR QUOTE

Page 24

The EU RAR (EC, 2008) estimated that in the EU, approximately 86 percent of manufactured 

HHCB is used as part of a fragrance mixture, and 14 percent is used directly in the bulk 

formulation of products. HHCB is, however, generally estimated to make up only two to four 

percent of fragrance compounds because fragrance mixtures can contain from 50 to 

300 ingredients (HERA, 2004a; Bickers et al., 2003). As shown in Table 2-7, the majority of 

fragrance oils (25 percent) are used in detergents, followed by fabric softeners (14 percent) and 

personal care products (13 percent), whereas on a weight percent (wt%) basis, perfume 

extracts have the highest level of fragrance incorporation at 20 wt% (Table 2-8). Given the 

importance of HHCB as a fragrance ingredient, EPA expects that the use of HHCB has a similar 

distribution pattern as fragrance oils in general with the largest percentage of volume used in 

detergents, fabric softeners, and personal care products.

Table 2-7. Estimated Distribution of Fragrance Oils By Use.

Use Fragrance Oils (percent)

Detergent 25

Fabric softeners 14

Personal care 13

Bath and shower 10

Hair care 10

Soaps 9

Industrial and household cleaner 8

Other 6

Fine fragrances 5

Source: Balk et al. (2001) (as cited in OSPAR, 2004).
DRAFT DO NOT CITE OR QUOTE

Page 25

Table 2-8. Cosmetic Product Types and Upper Levels of Fragrance Incorporation.

Cosmetic Product Type Fragrance Level (wt%)a,b

Perfume extracts 20.00

Toilet waters 8.00

Fragranced cream 4.00

Bath products 2.00

Toilet soap 1.50

Shower gel 1.20

Antiperspirant/deodorant 1.00

Hair spray 0.50

Shampoo 0.50

Body lotion 0.40

Face cream 0.30

a

Industry survey on typical quantities used per application of different cosmetics (COLIPA, 1987; as cited in Cadby

et al., 2002). b

Estimates of typical fragrance levels in different products and maximum likely proportion of fragrance remaining 

on skin after normal product use (RIFM, 1996; as cited in Cadby et al., 2002).

Source: Cadby et al. (2002).

Several recent studies have measured the levels of HHCB in a variety of consumer products. 

One study found HHCB in 55% of the personal care products tested, including 100% of 

deodorants tested (Roosens, et. al., 2007). A 2002 survey found that HHCB may be found at 

levels between 0.02 and 0.90% in household cleaning products (IFRA, 2002 as reported in HERA, 

2004b). Another recent study by Reiner and Kannan (2006) analyzed products for 

concentrations of HHCB and a degradant (breakdown product), HHCB-lactone. In this study,

HHCB concentrations were from <5 to 646,000 mg/kg sample and the degradant, the HHCBlactone, was <5 mg/kg in the same products. This study also showed that products within a 

single class might contain differing levels of HHCB. For example, it was noted that HHCB 

concentrations in two different laundry detergents were 79 and 84,900 mg/kg sample. 

A wide range of consumer products contain HHCB. However, these studies indicate that levels 

varied significantly across and even within product types. This variation is not unexpected given 

the complexity of fragrance mixtures which often contain hundreds of constituents. 
DRAFT DO NOT CITE OR QUOTE

Page 26

CONCLUSIONS ON PRODUCTION VOLUME AND USES

Use volume of HHCB has increased gradually during 2000-2011 due to increased demand for 

fragranced products. HHCB is used as part of fragrance mixtures in products such as detergents 

and fabric softeners, and personal care products. The concentration of HHCB in any of these 

products ranges widely across product types as well as within single product categories (e.g., 

laundry detergent). 
DRAFT DO NOT CITE OR QUOTE

Page 27

Chapter 3: Environmental Assessment

INTRODUCTION

This assessment evaluated the environmental risk to the aquatic and terrestrial environments

from the use of HHCB as a fragrance ingredient in consumer and commercial products. The 

HHCB environmental risk assessment is composed of the following: (A) an exposure 

assessment; (B) an ecological hazard assessment; (C) an environmental risk characterization;

and (D) a discussion of uncertainties and data limitations. Environmental risks were assessed 

by comparing levels of HHCB measured in the environment to the COCs determined from 

aquatic, benthic, and terrestrial toxicity studies. 

A. ENVIRONMENTAL EXPOSURE ASSESSMENT

For this assessment, environmental monitoring data on measured levels of HHCB in 

wastewater, surface water, sediment, soil, and biota were used to characterize environmental

exposure to HHCB.

ESTIMATED ENVIRONMENTAL RELEASES OF HHCB 

Releases to the environment potentially occur throughout the life cycle of HHCB. In the US, the 

life cycle of HHCB consists of processing of imported HHCB as an ingredient in the compounding 

of fragrance oils, blending of these fragrance oils into end-use products for commercial and 

consumer use, consumer or commercial use of products, and disposal of HHCB-containing 

products to the environment. Compounding of fragrance oils and formulation of end-use 

products occurs at industrial sites. A small amount of HHCB is released from these sites to the 

environment, including to water, mostly as a result of cleaning of compounding and blending 

equipment and cleaning of containers used to transport material. Specifically, the amount 

released from industrial sites was estimated at a maximum of 10 percent of the use volume 

(see Appendix E). EPA assumed that the entire HHCB content of all of the various types of enduse products was released down-the-drain after use. Therefore, this assessment estimates that

90% of HHCB is released to municipal wastewater after commercial and consumer use of 

products containing HHCB.

MEASURED LEVELS OF HHCB IN THE ENVIRONMENT

Studies published between 2004 and May 2012 that reported measurements of HHCB in the 

environment in the US were reviewed for this assessment. Additionally, monitoring data in US 

waterways were collected from the USGS NWIS database up to May 2012. All data collected 

from 2001-2012 contained in three USGS NWIS parameter codes (62075, 62823, 63209) for 

HHCB were included in this assessment. This database includes monitoring data as well as data 

collected for studies of organic anthropogenic waste indicators, organic wastewater 

contaminants, and biologically-active chemical mixtures in urban streams. 
DRAFT DO NOT CITE OR QUOTE

Page 28

Although ten times more data was available from the USGS as compared to data from

published literature studies for effluent, surface water and sediment, the measured HHCB 

concentrations were within the same order of magnitude. EPA assumed that data from 

scientific publications were comparable to the USGS data and were used along with the USGS 

data as the basis of the exposure and risk characterizations. Although sampling techniques 

varied depending upon the environmental media, gas chromatography/mass spectrometry 

(GC/MS) was the primary analytical methodology employed in the detection and quantification 

of HHCB across all media. Concentrations of HHCB found in wastewater, surface water, 

sediment, biosolids, soil, and aquatic biota are summarized below and in Table 3-2. More 

detailed information on the monitoring data is provided in Appendix G. 

Wastewater

In published studies, measured concentrations of HHCB in US WWTP influent and effluent 

varied widely. Influent values ranged from 43 to 12,700 ng/L and effluent values ranged from 

10 to 13,000 ng/L (Table 3-2; see also: Appendix G, Table G-1). Effluent concentration was

dependent on the type of waste treatment process employed and varied depending upon the 

season or month when the sample was collected. Reported values from a Canadian study 

confirmed that effluent values were process dependent and concentrations were seasonally 

dependent and within an order of magnitude of those measured in the US (Smyth et al., 2008). 

The mean HHCB concentrations in wastewater effluent from stream and outfall discharge sites 

calculated from the USGS NWIS data (more than 40 data points) were 980-1180 ng/L. The 

highest values were measured at outfall sites (95th percentile: 3400 ng/L). The mean effluent 

concentrations were comparable at stream and outfall sites, but the range of concentrations 

was greater at outfall sites. 

Surface Water

In published studies, surface water concentrations of HHCB were found to range from nondetect (ND) to 1,600 ng/L and were dependent upon their proximity to WWTP outfalls (Table 3-

2; see also: Appendix G, Table G-2). The lowest reported concentrations were ND levels at 

locations upstream from WWTPs; higher concentrations of HHCB were found downstream from 

the WWTPs. A maximum value of 1,600 ng/L was reported by Barber (2011) for a location at 

the North Shore Channel of the Chicago River at an area impacted by water reclamation plant 

effluent. These data support the assertion that the efficiency of HHCB removal in WWTPs plays 

a significant role in the observed surface water concentrations.

From more than 6000 data points for HHCB in surface water (collected by the USGS from 

locations in 46 states), the mean calculated value for HHCB concentration in surface water was 

< 1.1 μg/L for all sites; the highest concentrations were measured at outfall sites. The 95th

percentile groundwater concentrations at well sites and surface water concentrations at 

streams and lakes, reservoirs, and impoundment sites were ≤0.35 μg/L (Table 3-2; see also: 

Appendix G, Table G-3 and Figures G-2 and G-3). Lower concentrations were consistently 
DRAFT DO NOT CITE OR QUOTE

Page 29

recorded in surface water samples collected from streams and lakes, reservoirs, and 

impoundment sites when compared to values measured at outfall sites. 

Sediment

Sediment concentrations varied from ND to 388 ng/g dw in published studies (Table 3-2; see 

also: Appendix G, Table G-3). Aside from ND values reported in three tidal tributaries of the 

Chesapeake Bay, the lowest reported concentrations of HHCB (1.43 to 2.13 ng/g dw) were 

found in a lake in Texas, a non effluent impacted site. One of the highest concentrations (388 

ng/g dw) was reported along the upper Hudson River in New York, a river that receives treated 

wastewater discharge. Likewise, sediment along the Cuyahoga River in Ohio was found to have 

higher concentrations of HHCB in samples collected downstream of seven WWTPs (average= 

144 ng/g dw) when compared to those collected upstream of the same WWTPs (average = 37 

ng/g dw). These studies suggest that effluents from WWTPs are a major source of synthetic 

musks that enter the environment and are present in the sediments of various water bodies.

Over 600 bottom material measurements (collected from USGA locations in 25 states) were 

available from the USGS NWIS (Table 3-2; see also: Appendix G, Table G-7 and Figure G-3). The 

95th percentile concentration at lake/reservoir/impoundment sites was 213 μg/kg, and the 

mean concentration was 87 μg/kg. The 95th percentile concentration at stream sites was 200 

μg/kg, and the mean concentration was 68 μg/kg. The HHCB concentrations (5%ile, mean, and 

95%ile) were similar at both types of sites. 

Biosolids and Soil

Published concentrations of HHCB in biosolids ranged from <100 to >100,000 μg/kg dw (Table 

3-2; see also: Appendix G, Table G-4). These values were dependent on a number of factors, 

including: WWTP operations (e.g., municipal or industrial receiving waste stream), season, date 

of collection, location, population served, water volume, treatment type, and preparation 

methodologies prior to land application. Because these treatment variables cannot be 

separated from all values within the reported data set, it was not possible to assign sensitivity 

to a single property. However, EPA observed that the US values for measurements of biosolids 

were similar to those obtained in Ontario, Canada and showed similar differences depending 

upon location, treatment process, and season (Smyth et al., 2007). 

Soil concentrations of HHCB have been reported in a limited number of studies, with values 

ranging from <0.33 to 2,770 μg/kg dw in biosolid amended soils (Table 3-2; see also: Appendix 

G, Table G-5). The available data indicate that land application of treated waste water effluent 

or biosolids results in detectable quantities of HHCB in those soils.

Biota

The measured concentrations of HHCB in wildlife varied by location, species, and method of 

reporting (e.g., lipid weight, tissue weight, or wet weight) (Table 3-2; see also: Appendix G, 

Table G-6). Monitoring studies of aquatic biota were available only from the scientific 

literature, and the greatest number of sampling measurements was collected for fish (146 

sampling measurements). On a lipid weight basis, measured values across various fish species 
DRAFT DO NOT CITE OR QUOTE

Page 30

and locations ranged from <1 to 51.1 ng/g, whereas levels in wild caught and farm raised 

shrimp were 330 and 424 ng/g, respectively (Appendix G, Table G-6). Mean tissue weight 

concentrations of HHCB in different fish species ranged from 100 to 1800 ng/g (Appendix G, 

Table G-6). On a wet weight basis, HHCB has been reported in a number of different species 

and trophic levels, including: sea otters, seals, dolphins, sharks, whales, fish, etc. The reported 

values were relatively consistent across these species and ranged from <1 to 25 ng/g (Appendix 

G, Table G-6). 

USGS Data Analysis

In addition to reported quantitative values, the USGS dataset includes values that are reported 

as less than the USGS laboratory reporting level7 (LRL); data that are between the LRL and the 

long-term method detection level (LT-MDL); and data that are below the LT-MDL (Oblinger et 

al., 1999). The value of the LRL is reported with a “less than” remark code for samples in which 

the analyte was not detected. “Estimated” remark codes are added to all values falling outside 

the calibration range because of increased measurement uncertainty or values below the LTMDL determined using information-rich methods. 

For monitoring data sets where the geometric standard deviation was less than 3.0, values 

recorded as “less than LRL” or “estimated” were replaced by the LRL divided by the square root 

of two, as per the EPA OPPT’s guidance (USEPA OPPT guidance, 1994). Where the geometric 

standard deviation was greater than 3.0, EPA replaced values recorded as “less than LRL” or 

“estimated” with the LRL divided by two (USEPA OPPT guidance, 1994). This practice presents a 

conservative low end value protecting against false negative values; therefore, it should be 

noted that these values do not necessarily represent quantitative measured concentrations and 

are biased towards the LRL.

High quality monitoring data (i.e., greater than 6800 data points) were available from the USGS 

NWIS database for surface water and sediment. Data was available from the USGS NWIS for 

effluent to a lesser degree (47 data points) as shown in Table 3-2. Documented protocols and 

guidelines for sample collection and analysis were employed by the USGS such that these data 

sets are deemed to be acceptable for use in the exposure assessment. 

Conclusions of Environmental Exposure

HHCB has been detected and measured in wastewater, surface water, sediment, sewage 

sludge, soil, and aquatic biota in numerous studies. These data strongly suggest that HHCB is 

ubiquitous and potentially widespread in the environment. The data are assumed to be 

representative. 

 7 The LRL for water sampling was 0.5 μg/L for sampling dated 7/16/2001 – 9/30/2009 and was updated to 

0.05 μg/L for samples dating from 10/1/2009 to the present. 
DRAFT DO NOT CITE OR QUOTE

Page 31

Table 3-2. Summary of Measured Concentrations of HHCB in Environmental Media and Biota.

Environmental Compartment

Number of 

samples (all 

US studies)

Units Minimum Maximum Range of 

Means Reference

Wastewater ng/L

- Influent >135 43 12,700 420 - 9030 Hope (2012), Horii (2007), 

Chase (2012), Reiner 

- Effluent >135 10 13,000 21 - 5310 (2007), Osemwengie (2004)

- Effluent discharged to ocean n/a ND 2500 ---d Anderson (2012)

- Effluent (at stream or outfall) 47 90a 3400b 980 - 1180

USGS NWIS

(code 62075 and 62823)

Surface Water ng/L

- Freshwater 94 ND 1600 0.36 - 1600

Osemwengie (2004), 

Anderson (2012), Barber 

(2011), Peck (2004), Reiner 

(2011), Tomasek (2011), 

Klecka (2010), Chase (2012)

- Outfall, Stream, Lake, Reservoir, 

or Impoundment 4720 20a 2300b 120 - 1080

USGS NWIS 

(code 62075 and 62823)

- Groundwater (well) 1548 40a 350b 320 USGS NWIS (code 62075)

Sediment ng/g

- Freshwater >59 ND 388 1.1 - 144c

Tertuliani (2008), 

Klecka (2010), 

Sapozhnikova (2010), 

Chase (2012), Reiner 

(2011), Peck (2006)

- Bottom material (Lake, 

Resevoir, Impoundment, Stream) 606 17.68a 212.87b 67.99 - 87.46 USGS NWIS (code 63209)

Biosolids/Sludge >58 μg/kg 13 427,000

13 -

177,000c

Kinney (2006, 2008), 

Buyuksoonmez (2005), 

DiFrnacesco (2004)

Soil μg/kg

- Biosolids amended 6 1050 2770 ---d

Kinney (2008),

- Wastewater treated n/a <0.33 5.69 --- Chase (2012) d

Biota (aquatic) ng/g

- Fish (wet weight or lipid 

weight) 116 <1 51.1 3.0 - 4.8

Reiner (2011), 

Sapozhnikova (2010), 

Ramirez (2009), Kannan 

(2005), Osemwengie 

(2004), Kinney (2008)

- Fish (tissue conc.) 30 n/a 2100 100 - 1800

- Eel 1 125 125 ---d

- Zebra Mussel 4 10.3 19.3 ---d

- Shrimp 9 330 424 ---d

- Marine mammal 25 <1 25 2.8 - 14

- Shark 3 4.6 5.2 4.8

- Mink 4 2.2 5.3 3.7

- Bird 5 1.9 4.2 2.3 - 4.0

ND not detected

n/a not available

a 5th percentile value b 95th percentile valuec Range of averages

d Range of means not available
DRAFT DO NOT CITE OR QUOTE

Page 32

B. ENVIRONMENTAL HAZARD ASSESSMENT

The environmental hazard reviewed and summarized in this section is based on previous 

summaries including Balk and Ford (1999a); Balk and Ford (1999b); EU RAR (EC, 2008); HERA 

Project Report (HERA, 2004a); and Robust Summaries submitted under the US EPA HPV 

program (US EPA, 2003). In addition, a literature search was performed to identify peerreviewed articles published between 2005 and May 2012.

A summary of the available ecotoxicity data for HHCB that met US EPA HPV Program’s 

acceptability criteria for ecotoxicity studies (US EPA, 1999c) is provided below. Examples of the 

scientific quality criteria for the HPV Program are: a clear description of the endpoints, 

inclusion of appropriate controls, identification of test substance and test organism, stated 

exposure duration time and administration route, and transparent reporting of effect 

concentrations. EPA also accepted the classification of studies previously reviewed in the EU 

RAR (EC, 2008).

8

HHCB has been tested for acute and chronic aquatic, sediment, and soil toxicity. For aquatic 

toxicity studies that met EPA’s acceptability criteria, a hazard rating was assigned based on 

established EPA methodology for existing chemicals (US EPA, 2009a). Hazard rating methods

are not available for sediment or soil toxicity.

Acute Toxicity to Aquatic Species

A summary of the acute aquatic toxicity studies considered for this assessment is provided in

Table 3-3. The available studies in fish, invertebrates, and algae indicate that HHCB exhibits 

high acute toxicity in the environmental media and for the aquatic species tested. 

Table 3-3. Aquatic Toxicity Data for HHCB—Acute Toxicity.

Test 

Organism Species Test Type Endpoint Value (mg/L) Hazard 

Rating Reference

Bluegill 

sunfish

L. macrochirus OECD TG 

204

LC50 0.452 High Balk and Ford 

(1999b)

Fathead 

minnow

Pimphales 

promelas

OECD TG 

210

LC50 >0.140 High Balk and Ford 

(1999b)

Acute 96-hour EC50 0.39 High Dietrich and 

Hitzfield (2004)

 8 The US participated on the following international activities related to HHCB: (1) evaluating the screening 

information data set (SIDS); (2) participating at the SIDS Initial Assessment Meeting (SIAM) 28; and (3) accepting

the SIDS Initial Assessment Report (SIAR) under international agreements

(http://webnet.oecd.org/hpv/ui/handler.axd?id=b36caa92-a554-4853-b98d-744e72cb9d56).
DRAFT DO NOT CITE OR QUOTE

Page 33

Table 3-3. Aquatic Toxicity Data for HHCB—Acute Toxicity.

Test 

Organism Species Test Type Endpoint Value (mg/L) Hazard Rating Reference

Zebrafish D. rerio Acute LC50

96-hour LC50

96-hour EC50

0.452

>0.67

>0.67

High Dietrich and 

Hitzfield (2004)

Acute 48-hour 

NOEC/LOEC >1.0/NEP

a,b Moderate

Carlsson and 

Norrgren 

(2004)

Water flea Daphnia magna OECD TG 

202, part II, 

proposed 

updated 

version,

June 1993

48-hour ECR

50 0.282 High Balk and Ford 

(1999b)

Marine 

copepod

Acartia tonsa Acute, static 

ISO 1997 

standard

48-hour LCR

10

LCR

50

0.12P

a

0.47P

a

High Wollenberger

et al. (2003)

Estuarine 

Copepods

Norellia spinipes Life cycle 96-hour LCR

50 1.9P

a High Breitholtz et al.

(2003)

Freshwater 

mussel

Lampsilis cardium Acute, static 

renewal

24-hour LCR

50

48-hour LCR

50

1.0->1.75P

a

0.999->1.75P

a Moderate

Gooding et al.

(2006)

Acute, static 

renewal

96-hour ECR

50 0.153-0.831P

a High Gooding et al.

(2006)

Green algae Pseudokirchneriella 

subcapitata

Acute

OECD TG 

201

72-hour ECR

50 0.854 

(growth)

0.72 (biomass)

High Balk and Ford 

(1999b)

P

a

PNominal concentration. Note that these studies were not included for risk assessment purposes because of the 

uncertainty in actual HHCB exposure concentrations to the test organisms.

P

b

PNot established.

LOEC = lowest-observed-effect concentration; NOEC = no-observed-effect concentration

Because there were no discernible differences in the quality and acceptability of the available 

acute toxicity studies, the lowest acute toxicity value was selected to derive the acute COC. The 

48-hour LCR

50R in Daphnia (Balk and Ford, 1999b) was divided by an assessment factor 

(uncertainty factor [UF]) of 5 for invertebrates, according to Office of Pollution Prevention and 

Toxics (OPPT) guidance (US EPA, 2009b), to give an acute COC of 0.0564 mg/L or 56.4 ppm.

Chronic Toxicity to Aquatic Species

Studies with data of sufficient quality and acceptability on the chronic aquatic toxicity of HHCB 

are summarized in Table 3-4. Depending on the species and study type, HHCB received a 

hazard rating of moderate to high for chronic aquatic toxicity.
DRAFT DO NOT CITE OR QUOTE

Page 34

Table 3-4. Aquatic Toxicity Data for HHCB—Chronic Toxicity.

Test 

Organism Species Test Type Endpoint Value (mg/L) Hazard 

Rating Reference

Bluegill 

sunfish

L. macrochirus OECD TG 204 21-day 

NOEC/LOEC

ChV

0.093/0.182

0.13

Moderate

Moderate

Balk and Ford 

(1999b)

Fathead 

minnow

P. promelas OECD TG 210 36-day 

NOEC/LOEC 

ChV

0.068/0.140

0.098

Moderate

High

Balk and Ford 

(1999b)

Rainbow 

trout

Oncorhynchus 

mykiss

Chronic 21-day EC50 0.282 Moderate Dietrich and 

Hitzfield (2004)

Water flea D. magna Chronic 21-day 

NOEC/LOEC

ChV

0.111/0.205

0.151

Moderate

Moderate

Balk and Ford 

(1999b)

Marine 

Copepods

A. tonsa OECD draft 

Life cycle 

6-day 

NOEC/LOEC

EC10

EC50

ChV 

0.038/0.075

0.0438

0.115

0.0534

High

High

Bjornestad 

(2007) (as cited 

in EC, 2008)

Larval 

development

5-day 

NOEC/LOEC

5-day EC10

5-day EC50

0.30/NEa

0.037

0.059

Moderate

Highc

Highc

Wollenberger

et al. (2003)

d

Estuarine 

Copepods

N. spinipes Life cycle 22-day 

NOEC/LOEC

ChV

0.007/0.02b

0.012

Highc Breitholtz et al.

(2003)

d

Green algae P. subcapitata OECD TG 201 21-day 

NOEC/LOEC

ChV

0.201/0.466

0.306

Moderate

Moderate

Balk and Ford 

(1999b)

a

Not established.

b

Nominal concentration. Note that these studies were not considered for risk assessment purposes because of 

the uncertainty in actual HHCB exposure concentrations to the test organisms.

c

Exposures likely from both water and sediment. d

Complicated by the disappearance of HHCB from the water phase.

Chronic toxicity values based on nominal (not measured) concentrations from otherwise welldocumented studies were not considered for this risk assessment because the actual 

concentration of HHCB is uncertain. For example, the measured concentrations at the end of 

two chronic copepod tests (Breitholtz et al., 2003; Wollenberger et al., 2003) were only 2 to 19 

percent of the nominal concentration. Volatilization or adsorption to organic material likely 

accounts for the disappearance of HHCB from the water phase. 
DRAFT DO NOT CITE OR QUOTE

Page 35

The principal study, a partial life cycle test (OECD Draft Guideline) with the marine copepod,

Acartia tonsa, (Bjornestad, 2007) was designed to prevent the formation of organic debris and 

the loss of HHCB by sorption and volatilization. One day before addition of copepod eggs 

(80 eggs in each flask and 7,000 cells/mL of the algae Rhodomonas salina), Teflon-sealed flasks 

were pre-saturated with HHCB (94.8 percent purity) mixed with 0.1 percent of radio-labeled

[3-14C]-HHCB in ethanol at test concentrations of 37.5, 75, 150, 300, and 600 μg/L in natural 

seawater. Samples of the test solution were taken daily to measure the HHCB concentration.

After 5.5 days, the number of nauplii, copepodites, and non-hatched eggs were counted and 

the lengths of the larvae were measured. 

In this study, the actual test concentrations were >80 percent of nominal during the entire test.

Larval mortality was 27.1 and 29.1 percent in the control and solvent control, respectively, 

within the quality criterion for an acceptable study. In seawater and solvent controls, the larval 

development ratio (LDR; ratio of the number of copepodites to the sum of nauplii and 

copepodites) was 73 and 72 percent, respectively. Larval development followed a doseresponse relationship and was inhibited 7.5 percent at 37.5 μg/L HHCB and 27 percent at 75 

μg/L HHCB, as compared to the solvent control. Growth of nauplii was adversely affected at 

300 μg/L HHCB and no copepodites were observed. The EC10 (concentration affecting 10% of 

organisms) for LDR was 43.8 μg/L (95 percent confidence interval 30.1-55.3 μg/L), and the EC50

was 115 μg/L (131-153 μg/L).

Of the available studies, the lowest chronic value (ChV) identified was 0.0534 mg/L, as reported

by Bjornestad (2007) (as cited in EC, 2008), for effects on the larval development rate in marine 

copepods. To derive a chronic COC, the ChV was divided by an assessment factor (UF) of 10 for 

invertebrate species, according to OPPT guidance (US EPA, 2009b), to yield a chronic COC of 

0.00524 mg/L or 5.24 ppm for aquatic species. This value is similar to the predicted no-effect 

concentration (PNEC) of 0.0044 mg/L or 4.4 ppm derived in the EU RAR (EC, 2008) using the 

same principal study. 

Toxicity to Sediment-Dwelling Organisms

Toxicity studies in sediment-dwelling organisms including amphipods, midges, oligochaete 

worms, polychaete worms, and mud snails are summarized in Table 3-5. The most sensitive 

species identified was the New Zealand mud snail (P. antipodarum) (Pedersen et al., 2009). The 

effects of sediment-associated HHCB on growth, feeding rate, survival, and reproduction in the 

gastropod, Potamopyrgus antipodarum, were measured over 12 weeks in snails exposed to 0, 

0.1, 1, 10, 30, and 100 mg HHCB/g dw sediment. Adult survival and growth were not affected 

by HHCB. Juvenile growth and survival, reproduction, time to first reproduction, and adult 

feeding rate declined with increasing HHCB concentration. Effects on reproduction occurred at 

≥10 mg HHCB/kg dw of sediment; a no observed effect concentration (NOEC) of 1.0 mg 

HHCB/kg dw of sediment was reported, along with a calculated chronic value of 3.2 mg 

HHCB/kg dw of sediment (Pedersen et al., 2009).
DRAFT DO NOT CITE OR QUOTE

Page 36

A chronic COC for sediment was derived by applying an assessment factor of 10 to the chronic 

value for sediment dwelling organisms (3.2 mg HHCB/kg dw sediment), for a calculated chronic 

COC for sediment of 0.32 mg/kg dw (US EPA, 2009b).

Table 3-5. Sediment Toxicity Data for HHCB.

Test 

Organism Species Test Type Endpoint Value

(mg/kg dw) Reference

Amphipod Hyalella 

azteca

OECD TG 218 LC50

NOEC/LOEC

EC50

ChV

62.5

7.1/16.3

53.5

10.8

Egeler (2004) (as cited in 

EC, 2008)

Midge Cophixalus 

riparius

OECD TG 218 NOEC /LOEC

EC50

ChV

200/400

402

283

Egeler and Gilberg (2004)

(as cited in EC, 2008)

Oligochaete Lumbriculus 

variegatus

OECD TG 218 ChV 

(reproduction; 

nominal)

40.2

Egeler and Gilberg (2004)

(as cited in EC, 2008)

Polychaete Capitella sp. Lifetime NOEC/LOEC

ChV

1.5/26

6.2

Ramskov et al. (2009)

Sexual 

maturation

NOEC/LOEC

ChV

26/123

56.6

New Zealand 

mud snails

Potamopyrgus 

antipodarum

6-week 

exposure

NOEC/LOEC

ChV

10/30

17.3

Pedersen et al. (2009)

12-week 

exposure

NOEC/LOEC 

ChV

1.0/10

3.2

Reproduction/ 

development

NOEC/LOEC 

ChV

10/30

17.3

Toxicity to Terrestrial Organisms

The toxicity of HHCB to terrestrial organisms has been evaluated in springtails, earthworms, 

nematodes, and plants. A summary of these studies is provided in Table 3-6. The earthworm 

was the most sensitive terrestrial invertebrate species tested (Chen et al., 2011b). In the 

principal study, adult earthworms (Eisenia fetida) were exposed for 14 days to concentrations 

of 0, 100, 140, 196, 274.4, 384.2, 537.8, 752.9, 1,054.1, and 1,475.8 μg HHCB (99 percent

purity)/g (air-dried soil) in acetone (Chen et al., 2011b). The 7-day LC50 was reported as 489 μg 

HHCB/g and the 14-day LC50 as 392.4 μg HHCB/g. Significant effects on the growth rate of 

earthworms occurred at 100 mg HHCB/kg dw soil, the highest concentration, after a 28-day 

exposure to HHCB. A significant decrease in cocoon production occurred in earthworms 

exposed to 50 and 100 mg/kg dw soil. The calculated chronic value for earthworms of 38.7 

mg/kg dw soil.
DRAFT DO NOT CITE OR QUOTE

Page 37

Table 3-6. Soil Toxicity Data for HHCB.

Test 

Organism Species Test Type Endpoint Value Reference

Springtail Folsomia 

candida

ISO/CD 11267 NOEC/LOEC 

ChV

45/105 mg/kg dwa

68.7 mg/kg dwa

Balk and Ford 

(1999b)

Earthworm Eisenia fetida OECD TG 207 4-week NOEC 

ChV

45/105 mg/kg dwa

68.7 mg/kg dwa

OECD TG 207-

Filter paper 

contact 

screening test

24-hour LC50

48-hour LC50

72-hour LC50

32.60 μg/cm2

11.87 μg/cm2

6.14 μg/cm2

Chen et al. (2011b)

OECD TG 207 7-day LC50

14-day LC50

28-day NOEC/LOEC

ChV

489 mg/kg dw

392.4 mg/kg dw

30/50 mg/kg dw

38.7 mg/kg dw

Chen et al. (2011a)

Nematode Caenorhabditis 

elegans

Acute 

(lethality)

24-hr LC50 >194.6 mg/L Mori et al. (2006)

Chronic 

(growth, 

maturation, 

reproduction)

NOEC 9.8 mg/L

Wheat Triticum

aestivum

Seed 

germination

EC50

EC50

143.4 mg/L

422.3 mg/L

An et al. (2009)

Seedling 

development

NOEC/LOEC NEb/0.2 mg/L

NOAEC/LOAEC 

(seed germination) 

ChV (seed 

germination)

LC50 (seed 

germination)

EC50 (root 

elongation)

EC50 (shoot 

elongation )

77/194 mg/kg dw

122 mg/kg dw

846.6 mg/kg dw

2,123 mg/kg dw

928.5 mg/kg dw

Chen et al. (2010)

a

Measured concentration.

b

Not established.

LOAEC = lowest-observed-adverse-effect concentration; NOAEC = no-observed-adverse-effect concentration 

For plants, two studies were available that evaluated the toxicity of HHCB to wheat growth and 

development. Chen et al. (2010) identified a no-observed-adverse-effect concentration 

(NOAEC) for wheat seed germination of 77 mg/kg dw soil, with a calculated ChV of 122 mg/kg

dw soil.
DRAFT DO NOT CITE OR QUOTE

Page 38

Chronic COCs for soil invertebrates and terrestrial plants were derived by dividing the ChVs

reported by Chen et al. (2010) and Chen et al. (2011a) for earthworms and wheat respectively 

by UFs of 10. The resulting COCs for earthworms and wheat were 3.87 and 12.2 mg/kg dw soil.

Additional Studies

Studies were available on the effects of HHCB on endocrine mechanisms and other molecular 

pathways and are briefly summarized here.

The ability of HHCB to bind and interfere with steroid hormone receptors was investigated in 

liver cells of several aquatic species and also in cell lines. Weak estrogenic activity of HHCB 

(purity unknown) was observed in competitive estrogen receptor binding assays with rainbow 

trout, carp, and the amphibian, Xenopus leaves, and only at relatively high concentrations

(Dietrich and Chou, 2001; as cited in EC, 2008; Schreurs et al., 2002; Schreurs et al., 2004). 

HHCB showed no in vivo estrogenic activation of vitellogenin production in carp (Cyprinus 

carpio) (Seinen et al., 1999). HHCB did induce vitellogenin gene expression and protein 

synthesis in the livers of male medaka, Oryzias latipes, at 0.5 mg/L (nominal) (Yamauchi et al., 

2008). Antiestrogenic activity has been observed in various cell lines at lower concentrations

(Schreurs et al., 2002). HHCB produced a dose-dependent antagonistic effect on estrogen 

binding in juvenile zebrafish at concentrations at or below the no-observed-effect levels from 

early-life stage or growth studies in fish (Schreurs et al., 2004). No in vivo estrogenic activity 

was detected in a transgenic zebrafish assay, but anti-estogenic effects were observed in vivo

(Schreurs et al., 2004). HHCB (53.5 percent in DEP) was shown to inhibit estrogen-induced 

vitellogenin production in rainbow trout within the range of concentrations that have been 

detected in tissues of fish from contaminated locations (Simmons et al., 2010). A study of the 

interactions of HHCB with fish metabolic systems in carp showed that HHCB significantly 

inhibited Phase I and Phase II enzymes involved in the synthesis and metabolism of steroids

(Schnell et al., 2009).

The ecdysteroid agonist and antagonist activity of HHCB was assessed in an assay with the 

Drosophila melanogaster BII-cell line. HHCB did not show specific agonistic or antagonistic 

activity in this bioassay (Breitholtz et al., 2003). 

HHCB inhibited multixenobiotic resistance (mxr) transporters in gills of the marine mussel,

Mytilus californianus, an effect that may increase the accumulation of other toxicants in the 

tissues (Luckenbach et al., 2004). The inhibitory effects of HHCB were only partially reversed 

after a recovery period in clean seawater (Luckenbach and Epel, 2005). 

Summary of Environmental Hazard Assessment

Ecotoxicity studies for HHCB have been conducted in fish, aquatic invertebrates, aquatic plants, 

sediment invertebrates, soil invertebrates, and terrestrial plants. HHCB was found to have high 

acute toxicity and moderate to high chronic toxicity to aquatic organisms. The COCs derived for 

aquatic, soil, and sediment-dwelling organisms are summarized in Table 3-7. 
DRAFT DO NOT CITE OR QUOTE

Page 39

Table 3-7. COCs for Environmental Toxicity.

Acute, aquatic species 56.4 μg/L (56.4 ppm)

Chronic, aquatic species 5.24 μg/L (5.24 ppm)

Chronic, sediment-dwelling species 320 μg/kg dw

Chronic, terrestrial invertebrates 3.87 mg/kg dw

Chronic, terrestrial plants 12.2 mg/kg dw

C. ENVIRONMENTAL RISK CHARACTERIZATION

The goal of this environmental risk characterization was to determine whether there are risks 

to the aquatic, benthic, and terrestrial environment from the measured levels of HHCB found in 

surface water, sediment, or soil. The data for environmental monitoring and toxicity were used 

in this risk assessment to address the following key questions:

1. How do levels of HHCB measured in surface water in the US compare to levels that 

produce adverse effects in aquatic invertebrates, fish, or plants?

2. How do levels of HHCB measured in sediment in the US compare to levels that produce 

adverse effects in sediment invertebrates?

3. How do levels of HHCB measured in soil in the US compare to levels that produce 

adverse effects in soil invertebrates and plants?

To answer these key questions, environmental risk was characterized by calculating risk 

quotients or RQs (EPA, 1998; Barnthouse et al., 1982); the RQ is defined as:

RQ = Environmental Concentration / Effect Level

The environmental concentration for each compartment (i.e., water, sediment, soil) was 

determined based on measured concentrations of HHCB (Table 3-2). The COCs reported in

Table 3-7 served as the effect levels for aquatic, benthic, and terrestrial species. No distinction 

is made between marine and freshwater species because HHCB is expected to be widely 

distributed in the environment, there is not a large difference in sensitivity between marine and 

freshwater species, and the most sensitive test species is assumed to be representative of other 

species. 

Frequency, timing, and duration of exposure affects the potential for adverse effects in aquatic 

or terrestrial organisms. Aquatic exposure is assumed to be continuous based on constant 

wastewater production and a 2 to 4 day half-life of HHCB in water. This assumption is 

supported by the fact that HHCB is frequently measured in surface water downstream from 

WWTPs even though fate studies indicate that a significant percentage of HHCB may be lost to
DRAFT DO NOT CITE OR QUOTE

Page 40

volatilization and metabolism. Exposure of sediment dwelling-organisms to HHCB is also 

assumed to be continuous. 

In order to characterize the degree of potential risk due to HHCB exposure, RQs were calculated 

for the maximum levels and range of means from published studies, and also for the mean and 

95th percentile for surface water and sediment from the USGS NWIS data (Table 3-2). Surface 

water concentrations of HHCB at outfalls are assumed to represent the maximum 

environmental concentrations for HHCB. Monitoring data was limited for HHCB in US marine 

waters. The RQ for soil was calculated only for the maximum concentrations in biosolidamended and wastewater-treated soils. If the RQ is greater than one, there is a potential risk 

concern. 

The calculated RQs are presented as a range of values in Table 3-8 to reflect the range of 

concentrations of HHCB measured in the environment. 

Table 3-8. Calculated Environmental Risks of HHCB.

RQ

Maximum 

Published 

Concentration

RQ

Range of Mean 

Published 

Concentrations

RQ

95th

Percentile

USGS 

NWIS

RQ

Mean

USGS 

NWIS

RQ Range

Acute risk, aquatic 0.028 6.4 × 10-6 to 

0.028

0.041 0.006 to 

0.019

6.4 × 10-6 to 0.041

Chronic risk, aquatic 

species

0.31 6.9 × 10-5 to 

0.31

0.44 0.06 to 

0.21

6.9 × 10-5 to 0.44

Chronic risk, sedimentdwelling species 1.21a 0.003 to 0.45 0.67 0.21 to 

0.27

0.003 to 1.21

Chronic risk, terrestrial 

invertebrates

-biosolids-amended soil 0.72 No datab No data No data 0.72

-wastewater-treated soil 0.0015 No data No data No data 0.0015

Chronic risk, terrestrial 

plants

-biosolids-amended soil 0.23 No data No data No data 0.23

-wastewater-treated soil 0.0005 No data No data No data 0.0005 a

Values above 1.0 indicate excess risk and are shown in bold. b

‘No data’ indicates that individual concentrations or mean values were not available for that compartment.

The acute and chronic RQs for aquatic species were calculated from the USGS NWIS data using 

the concentrations measured in surface water near WWTP outfalls. Though the concentration 

of HHCB in effluent discharge may be diluted in surface water at the WWTP outfall, the

calculated aquatic RQs are considered conservative estimates for surface water based on actual 
DRAFT DO NOT CITE OR QUOTE

Page 41

surface water concentrations. Using this RQ approach, no risk concerns were identified for 

aquatic species (Table 3-8). 

The range of RQs for sediment-dwelling species was 0.12 to 1.22 (Table 3-8). The RQ of 

1.22 was calculated using the maximum published level of HHCB in sediment of 388 ng/g dw

(Table 3-2). This HHCB sediment concentration is from a single sample taken from the upper 

Hudson River near Troy, NY (Reiner, 2011) and a second sample (from the same study) taken 

near Albany, NY was 351 ng/g. Both of these values are above the 95th percentile value from 

the USGS monitoring data for bottom material. It should be noted that within the USGS data 

there are nine measured samplings (at 6 locations) above the COC for sediment dwelling 

species. These values are above the 95th percentile for bottom material and account for 1.5% 

of this data set. These values may represent “hotspots” for HHCB. Therefore, there is a 

potential risk concern for chronic effects of HHCB on sediment-dwelling species in a small 

percentage of highly contaminated areas.

Limited data are available on measured levels of HHCB in biosolids or wastewater- amended 

soil. The RQ was calculated for HHCB levels in a single study of agricultural soil amended with 

biosolids (see data in Table 3-2). The maximum RQs for this site were 0.72 for terrestrial 

invertebrates and 0.23 for terrestrial plants, which indicate no risk concerns for chronic effects 

of HHCB on terrestrial invertebrate species and plants using this limited data. 

D. KEY SOURCES OF UNCERTAINTY AND DATA LIMITATIONS

The strength of this assessment is the calculation of risk based on a considerable quantity of 

surface water and sediment monitoring data, a detailed fate characterization, and robust 

ecological toxicity studies. However, uncertainties need to be adequately characterized to 

understand the limitations of the approach. This section discusses the qualitative limitations of 

the monitoring data, ecological toxicity data, and risk characterization.

Available monitoring data for HHCB in water, sediment, and soil were used in the assessment. 

A robust dataset of over 6800 sampling measurements were available from the USGS NWIS for 

wastewater effluent, surface water, and bottom material along with over 500 sampling 

measurements from the scientific literature. The data from both published studies and the 

USGS NWIS database were used because the mean values for surface water and sediment were 

within the same order of magnitude. The concentrations reported in the literature and in the

USGS NWIS data are assumed to represent reasonable estimates, but the data may not reflect 

the actual distribution of HHCB in the environment. Calculation of the RQ based on the 

maximum value, the 95%ile, and the highest mean value was performed in an effort to capture

the full range of concentrations in the environment. Unless actual environmental 

concentrations of HHCB in surface water or sediment are greater than one or two orders of 

magnitude higher than expected, EPA assumes that additional measurements or estimates will 

not affect the risk calculation. 
DRAFT DO NOT CITE OR QUOTE

Page 42

Limited soil monitoring data for HHCB prevented a certain calculation of risk for terrestrial 

organisms associated with land application of biosolids. Two studies with 6 total samples 

reported HHCB concentrations in soil following land application of either biosolids or 

wastewater effluent. The locations are limited to two sites and may not represent the US given 

differences in sludge application practices, WWTP contaminant removal, or seasonal factors, 

for example. Measured concentrations of HHCB in non-amended soil were not available for 

comparison. 

Ecological toxicity studies of HHCB are available in a variety of aquatic, sediment, and terrestrial 

species of different phyla and classes. The lowest toxicity values of acceptable study quality 

were selected for derivation of the COC and standard uncertainty factors were applied to 

account for between-species uncertainty. Among the species tested, the acute and chronic 

aquatic toxicity values varied by about a factor of 10, suggesting that selection of marine or 

freshwater species as the endpoint of concern would not have a significant effect on the risk 

assessment. Sediment toxicity values varied by a factor of about 100 between species and the 

selection of the lowest chronic toxicity value is considered protective of the most sensitive 

species. In soil toxicity tests, springtail and earthworms showed similar sensitivity to HHCB. 

This risk characterization developed conservative estimates of risk for the aquatic environment 

and benthic sediments. The risk estimate for terrestrial environments was based on very limited

monitoring and toxicity data. 

CONCLUSIONS OF ENVIRONMENTAL ASSESSMENT

This assessment evaluated the environmental risk to the aquatic and terrestrial environments

from the use of HHCB as a fragrance ingredient in consumer and commercial products. 

Scenarios of interest included release of HHCB in wastewater to surface water or application of 

biosolids to land. The endpoints of interest were acute and chronic toxicity to aquatic species, 

chronic toxicity to sediment organisms, and chronic toxicity to terrestrial invertebrates and 

plants. The risk assessment considered only direct exposure to aquatic species, sedimentdwelling species, terrestrial species, and plants via contact with contaminated water, sediment, 

or soil. 

The acute and chronic risk for aquatic species did not exceed the COCs at maximum measured 

surface water concentrations. The chronic COC was exceeded for sediment-dwelling species at 

the upper range of measured environmental concentrations for less than 1.5% of the more than 

600 sediment measurements from the USGS. The data for chronic risk for terrestrial 

invertebrates and plants did not exceed the COC at maximum measured soil concentrations; 

this conclusion was based on limited data.
DRAFT DO NOT CITE OR QUOTE

Page 43

References

An, J., Zhou, Q., Sun, Y., and Xu, Z. (2009). Ecotoxicological effects of typical personal care products on 

seed germination and seedling development of wheat (Triticum aestivum L.). Chemosphere 76(10), 

1428-1434. 

Anderson, P., Denslow, N., Drewes, J. E., Olivieri, A., Schlenk, D., Scott, G. I., and Snyder, S. (2012). 

Monitoring strategies for chemicals of emerging concern (CECs) in California's aquatic ecosystems -

recommendations of a science advisory panel. Technical report 692. Southern California Coastal Water 

Research Project, Costa Mesa, CA. 

Andresen, J. A., Muir, D., Ueno, D., Darling, C., Theobald, N., and Bester, K. (2007). Emerging pollutants 

in the North Sea in comparison to Lake Ontario, Canada, data. Environ. Toxicol. Chem. 26(6), 1081-1089. 

Api, A. M., and Ford, R. A. (1999). Evaluation of the oral subchronic toxicity of HHCB (1,3,4,6,7,8-

hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-gamma-2-benzopyran) in the rat. Toxicol. Lett. 111(1-2), 

143-149. (as cited in EC, 2008).

Api, A. M., and San, R. H. (1999). Genotoxicity tests with 6-acetyl-1,1,2,4,4,7-hexamethyltetraline and 

1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-gamma-2-benzopyran. Mutat. Res. 446(1), 67-

81. (as cited in EC, 2008).

Artola-Garicano, E., Sinnige, T. L., van Holsteijn, I., Vaes, W. H., and Hermens, J. L. (2003). 

Bioconcentration and acute toxicity of polycyclic musks in two benthic organisms (Chironomus riparius 

and Lumbriculus variegatus). Environ. Toxicol. Chem. 22(5), 1086-1092. 

Aschmann, S. M., Arey, J., Atkinson, R., and Simonich, S. L. (2001). Atmospheric lifetimes and fates of 

selected fragrance materials and volatile model compounds. Environ. Sci. Technol. 35(18), 3595-3600. 

Ash, Michael, and Irene (2009). Specialty chemicals source book, 4th ed. Vol 2., 1026.

AWWA (2012). Water use statistics. Drinktap.org, Water information. American Water Works 

Association. http://www.drinktap.org/consumerdnn/Default.aspx?tabid=85. Accessed September 24, 

2012. 

Balk, F., Blok, J., and Salvito, D. (2001). Environmental risks of musk fragrance ingredients. In American 

Chemical Society symposium series 791, pharmaceutical and personal care products in the environment: 

Scientific and regulatory issues. American Chemical Society, Washington, DC (as cited in OSPAR, 2004).

Balk, F., and Ford, R. A. (1999a). Environmental risk assessment for the polycyclic musks AHTN and HHCB 

in the EU. I. Fate and exposure assessment. Toxicol. Lett. 111(1-2), 57-79. 

Balk, F., and Ford, R. A. (1999b). Environmental risk assessment for the polycyclic musks, AHTN and 

HHCB. II. Effect assessment and risk characterisation. Toxicol. Lett. 111(1-2), 81-94. 
DRAFT DO NOT CITE OR QUOTE

Page 44

Balk, F., Salvito, D., and Blok, H. (2004). Recent studies conducted by the Research Institute for 

Fragrance Materials (RIFM) in support of the environmental risk assessment process. In The Handbook 

of Environmental Chemistry Vol. 3, Part X. pp. 311-331. Springer-Verlag, Berlin.

Barber, L. B., Brown, G. K., Nettesheim, T. G., Murphy, E. W., Bartell, S. E., and Schoenfuss, H. L. (2011). 

Effects of biologically-active chemical mixtures on fish in a wastewater-impacted urban stream. Sci. Total 

Environ. 409(22), 4720-4728. 

Barnthouse, L. W., DeAngelis, D. L., Gardner, R. H., O'Neill, R. V., Suter, G. W., and Vaughan, D. S. (1982). 

Methodology for environmental risk analysis. ORNL/TM-8167. Oak Ridge National Laboratory, 

Environmental Sciences Division, Publication No. 2023. Prepared for U.S. Environmental Protection 

Agency, Office of Research and Development, Washington, DC, Oak Ridge, TN. 

Basketter, D. (1996). Galaxolide: Skin sensitization study in guinea pigs. Unilever Research, Bedford, U.K. 

Document reference: D95/051. Study SSM84.080 (as cited in EC, 2008).

Bickers, D. R., Calow, P., Greim, H. A., Hanifin, J. M., Rogers, A. E., Saurat, J., Sipes, I. G., Smith, R. L., and 

Tagami, H. (2003). The safety of fragrance materials. Regul. Toxicol. Pharmacol. 37, 218-273. 

Bitsch, N., Dudas, C., Korner, W., Failing, K., Biselli, S., Rimkus, G., and Brunn, H. (2002). Estrogenic 

activity of musk fragrances detected by the E-screen assay using human mcf-7 cells. Arch. Environ. 

Contam. Toxicol. 43(3), 257-264. (as cited in EC, 2008).

Bjornestad, E. (2007). Acartia tonsa larval development test with "HHCB". Project No. 54464, GLP Study 

No. 91328/700. DHI Denmark. Report to International Flavors and Fragrances. Hilversum, NL (as cited in 

EC, 2008).

Boethling, R., Fenner, K., Howard, P., Klecka, G., Madsen, T., Snape, J. R., and Whelan, M. J. (2009). 

Environmental persistence of organic pollutants: guidance for development and review of POP risk 

profiles. Integr. Environ. Assess. Manag. 5(4), 539-556.

Breitholtz, M., Wollenberger, L., and Dinan, L. (2003). Effects of four synthetic musks on the life cycle of 

the harpacticoid copepod Nitocra spinipes. Aquat. Toxicol. 63(2), 103-118. 

Buerge, I. J., Buser, H. R., Muller, M. D., and Poiger, T. (2003). Behavior of the polycyclic musks HHCB and 

AHTN in lakes, two potential anthropogenic markers for domestic wastewater in surface waters. 

Environ. Sci. Technol. 37(24), 5636-5644. 

Buyuksonmez, F., and Sekeroglu, S. (2005). Presence of pharmaceuticals and personal care products 

(PPCPs) in biosolids and their degradation during composting. J. Res. Sci. Technol. 2(1), 31-40. 

Cadby, P. A., Troy, W. R., and Vey, M. G. (2002). Consumer exposure to fragrance ingredients: providing 

estimates for safety evaluation. Regul. Toxicol. Pharmacol. 36(3), 246-252. 

Carlsson, G., and Norrgren, L. (2004). Synthetic musk toxicity to early life stages of Zebrafish (Danio 

rerio). Arch. Environ. Contam. Toxicol. 46(1), 102-105. 
DRAFT DO NOT CITE OR QUOTE

Page 45

Chase, D. A., Karnjanapiboonwong, A., Fang, Y., Cobb, G. P., Morse, A. N., and Anderson, T. A. (2012). 

Occurrence of synthetic musk fragrances in effluent and non-effluent impacted environments. Sci. Total 

Environ. 416, 253-260. 

Chen, C., Xue, S., Zhou, Q., and Xie, X. (2011a). Multilevel ecotoxicity assessment of polycyclic musk in 

the earthworm Eisenia fetida using traditional and molecular endpoints. Ecotoxicology 20(8), 1949-1958. 

Chen, C., Zhou, Q., Bao, Y., Li, Y., and Wang, P. (2010). Ecotoxicological effects of polycyclic musks and 

cadmium on seed germination and seedling growth of wheat (Triticum aestivum). Journal of 

Environmental Sciences 22(12), 1966-1973. 

Chen, C., Zhou, Q., Liu, S., and Xiu, Z. (2011b). Acute toxicity, biochemical and gene expression responses 

of the earthworm Eisenia fetida exposed to polycyclic musks. Chemosphere 83(8), 1147-1154. 

Christian, M. S., Hoberman, A. M., and Parker, R. M. (1997). Oral (gavage) developmental toxicity study 

of 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethyl cyclopenta-gamma-2-benzopyran (HHCB) in rats. Argus 

Research Laboratories Inc. Study 1318-001. Report to the Research Institute for Fragrance Materials, Inc. 

(as cited in EC, 2008).

Christian, M. S., Parker, R. M., Hoberman, A. M., Diener, R. M., and Api, A. M. (1999). Developmental 

toxicity studies of four fragrances in rats. Toxicol. Lett. 111(1-2), 169-174. (as cited in EC, 2008).

COLIPA (1987). Survey of cosmetic usage. The European Cosmetic Toiletry and Perfumery Association, 

submitted to the Scientific Committee on Cosmetology (as cited in Cadby et al., 2002).

Correa, R. S., White, R. E., and Weatherley, A. J. (2005). Biosolids effectiveness to yield ryegrass based on 

their nitrogen content. Scientia Agricola 62(3), 274-280. 

Curry, P. T., and Putman, D. L. (1995). In vitro mammalian cytogenetic test with 1,3,4,6,7,8-hexahydro4,6,6,7,8,8-hexamethylcyclopenta-gamma-2-benzopyran (HHCB). Microbiological Associated Inc. 

Laboratory Study number G94AP81.337005. Report to the Research Institute for Fragrance Materials, 

Inc. (as cited in EC, 2008).

Deblonde, T., Cossu-Leguille, C., and Hartemann, P. (2011). Emerging pollutants in wastewater: A review 

of the literature. Int. J. Hyg. Environ. Health 214(6), 442-448. 

Dietrich, D. R., and Chou, Y. (2001). Ecotoxicology of musks. In Pharmaceuticals and personal care 

products in the environment: Scientific and regulatory issues. American Chemical Society Symposium 

Series 791. American Chemical Society, Washington, DC (as cited in EC, 2008).

Dietrich, D. R., and Hitzfield, B. C. (2004). Bioaccumulation and ecotoxicity of synthetic musks in the 

aquatic environment. In The Handbook of Environmental Chemistry Vol. 3, Part X pp. 233-244. SpringerVerlag, Berlin, Germany.

DiFrancesco, A. M., Chiu, P. C., Standley, L. J., Allen, H. E., and Salvito, D. T. (2004). Dissipation of 

fragrance materials in sludge-amended soils. Environ. Sci. Technol. 38(1), 194-201. (as cited in EC, 2008).
DRAFT DO NOT CITE OR QUOTE

Page 46

Dsikowitzky, L., Schwarzbauer, J., and Littke, R. (2002). Distribution of polycyclic musks in water and 

particulate matter of the Lippe River (Germany). Org. Geochem. 33(12), 1747-1758. 

Duedahl-Olesen, L., Cederberg, T., Pedersen, K. H., and Hojgard, A. (2005). Synthetic musk fragrances in 

trout from Danish fish farms and human milk. Chemosphere 61(3), 422-431. 

EC (2008). European Union risk assessment report for 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-

hexamethylcyclopenta-a-2-benzopyran (1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylin-deno[5,6-

C]pyran-HHCB), CAS No. 1222-05-5, EINECS No. 214-916-9, Risk assessment, final approved version.

European Commission. Office for Official Publications of the European Communities, Luxembourg, The 

Netherlands. http://esis.jrc.ec.europa.eu/doc/risk_assessment/REPORT/hhcbreport414.pdf. Accessed 

September 27, 2012. 

ECHA (2009a). Background document for 5-tert-butyl-2,4,6-trinitro-m-xylene (musk xylene). European 

Chemicals Agency. http://echa.europa.eu/documents/10162/f093e074-1fdc-4d00-af61-f644d2293d4f. 

Accessed September 27, 2012. 

ECHA (2009b). Data on manufacture, import, export, uses and releases of musk xylene (CAS No. 81-15-2) 

as well as information on potential alternatives to its use. European Chemicals Agency. 

Egeler, P. (2004). HHCB/Galaxolide: A study on the toxicity to the sediment dweller Hyalella azteca. ECT 

Study number: AF1HA. ECT Oekotoxikologie GmbH report to International Flavors and Fragrances, IFF. 

(as cited in EC, 2008).

Egeler, P., and Gilberg, D. (2004). HHCB/Galaxolide: A study on the toxicity to the sediment dweller 

Chronomus riparius. ECT Study number: AF1ME. ECT Oekotoxikologie GmbH report to International 

Flavors and Fragrances, IFF. (as cited in EC, 2008).

Envirogen (1998). Fate of HHCB in soil microcosms. Envirogen, Inc. Princeton Research Centre, report 

submitted to International Flavors and Fragrances, Lawrenceville, NJ. (as cited in EC, 2008).

ESIS (2012). CAS# 1222-05-5. European Chemical Substances Information System. European 

Commission, Joint Research Centre, Institute for Health and Consumer Protection. 

http://esis.jrc.ec.europa.eu/. Accessed September 17, 2012. 

European Union (2011). Commission regulation (EU) No 143/2011 of 17 February 2011 amending Annex 

XIV to Regulation (EC) No 1907/2006 of the European Parliament and of the Council on the Registration, 

Evaluation, Authorisation and Restriction of Chemicals (REACH) Official Journal of the European Union

18(2), L 44/42 - L 44/46. http://eurlex.europa.eu/LexUriServ/LexUriServ.do?uri=OJ:L:2011:044:0002:0006:en:pdf. Accessed September 27, 

2012. 

Fahlbusch, K., Hammerschmidt, F., Panten, J., Pickenhagen, W., Schatowski, D., Bauer, K., Garbe, D., and 

Surburg, H. (2012). Flavors and fragrances. In Ullmann's encyclopedia of industrial chemistry. Wiley 

Online Library. http://onlinelibrary.wiley.com/doi/10.1002/14356007.a11_141/pdf. Accessed 

September 27, 2012.
DRAFT DO NOT CITE OR QUOTE

Page 47

Federle, T. W., Itrich, N. R., Lee, D. M., and Langworthy, D. (2002). Recent advances in the environmental 

fate of fragrance ingredients. Poster of P&G presented at SETAC 23rd Annual Meeting, Salt Lake City, 

USA (as cited in EC, 2008)

Fehrenbacher, M. C., and Hummel, A. A. (1996). Evaluation of the mass balance model used by the 

Environmental Protection Agency for estimating inhalation exposure to new chemical substances. Am. 

Ind. Hyg. Assoc. J. 57(6), 526-536. 

Folk, C. M., and Dammers, K. S. (1987). Human phototoxicity study of fragrance materials. Hill Top 

Research Inc. Report to the Research Institute for Fragrance Materials, Inc. Unpublished report, 11 

February. (as cited in EC, 2008).

Ford, R. A., and Bottomley, A. (1997). A method for evaluation of the potential toxicity to the neonate 

from exposure to xenobiotics via mother's milk - application to three fragrance materials. The 

Toxicologist 36(1, Part 2), 367. (as cited in EC, 2008).

Ford, R. A., Hawkins, D. R., Schwarzenbach, R., and Api, A. M. (1999). The systemic exposure to the 

polycyclic musks, AHTN and HHCB, under conditions of use as fragrance ingredients: evidence of lack of 

complete absorption from a skin reservoir. Toxicol. Lett. 111(1-2), 133-142. (as cited in EC, 2008).

Fromme, H., Otto, T., and Pilz, K. (2001). Polycyclic musk fragrances in different environmental 

compartments in Berlin (Germany). Water Res. 35(1), 121-128. 

Gabriel, K. L., and Mark, R. (1987). Phototoxicity study in human subjects Biosearch Incorporated, 

Project Number 87-5460H. Report to the Research Institute for Fragrance Materials, Inc. Unpublished 

report, 23 January. (as cited in EC, 2008).

Gatermann, R., Hellou, J., Huhnerfuss, H., Rimkus, G., and Zitko, V. (1999). Polycyclic and nitro musks in 

the environment: a comparison between Canadian and European aquatic biota. Chemosphere 38(14), 

3431-3441. (as cited in Peck et al., 2006).

Gomez, E., Pillon, A., Fenet, H., Rosain, D., Duchesne, M. J., Nicolas, J. C., Balaguer, P., and Casellas, C. 

(2005). Estrogenic activity of cosmetic components in reporter cell lines: parabens, UV screens, and 

musks. J. Toxicol. Environ. Health A 68(4), 239-251. (as cited in EC, 2008).

Gooding, M. P., Newton, T. J., Bartsch, M. R., and Hornbuckle, K. C. (2006). Toxicity of synthetic musks to 

early life stages of the freshwater mussel Lampsilis cardium. Arch. Environ. Contam. Toxicol. 51(4), 549-

558. 

Green, D. M., and Brian, S. I. (2001). In vitro human skin absorption of radiolabeled fragrance material 

HHCB. Report No. RIFM/4/00, 100901. An-ex Analytical Services Ltd., Report to the Research Institute 

for Fragrance Materials, Inc. (as cited in EC, 2008).

Gudi, R., and Ritter, P. (1997). Micronucleus cytogenetic assay of HHCB in mice. Microbiological 

Associates Inc. Report No. G96BV67.122. Report to the Research Institute for Fragrance Materials, Inc. 

(as cited in EC, 2008).
DRAFT DO NOT CITE OR QUOTE

Page 48

Guillaume, R. Q., Ede, M., and Majors, P. A. (1973a). Repeated insult patch test group of No. 118, SC-04-

0845 (07-0290) (Galaxolide). Hill Top Research, Inc. Report No. 73-391-70. Unpublished report to 

International Flavors and Fragrances, Inc. Private communication to the Research Institute for Fragrance 

Materials. (as cited in EC, 2008).

Guillaume, R. Q., Ede, M., and Majors, P. A. (1973b). Repeated insult patch test group of No. 118, SC-04-

0845 (07-0290) (Galaxolide). Hill Top Research, Inc. Report No. 73-390-70. Unpublished report to 

International Flavors and Fragrances, Inc. Private communication to the Research Institute for Fragrance 

Materials. (as cited in EC, 2008).

Harrison, L. B., and Stolman, L. P. (1986). Phototoxicity testing in human subjects. Harrison Research 

Laboratories Inc. Report to the Research Institute for Fragrance Materials, Inc. Unpublished report, 19 

December. (as cited in EC, 2008).

Hawkins, D. R. (1997). 14C-HHCB – Assessment of the fate of a fragrance chemical in a pig after 

intravenous administration. Project No. RIF 47/971951. Huntingdon Life Sciences Ltd. Report to the 

Research Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

Hawkins, D. R., Elsom, L. F., Winwick, T., and Girkin, R. (1997). 14C-HHCB - Absorption, distribution and 

excretion in rats after single intravenous doses. Project No. RIF 43/963877. Huntingdon Life Sciences Ltd. 

Report to the Research Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

Hawkins, D. R., Elsom, L. F., Winwick, T., and Nicoll, S. A. (1996b). 14C-HHCB: Investigation of the 

transfer across the placenta and into milk of rats during and after pregnancy following repeated oral 

administration. RIF 35/952157. Huntingdon Life Sciences Ltd. Report to the Research Institute for 

Fragrance Materials, Inc. Unpublished report, 3, September. (as cited in EC, 2008).

Hawkins, D. R., Kirkpatrick, D., and Girkin, R. (1995). 14C-HHCB - Studies on the dermal absorption in the 

rat. Project No. RIF32/951257. Huntingdon Life Sciences Ltd. Report to the Research Institute for 

Fragrance Materials, Inc. (as cited in EC, 2008).

Hawkins, D. R., Kirkpatrick, D., Girkin, R., and Brooker, M. D. (1996a). 14C-HHCB: Studies on the dermal 

absorption in the rat. Project No. RIF32/951257. Huntingdon Life Sciences Ltd. Report to the Research 

Institute for Fragrance, Inc. (as cited in EC, 2008).

Haynes, G. (1984). Acute dermal irritation study. Study Reference No. 307-338/8403. Toxicol 

Laboratories. Report to the Research Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

Haynes, G. (1985). Acute dermal irritation study. Study Reference No. 70-101/8503. Toxicol 

Laboratories. Report to the Research Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

Haynes, G. (1986). Acute dermal irritation study. Study Reference No. 150-173/8602. Toxicol 

Laboratories. Report to the Research Institute for Fragrance Materials, Inc. (as cited in EC, 2008).
DRAFT DO NOT CITE OR QUOTE

Page 49

Heberer, T., Gramer, S., and Stan, H. J. (1999). Occurrence and distribution of organic contaminants in 

the aquatic system in Berlin. Part III: Determination of synthetic musks in Berlin surface water applying 

solid-phase microextraction (SPME) and gas chromatography-mass spectrometry (GC-MS). Acta 

Hydrochimica et Hydriobiologica 27(3), 150-156. 

Helm, P. A., Howell, E. T., Li, H., Metcalfe, T. L., Chomicki, K. M., and Metcalfe, C. D. (2012). Influence of 

nearshore dynamics on the distribution of organic wastewater-associated chemicals in Lake Ontario 

determined using passive samplers. J. Great Lakes Res. Accessed September 27, 2012. 

HERA (2004a). Human and environmental risk assessment on ingredients of household cleaning 

products: Polycyclic musks AHTN (CAS 1506-02-1) and HHCB (CAS 1222-05-5). Environmental section. 

Version 2.0. Human and Environmental Risk Assessment. http://www.heraproject.com/files/28-E36551E10-F8EF-E807-E4199B9BB0076A9F.pdf. Accessed September 27, 2012. 

HERA (2004b). Human and environmental risk assessment on ingredients of household cleaning 

products: HHCB (1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-y-2-benzopyran and related 

isomers). Version 2.0. Human and Environmental Risk Assessment. 

http://www.heraproject.com/files/29-HH-04-

pcm%20HHCB%20HERA%20Human%20Health%20DISCL%20ed2.pdf. Accessed September 27, 2012. 

Hope, B. K., Pillsbury, L., and Boling, B. (2012). A state-wide survey in Oregon (USA) of trace metals and 

organic chemicals in municipal effluent. Sci. Total Environ. 417-418, 263-272. 

Hopkins, M. N., Canning, M., Wright, J. A., Lambert, A. H., and Trenchard-Morgan, S. (1996). HHCB: 13-

Week oral (dietary) toxicity study of HHCB in the rat with a 4 week treatment period. Addendum 

attached. Toxicol Laboratories. Report to the Research Institute for Fragrance Materials, Inc. (as cited in 

EC, 2008).

Horii, Y., Reiner, J. L., Loganathan, B. G., Senthil Kumar, K., Sajwan, K., and Kannan, K. (2007). Occurrence 

and fate of polycyclic musks in wastewater treatment plants in Kentucky and Georgia, USA. 

Chemosphere 68(11), 2011-2020. 

HSDB (2007). Galaxolide. CASRN 1222-05-5. Hazardous Substance Data Bank. 

http://toxnet.nlm.nih.gov/cgi-bin/sis/search/a?dbs+hsdb:@term+@DOCNO+7514. Accessed September 

17, 2012. 

Hu, Z., Shi, Y., Niu, H., Cai, Y., Jiang, G., and Wu, Y. (2010). Occurrence of synthetic musk fragrances in 

human blood from 11 cities in China. Environ. Toxicol. Chem. 29(9), 1877-1882. 

Hutter, H. P., Wallner, P., Hartl, W., Uhl, M., Lorbeer, G., Gminski, R., Mersch-Sundermann, V., and 

Kundi, M. (2010). Higher blood concentrations of synthetic musks in women above fifty years than in 

younger women. Int. J. Hyg. Environ. Health 213(2), 124-130. 

Hutter, H. P., Wallner, P., Moshammer, H., Hartl, W., Sattelberger, R., Lorbeer, G., and Kundi, M. (2005). 

Blood concentrations of polycyclic musks in healthy young adults. Chemosphere 59(4), 487-492. 

IFF (2012). Synthetics Compendium. International Flavos and Fragrances Inc. 

http://www.iff.com/custom/iff/pdfs/Synthetics_Compendium_A4_Sheets.pdf. 
DRAFT DO NOT CITE OR QUOTE

Page 50

IFRA (2011). EU Regulation follows fragrance industry's voluntary global ban. International Fragrance 

Association. www.ifraorg.org/view_document.aspx?docId=22609. Accessed August 29, 2012. 

IFRA (2012a). Re: Response to USEPA's 2012 TSCA work plan chemical: 1,3,4,6,7,8-hexahydro4,6,6,7,8,8-hexamethylcyclopenta[g]-2-benzopyran (HHCB); CAS 1222-05-5. Letter to the US EPA. 

March 30, 2012. International Fragrance Association North America. 

IFRA (2012b). HHCB. Email from Jane Wishneff, IFRA, to Leslie Cronkhite, US EPA. June 29, 2012. 

International Fragrance Association. 

IFRA (2012c). REACH Exposure scenarios for fragrance substances. International Fragrance Association, 

Brussels, Belgium. 

IFRA (2012d). IFRANA RIFM responses to EPA Questions 6.14. Letter to the US EPA, June 14, 2012.

International Fragrance Association North America. 

Jager, T. (1998). Mechanistic approach for estimating bioconcentration of organic chemicals in 

earthworms (Oligochaeta). Environ. Toxicol. Chem. 17(10), 2080-2090. (as cited in EC, 2008).

Jenkins, W. R. (1991). Abbalide: Assessment of its biodegradability, modified sturm test. Life Science 

Research Report 90/BAK003/1361. Bush Boake Allen, Inc. (as cited in EC, 2008).

Jones, K., Bottomley, A. M., and Gopinath, C. (1996). HHCB: Effects on peri- and post natal development 

including maternal function in the rat (gavage administration). RIF 36/961217. Huntingdon Life Sciences 

Ltd. Report to the Research Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

Kang, C. S., Lee, J., Kim, S., Lee, K., Lee, J. S., Park, P. S., Yun, S. H., Kannan, K., Yoo, Y. W., Ha, J. Y., and 

Lee, S. W. (2010). Polybrominated diphenyl ethers and synthetic musks in umbilical cord serum, 

maternal serum, and breast milk from Seoul, South Korea. Chemosphere 80, 116-122. 

Kannan, K., Reiner, J. L., Yun, S. H., Perrotta, E. E., Tao, L., Johnson-Restrepo, B., and Rodan, B. D. (2005). 

Polycyclic musk compounds in higher trophic level aquatic organisms and humans from the United 

States. Chemosphere 61(5), 693-700. 

Kevekordes, S., Mersch-Sundermann, V., Diez, M., Bolten, C., and Dunkelberg, H. (1998). Genotoxicity of 

polycyclic musk fragrances in the sister-chromatid exchange test. Anticancer Res. 18(1A), 449-452. (as 

cited in EC, 2008).

Kevekordes, S., Mersch-Sundermann, V., Diez, M., and Dunkelberg, H. (1997). In vitro genotoxicity of 

polycyclic musk fragrances in the micronucleus test. Mutat. Res. 395(2-3), 145-150. (as cited in EC, 

2008).

King, J. M. H. (1994). Assessment of the inherent biodegradability of 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-

hexamethylcyclopenta-gamma-2-benzopyran IPM in the sealed vessel test using acclimatised effluent.

Unilever BD/END/03. (as cited in EC, 2008).
DRAFT DO NOT CITE OR QUOTE

Page 51

Kinney, C. A., Furlong, E. T., Kolpin, D. W., Burkhardt, M. R., Zaugg, S. D., Werner, S. L., Bossio, J. P., and 

Benotti, M. J. (2008). Bioaccumulation of pharmaceuticals and other anthropogenic waste indicators in 

earthworms from agricultural soil amended with biosolid or swine manure. Environ. Sci. Technol. 42(6), 

1863-1870. 

Kinney, C. A., Furlong, E. T., Zaugg, S. D., Burkhard, M. R., Werner, S. L., Cahill, J. D., and Jorgensen, G. R. 

(2006). Survey of organic wastewater contaminants in biosolids destined for land application. Environ. 

Sci. Technol. 40(23), 7207-7215. 

Klecka, G., Persoon, C., and Currie, R. (2010). Chemicals of emerging concern in the Great Lakes Basin: an 

analysis of environmental exposures. Rev. Environ. Contam. Toxicol. 207, 1-93. 

Langdon, K. A., Warne, M. S., and Kookana, R. S. (2010). Aquatic hazard assessment for pharmaceuticals, 

personal care products, and endocrine-disrupting compounds from biosolids-amended land. Integr. 

Environ. Assess. Manag. 6(4), 663-676. 

Langworthy, D. E., Itrich, N. R., Simonich, S. L., and Federle, T. W. (2000). Biotransformation of the 

polycyclic musk HHCB in activated sludge and river water. Presented at SETAC, May 2000, Brighton, UK 

(as cited in EC, 2008)

Lee, H. B., Pert, T. E., and Sarafin, K. (2003). Occurrence of polycyclic and nitro musk compounds in 

Canadian sludge and wastewater samples. Water Qual. Res. J. Can. 38(4), 683-702. 

Levenstein, I. (1973a). Acute skin irritation study in rabbits. Leberco Laboratories. Assay No. 34531.

Unpublished Report to International Flavors and Fragrances, Inc. Private communication to the Research 

Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

Levenstein, I. (1973b). Acute eye irritation study in rabbits. Leberco Laboratories. Assay No. 34542.

Unpublished Report to International Flavors and Fragrances, Inc. Private communication to the Research 

Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

Levenstein, I. (1975a). Acute skin irritation study in rabbits. Leberco Laboratories. Assay No. 52697.

Unpublished Report to International Flavors and Fragrances, Inc. Private communication to the Research 

Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

Levenstein, I. (1975b). Acute eye irritation study in rabbits. Leberco Laboratories. Assay No. 52692.

Unpublished Report to International Flavors and Fragrances, Inc. Private communication to the Research 

Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

Liebl, B., Mayer, R., Ommer, S., Sonnichsen, C., and Koletzko, B. (2000). Transition of nitro musks and 

polycyclic musks into human milk. In Short and long term effects of breast feeding on child health. pp. 

289-305. Kluwer Academic/Plenum Publishers.

Lignell, S., Darnerud, P. O., Aune, M., Cnattingius, S., Hajslova, J., Setkova, L., and Glynn, A. (2008). 

Temporal trends of synthetic musk compounds in mother's milk and associations with personal use of 

perfumed products. Environ. Sci. Technol. 42(17), 6743-6748. 
DRAFT DO NOT CITE OR QUOTE

Page 52

Lindstrum, P., Ison, A., and Carabello, F. B. (1978a). Human phototoxicity studies using HHCB:(June 9), 

25%, 50%, and 100% of commercially diluted HHCB. Hill Top Research, Inc. Report No. 78-296-72.

Unpublished Report to International Flavors and Fragrances, Inc. Private communication to the Research 

Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

Lindstrum, P., Ison, A., and Carabello, F. B. (1978b). Human phototoxicity studies using HHCB: (July 19), 

100% commercially diluted Hill Top Research, Inc Report No. 78-299-72. Unpublished Report to 

International Flavors and Fragrances, Inc. Private communication to the Research Institute for Fragrance 

Materials, Inc. (as cited in EC, 2008).

Lindstrum, P., Ison, A., and Carabello, F. B. (1978c). Human phototoxicity studies using HHCB: (October 

3), 100% HHCB and 65% HHCB in two diluents. Hill Top Research, Inc Report No. 78-844-72. Unpublished 

Report to International Flavors and Fragrances, Inc. Private communication to the Research Institute for 

Fragrance Materials, Inc. (as cited in EC, 2008).

Litz, N. T., Muller, J., and Bohmer, W. (2007). Occurrence of polycyclic musks in sewage sludge and their 

behavior in soils and plants. J. Soils Sediments 7(1), 36-44. 

Luckenbach, T., Corsi, I., and Epel, D. (2004). Fatal attraction: synthetic musk fragrances compromise 

multixenobiotic defense systems in mussels. Mar. Environ. Res. 58(2-5), 215-219. 

Luckenbach, T., and Epel, D. (2005). Nitromusk and polycyclic musk compounds as long-term inhibitors 

of cellular xenobiotic defense systems mediated by multidrug transporters. Environ. Health. Perspect.

113(1), 17-24. 

Mersch-Sundermann, V., Kevekordes, S., and Jenter, C. (1998a). Lack of mutagenicity of polycyclic musk 

fragrances in Salmonella typhimurium. Toxicol. in Vitro 12(4), 389-393. (as cited in EC, 2008).

Mersch-Sundermann, V., Kevekordes, S., and Jenter, C. (1998b). Testing of SOS induction of artificial 

polycyclic musk fragrances in E. coli PQ37 (SOS chromotest). Toxicol. Lett. 95(3), 147-154. (as cited in EC, 

2008).

Minner, R. J., and Foster, G. V. (1977). Comparative acute toxicity studies in the female rat with five 

synthetic musk chemicals. Avon Products, New York. (as cited in EC, 2008).

Moon, H. B., An, Y. R., Choi, S. G., Choi, M., and Choi, H. G. (2012). Accumulation of PAHs and synthetic 

musk compound in minke whales (Balanoptera acutorostrata) and long-beaked common dolphins 

(Delphinus capensis) from Korean coastal waters. Environ. Toxicol. Chem. 31(3), 477-485. 

Moreno, O. M. (1975). Galaxolide 50: Acute oral toxicity in rats; dermal toxicity in rabbits. Project No. 

MB 75-770. MB Research. Report to the Research Institute for Fragrance Materials, Inc. (as cited in EC, 

2008).

Mori, T., Morita, F., Inokuchi, A., Takao, Y., Kohra, S., Tominaga, N., Takemasa, T., and Arizono, K. (2006). 

Ecotoxicological effect of polycyclic musks on Caenorhabditis elegans. J. Health Sci. 52(3), 276-282. 
DRAFT DO NOT CITE OR QUOTE

Page 53

NEBRA (2007). A national biosolids regulation, quality, end & use disposal survey. Final report. North 

East Biosolids and Residuals Association. http://www.nebiosolids.org/uploads/pdf/NtlBiosolidsReport20July07.pdf. Accessed September 17, 2012. 

Oblinger Childress, C. J., Foreman, W. T., Connor, B. F., and Maloney, T. J. (1999). New reporting 

procedures based on long-term method detection levels and some considerations for interpretations of 

water-quality data provided by the US Geological Survey National Water Quality Laboratory. US 

Geological Survey. Open-File report 99-193, 1-19. 

OECD (2009). HHCB, SIAM 28. SIDS Initial assessment profile (CAS No. 1222-05-5). Organisation for 

Economic Co-operation and Development, Screening Information Data Set (SIDS) for High Production 

Volume (HPV) chemicals, Germany. http://webnet.oecd.org/hpv/UI/handler.axd?id=b36caa92-a554-

4853-b98d-744e72cb9d56. Accessed September 18, 2012. 

OECD (2010). Emission scenario document on the blending of fragrance oils into commercial and 

consumer products. OECD Environment, Health and Safety Publications Series on Emission Scenario 

Documents Number 26. Organisation for Economic Co-operation and Development, Paris, France. 

http://www.oecd.org/env/chemicalsafetyandbiosafety/assessmentofchemicals/46021591.pdf. Accessed 

September 27, 2012. 

Osemwengie, L. I., and Gerstenberger, S. L. (2004). Levels of synthetic musk compounds in municipal 

wastewater for potential estimation of biota exposure in receiving waters. J. Environ. Monit. 6(6), 533-

539. 

Osemwengie, L. I., and Steinberg, S. (2001). On-site solid-phase extraction and laboratory analysis of 

ultra-trace synthetic musks in municipal sewage effluent using gas chromatography-mass spectrometry 

in the full-scan mode. J. Chromatogr. A 932, 107-118. 

OSPAR (2004). Musk xylene and other musks, no. 200. In Hazardous substance series. Oslo-Paris 

Commission. OSPAR Secretariat, London, UK. 

http://www.ospar.org/documents/dbase/publications/p00200_BD%20on%20musk%20xylene.pdf. 

Accessed September 27, 2012.

Parish, W. E. (1988). Photosensitization study on Galaxolide (1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-

hexamethylcyclopenta-gamma-2-benzopyran). Private communication to the Research Institute for 

Fragrance Materials, Inc. (as cited in EC, 2008).

Peck, A. M., and Hornbuckle, K. C. (2004). Synthetic musk fragrances in Lake Michigan. Environ. Sci. 

Technol. 38(2), 367-372. 

Peck, A. M., Linebaugh, E. K., and Hornbuckle, K. C. (2006). Synthetic musk fragrances in Lake Erie and 

Lake Ontario sediment cores. Environ. Sci. Technol. 40(18), 5629-5635. 2757450.

Pedersen, S., Selck, H., Salvito, D., and Forbes, V. (2009). Effects of the polycyclic musk HHCB on 

individual- and population-level endpoints in Potamopyrgus antipodarum. Ecotoxicol. Environ. Saf. 72(4), 

1190-1199. 
DRAFT DO NOT CITE OR QUOTE

Page 54

Peters, R. J. B. (2005). Man-made chemicals in maternal and cord blood. TNO-report, B&O-A R 

2005/129. TNO Built Environment and Geosciences, submitted to Greenpeace International and WWFUK. 

Potera, C. (2007). The sweet scent on baby's breath? Environ. Health. Perspect. 115(10), A491. 2022646.

Ramirez, A. J., Brain, R. A., Usenko, S., Mottaleb, M. A., O'Donnell, J. G., Stahl, L. L., Wathen, J. B., Snyder, 

B. D., Pitt, J. L., Perez-Hurtado, P., Dobbins, L. L., Brooks, B. W., and Chambliss, C. K. (2009). Occurrence 

of pharmaceuticals and personal care products in fish: results of a national pilot study in the United 

States. Environ. Toxicol. Chem. 28(12), 2587-2597. 

Ramskov, T., Selck, H., Salvito, D., and Forbes, V. E. (2009). Individual- and population-level effects of the 

synthetic musk, HHCB, on the deposit-feeding polychaete, Capitella sp. I. Environ. Toxicol. Chem. 28(12), 

2695-2705. 

Reiner, J. L., Berset, J. D., and Kannan, K. (2007a). Mass flow of polycyclic musks in two wastewater 

treatment plants. Arch. Environ. Contam. Toxicol. 52(4), 451-457. 

Reiner, J. L., and Kannan, K. (2006). A survey of polycyclic musks in selected household commodities 

from the United States. Chemosphere 62(6), 867-873. 

Reiner, J. L., and Kannan, K. (2011). Polycyclic musks in water, sediment, and fishes from the upper 

Hudson River, New York, USA. Water, Air, & Soil Pollution 214(1-4), 335-342. 

Reiner, J. L., Wong, C. M., Arcaro, K. F., and Kannan, K. (2007b). Synthetic musk fragrances in human milk 

from the United States. Environ. Sci. Technol. 41(11), 3815-3820. 

RIFM (1996). Cosmetics: Submission on the safety data on ingredients: 1,3,4,6,7,8-hexahydro4,6,6,7,8,8-hexamethylcyclopenta-gamma-2-benzopyran (HHCB) and 6-acetyl-1,1,2,4,4,7-

hexamethyltetraline (AHTN). Submission to DG XXIV, European Commission (as cited in Cadby et al., 

2002).

Rimkus, G. G. (1999). Polycyclic musk fragrances in the aquatic environment. Toxicol. Lett. 111(1-2), 37-

56. (as cited in Luckenbach and Epel, 2005).

Roosens, L., Covaci, A., and Neels, H. (2007). Concentrations of synthetic musk compounds in personal 

care and sanitation products and human exposure profiles through dermal application. Chemosphere

69(10), 1540-1547. 

Rubenkoenig, H. L., and Ede, M. (1964). Repeated insult patch test in humans. Hill Top Research Institute 

Inc. SC0998-G. Unpublished report to IFF Incorporated, 8 June. Private communication to the Research 

Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

San, R. C., Klug, M. L., and Wagner, V. O. (1994). Salmonella/Escherichia coli plate incorporation 

mutagenicity assay with a confirmatory assay with 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-

hexamethylcyclopenta-gamma-2-benzopyran (HHCB). Microbiological Associates, Inc. Laboratory Study 

Number G94AP81.501088. Report to the Research Institute for Fragrance Materials, Inc. (as cited in EC, 

2008).
DRAFT DO NOT CITE OR QUOTE

Page 55

San, R. H. C., and Sly, J. E. (1994). Unscheduled DNA synthesis assay in rat primary hepatocytes with 

1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-gamma-2-benzopyran (HHCB). Microbiological 

Associates, Inc. Laboratory Study Number G94AP81.380030. Report to the Research Institute for 

Fragrance Materials, Inc. (as cited in EC, 2008).

Sapozhnikova, Y., Liebert, D., Wirth, E., and Fulton, M. (2010). Polycyclic musk fragrances in sediments 

and shrimp tissues. Polycyclic Aromat. Compd. 30(5), 298-308. 

Sauer, C. (1980). Primary ocular irritation study in rabbits. Cosmopolitan Safety Evaluation, Inc. Study 

No. 0264. Unpublished Report to International Flavors and Fragrances, Inc. Private communication to 

the Research Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

Schaefer, E. C. (2005). 14C-HHCB: Dieaway of a semi-volatile organic compound in river water. Project 

No. 558E-109A. Wildlife International, Ltd. Report submitted to the Research Institute for Fragrance 

Materials, Inc. 

Schiavone, A., Kannan, K., Horii, Y., Focardi, S., and Corsolini, S. (2010). Polybrominated diphenyl ethers, 

polychlorinated naphthalenes and polycyclic musks in human fat from Italy: comparison to 

polychlorinated biphenyls and organochlorine pesticides. Environ. Pollut. 158, 599-606. 

Schnell, S., Martin-Skilton, R., Fernandes, D., and Porte, C. (2009). The interference of nitro- and 

polycyclic musks with endogenous and xenobiotic metabolizing enzymes in carp: an in vitro study. 

Environ. Sci. Technol. 43(24), 9458-9464. 

Schreiber, W. L. (2004). Perfumes. In Kirk-Othmer encyclopedia of chemical technology, 5th ed., Vol. 18.

John Wiley & Sons, Inc.

Schreurs, R. H., Quaedackers, M. E., Seinen, W., and van der, B. B. (2002). Transcriptional activation of 

estrogen receptor ERalpha and ERbeta by polycyclic musks is cell type dependent. Toxicol. Appl. 

Pharmacol. 183(1), 1-9. (as cited in EC, 2008).

Schreurs, R. H., Sonneveld, E., Jansen, J. H., Seinen, W., and van der, B. B. (2005a). Interaction of 

polycyclic musks and UV filters with the estrogen receptor (ER), androgen receptor (AR), and 

progesterone receptor (PR) in reporter gene bioassays. Toxicol. Sci. 83(2), 264-272. (as cited in EC, 2008).

Schreurs, R. H. M. M., Legler, J., Artola-Garicano, E., Sinnige, T. L., Lanser, P. H., Seinen, W., and van der, 

B. B. (2004). In vitro and in vivo antiestrogenic effects of polycyclic musks in zebrafish. Environ. Sci. 

Technol. 38(4), 997-1002. (as cited in EC, 2008).

Schreurs, R. H. M. M., Sonneveld, E., Van der Saag, P. T., and Van der Burg, B. (2005b). Examination of 

the in vitro (anti)estrogenic, (anti)androgenic and (anti)dioxin-like activities of tetralin, indane and 

isochroman derivatives using receptor-specific bioassays. Toxicol. Lett. 156(2), 261-275. (as cited in EC, 

2008).

Seinen, W., Lemmen, J. G., Pieters, R. H., Verbruggen, E. M., and van der, B. B. (1999). AHTN and HHCB 

show weak estrogenic--but no uterotrophic activity. Toxicol. Lett. 111(1-2), 161-168. (as cited in EC, 

2008).
DRAFT DO NOT CITE OR QUOTE

Page 56

Shanahan, R. W., and Alworth, K. (1987). Phototoxicity testing in human subjects. Essex Testing Clinic 

Inc. Report to the Research Institute for Fragrance Materials, Inc. Unpublished report. (as cited in EC, 

2008).

Simmons, D. B., Marlatt, V. L., Trudeau, V. L., Sherry, J. P., and Metcalfe, C. D. (2010). Interaction of 

Galaxolide(R) with the human and trout estrogen receptor-alpha. Sci. Total Environ. 408(24), 6158-6164. 

Simonich, S. L., Federle, T. W., Eckhoff, W. S., Rottiers, A., Webb, S., Sabaliunas, D., and de Wolf, W. 

(2002). Removal of fragrance materials during U.S. and European wastewater treatment. Environ. Sci. 

Technol. 36(13), 2839-2847. 

Smyth, S. A., Lishman, L. A., McBean, E. A., Kleywegt, S., Yang, J. J., Svoboda, M. L., Lee, H. B., and Seto, 

P. (2007). Fate of polycyclic and nitro musks during aerobic and anaerobic sludge digestion. In Moving 

Forward: Wastewater biosolids sustainability: Technical, managerial, and public synergy, June 24-27, 

2007. pp. 223-229. ILA International Water Association, Moncton, New Brunswick, Canada. 

http://www.bvsde.paho.org/bvsaar/cdlodos/pdf/fateofpolycyclic223.pdf. Accessed September 27, 

2012.

Smyth, S. A., Lishman, L. A., McBean, E. A., Kleywegt, S., Yang, J. J., Svoboda, M. L., Lee, H. B., and Seto, 

P. (2008). Seasonal occurrence and removal of polycyclic and nitro musks from wastewater treatment 

plants in Ontario, Canada. Journal of Environmental Engineering and Science 7(4), 299-317. 

Sommer, C. (2004). The role of musk and musk compounds in the fragrance industry. In Handbook of 

environmental chemistry, Vol. 3, Part X. pp. 1-16. Springer-Verlag, Berlin, Germany.

Somogyi, L. P., and Kishi, A. (2001). Aroma chemicals and the flavor and fragrance industry. SRI 

International (as cited in Peck et al., 2006)

Stevens, J. L., Northcott, G. L., Stern, G. A., Tomy, G. T., and Jones, K. C. (2003). PAHs, PCBs, PCNs, 

organochlorine pesticides, synthetic musks, and polychlorinated n-alkanes in U.K. sewage sludge: survey 

results and implications. Environ. Sci. Technol. 37(3), 462-467. 

SWECO (2008). Screening of musk substances. SWECO Environment screening report 2008:2. Swedish 

Environmental Protection Agency, Malmo, Sweden. 

http://www.naturvardsverket.se/upload/02_tillstandet_i_miljon/Miljoovervakning/rapporter/miljogift/

screening_mysk.pdf. Accessed September 27, 2012. 

US EPA (1988). Releases during cleaning of equipment. Prepared by PEI Associates, Inc. US 

Environmental Protection Agency, Office of Pesticides and Toxic Substances, Washington, DC. 

US EPA (1991). Preparation of engineering assessments. Volume I: CEB Engineering manual. US 

Environmental Protection Agency, Office of Toxic Substances. Chemical Engineering Branch, Economics 

and Technology Division., Washington, DC. 

US EPA (1992). Standard assumptions for PMN assessments. Memorandum from Chemical Engineering 

Branch Quality Panel to Chemical Engineering Branch Staff and Management. US Environmental 

Protection Agency, Washington, DC. 
DRAFT DO NOT CITE OR QUOTE

Page 57

US EPA (1994). Guidelines for statistical analysis of occupational exposure data. Final. US Environmental 

Protection Agency, Office of Pollution Prevention and Toxics 

http://www.epa.gov/oppt/exposure/pubs/stat_guide_occ.pdf. Accessed September 27, 2012. 

USEPA (1998). Guidelines for Ecological Risk Assessment. Fed. Regist. 63(93):26846-26924. 

http://www.epa.gov/raf/publications/pdfs/ECOTXTBX.PDF. Accessed April 1, 2012.

US EPA (1999a). Category for persistent, bioaccumulative, and toxic new chemical substances. US 

Environmental Protection Agency. Fed. Regist. 64(213), 60194-60204. 

http://www.epa.gov/fedrgstr/EPA-TOX/1999/November/Day-04/t28888.htm. Accessed September 27, 

2012. 

US EPA (1999b). Persistent bioaccumulative toxic (PBT) chemicals; lowering of reporting thresholds for 

certain PBT chemicals; addition of certain PBT chemicals; community right-to-know toxic chemical 

reporting. Fed. Regist. 64(209), 58666-58753. 

US EPA (1999c). Determining the adequacy of existing data. US Environmental Protection Agency, Office 

of Pollution Prevention and Toxics. http://www.epa.gov/hpv/pubs/general/datadfin.htm. Accessed 

September 27, 2012. 

US EPA (2000). Interim guidance for using ready and inherent biodegradability tests to derive input data 

for multimedia models and wastewater treatment plants (WWT) models (9/1/2000). Exposure 

assessment tools and models. http://www.epa.gov/oppt/exposure/pubs/halflife.htm. Accessed 

September 17, 2012. 

US EPA (2003). Robust test summaries for 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethylcyclopentagamma-2-benzopyran (HHCB) CAS# 1222-05-5. US Environmental Protection Agency. Submitted to the 

EPA under the HPV Challenge Program by International Flavors & Fragrances. Revised May 2004. 

http://www.epa.gov/HPV/pubs/summaries/cyclopen/c14820rs.pdf. Accessed September 18, 2012. 

US EPA (2009a). Methodology for risk-based prioritization under ChAMP. US Environmental Protection 

Agency, Office of Pollution Prevention and Toxics. 

US EPA (2009b). Interpretive assistance document for assessment of discrete organic chemicals. 

Sustainable futures summary assessment. US Environmental Protection Agency, Office of Pollution 

Prevention and Toxics. 

US EPA (2011). Estimation program interface (EPI) suite version 4.10. Exposure assessment tools and 

models. US Environmental Protection Agency, Office of Chemical Safety and Pollution Prevention. 

http://www.epa.gov/opptintr/exposure/pubs/episuite.htm. Accessed September 17, 2012. 

US EPA (2012a). TSCA Work plan chemicals: methods document. US Environmental Protection Agency, 

Office of Pollution Prevention and Toxics. 

US EPA (2012b). Work plan chemicals for assessment during 2013 and 2014. US Environmental 

Protection Agency, Office of Pollution Prevention and Toxics. 

http://www.epa.gov/oppt/existingchemicals/pubs/workplanlist.html. Accessed September 17, 2012. 
DRAFT DO NOT CITE OR QUOTE

Page 58

US EPA (2012c). Results from inert ingredient test orders issued under EPA's endocrine disruptor 

screening program: New data compensation claims; potential disapproval of inert uses pending public 

comment. Fed. Regist. 77(50), 15101-15104. http://www.gpo.gov/fdsys/pkg/FR-2012-03-14/pdf/2012-

6164.pdf. Accessed September 17, 2012. 

US EPA (2012d). CAS: 1222-05-5 and 81-15-2. Non confidential 2006 IUR data by chemical, including 

manufacturing, processing and use information. US Environmental Protection Agency, Office of Pollution 

Prevention and Toxics. http://cfpub.epa.gov/iursearch/index.cfm?s=chem. Accessed July 19, 2012. 

US EPA (2012e). Search results by CAS No: 1222-05-5 and 81-15-2. Non-confidential IUR production 

volume data 1986-2002. US Environmental Protection Agency, Office of Pollution Prevention and Toxics. 

http://www.epa.gov/oppt/iur/tools/data/2002-vol.html. Accessed September 27, 2012. 

USCB (2004). Toilet preparation manufacturing: 2002. 2002 Economic census manufacturing industry 

series. US Census Bureau. US Department of Commerce, Economics and Statistics Administration. 

http://www.census.gov/prod/ec02/ec0231i325620t.pdf. Accessed September 18, 2012. 

USCB (2007a). NAICS 325620 Toilet preparation manufacturing. Geographic distribution - Toilet 

preparation manufacturing: 2007. Industry statistics sampler. 2007 Census: Geographic distribution. US 

Census Bureau. US Department of Commerce. 

http://www.census.gov/econ/industry/geo/g325620.htm. Accessed September 18, 2012. 

USCB (2007b). NAICS 325611 Soap and other detergent manufacturing. Geographic distribution - Soap 

and other detergent manufacturing: 2007. Industry statistics sampler. 2007 Census: Geographic 

distribution. US Census Bureau. US Department of Commerce. 

http://www.census.gov/econ/industry/geo/g325611.htm. Accessed September 18, 2012. 

USCB (2007c). NAICS 325612 Polish and other sanitation good manufacturing. Geographic distribution -

polish and other sanitation good manufacturing: 2007. Industry statistics sampler. 2007 Census: 

Geographic distribution. US Census Bureau. US Department of Commerce. 

http://www.census.gov/econ/industry/geo/g325612.htm. Accessed September 18, 2012. 

USGS (2008). Occurrence of organic wastewater compounds in the Tinkers Creek Watershed and two 

other tributaries to the Cuyahoga River, Northeast Ohio. USGS scientific investigations report 2008-

5173. 60 p. . US Geological Survey, US Department of the Interior, Reston, VA. 

http://pubs.usgs.gov/sir/2008/5173/. Accessed September 27, 2012. 

USGS (2011). Wastewater indicator compounds in wastewater effluent, surface water, and bed 

sediment in the St. Croix National Scenic Riverway and implications for water resources and aquatic 

biota, Minnesota and Wisconsin, 2007-08. Scientific investigations report 2011-5208. 40 p. . US 

Geological Survey, US Department of the Interior, Denver, CO. http://pubs.usgs.gov/sir/2011/5208/. 

Accessed September 27, 2012. 

USGS (2012). [1,3,4,6,7,8-Hexahydro-4,6,6,7,8,8-hexamethylcyclopenta-γ-2-benzopyran]. National 

Water Information System: Web interface. U.S. Geological Survey. http://waterdata.usgs.gov/nwis. 

Accessed June 15, 2012. 
DRAFT DO NOT CITE OR QUOTE

Page 59

Wang, H., Zhang, J., Gao, F., Yang, Y., Duan, H., Wu, Y., Berset, J., and Shao, B. (2011). Simultaneous 

analysis of synthetic musks and triclosan in human breast milk by gas chromatography tandem mass 

spectrometry. Journal of Chromatography B 879, 1861-1869. 

WERF (2007). Fate of pharmaceuticals and personal care products through municipal wastewater 

treatment processes. 03-CTS-22UR. Water Environment Research Foundation, London. 

Wollenberger, L., Breitholtz, M., Ole, K. K., and Bengtsson, B. E. (2003). Inhibition of larval development 

of the marine copepod Acartia tonsa by four synthetic musk substances. Sci. Total Environ. 305(1-3), 53-

64. 

Wolven, A., and Levenstein, I. (1963). Acute eye irritation study in rabbits. Leberco Laboratories. Assay 

No. 36327. Unpublished Report to International Flavors and Fragrances, Inc. Private communication to 

the Research Institute for Fragrance Materials, Inc. (as cited in EC, 2008).

Yamauchi, R., Ishibashi, H., Hirano, M., Mori, T., Kim, J. W., and Arizono, K. (2008). Effects of synthetic 

polycyclic musks on estrogen receptor, vitellogenin, pregnane X receptor, and cytochrome P450 3A gene 

expression in the livers of male medaka (Oryzias latipes). Aquat. Toxicol. 90(4), 261-268. 

Yang, J. J., and Metcalfe, C. D. (2006). Fate of synthetic musks in a domestic wastewater treatment plant 

and in an agricultural field amended with biosolids. Sci. Total Environ. 363(1-3), 149-165. 

Yin, J., Wang, H., Zhang, J., Zhou, N., Gao, F., Wu, Y., Xiang, J., and Shao, B. (2012). The occurrence of 

synthetic musks in human breast milk in Sichuan, China. Chemosphere 87, 1018-1023. 

Zhang, X., Liang, G., Zeng, X., Zhou, J., Sheng, G., and Ful, J. (2011). Levels of synthetic musk fragrances in 

human milk from three cities in the Yangtze River Delta in Eastern China. J. Environ. Sci. (China) 23(6), 

983-990. 
DRAFT DO NOT CITE OR QUOTE

Page 60

Appendices
DRAFT DO NOT CITE OR QUOTE

Page 61

Appendix A: Human Health Toxicity Studies, Biomonitoring 

Data, and Risk Assessments

The human health hazard reviewed and summarized in this section (see Table A-1) relies heavily 

on the hazard information extracted from the 2008 EU RAR on HHCB (EC, 2008). Other reviews 

are also available (OECD, 2009; US EPA, 2003; HERA, 2004a; 2004b). This review is not intended 

to be an exhaustive discussion of the toxicity of HHCB, but rather a summary of the available 

data. The reader is referred to the original documents for more detailed information. 

A-1. HUMAN HAZARD CHARACTERIZATION

Toxicokinetics (Absorption, Distribution, Metabolism, Excretion) 

Available toxicokinetics data indicate that HHCB has poor dermal absorption and is extensively 

metabolized and excreted, with no evidence of significant bioaccumulation. An in vivo human 

study reported approximately 0.1 percent dermal absorption following exposure to HHCB 

(Hawkins et al., 1996a; Ford et al., 1999; as cited in EC, 2008), while an in vitro dermal 

absorption study with human epidermal membranes showed 5.2 percent absorption (Green 

and Brian, 2001; as cited in EC, 2008). Human biomonitoring data show that HHCB has been 

detected in blood, milk, and adipose tissue following dermal exposures (see Table A-2). Data in 

humans after oral and inhalation exposures, including half-life data, are not available. In 

animals, dermal exposures in rats showed 16 percent absorption, with distribution to the 

stomach, liver, fat, plasma, adrenal glands, kidneys, and thyroid, with excretion primarily in the 

feces (Ford et al., 1999; Hawkins et al., 1995; as cited in EC, 2008). Intravenous studies in rats 

and pigs indicated that HHCB is rapidly distributed and excreted primarily in the feces in the rat 

and in the urine in pigs (Hawkins et al., 1997; Hawkins, 1997; as cited in EC, 2008). No evidence 

of accumulation in the feces or urine was seen in either the rat or pig. Apparent half-lives of 

elimination in the blood were longer in the rat than in the pig, with clearance from fat and skin 

considerably slower. HHCB and metabolites were found in the milk of pregnant and lactating 

rats exposed via the oral route (Hawkins et al., 1996a; as cited in EC, 2008). Toxicokinetics data 

in animals after oral and inhalation exposures are not available. 

Acute Toxicity

Several acute toxicity studies in rats and rabbits by the oral and dermal routes with HHCB are 

available (Minner and Foster, 1977; Moreno, 1975; as cited in EC, 2008). Acute toxicity studies 

by the inhalation route are not available. The acute toxicity of HHCB is low via the oral route in 

rats and low via the dermal route in rabbits, with LD50 values >3,000 mg/kg-bw. 

Subchronic/Repeated-Dose Toxicity

An oral range-finding study in rats with mean achieved intakes up to 829 mg/kg-bw/day 

reported signs of liver toxicity at 347 mg/kg-bw/day (lowest-observed-adverse effect 

concentration [LOAEL] (Api and Ford, 1999; Hopkins et al., 1996; as cited in EC, 2008); however, 

the follow-up 13-week repeated-dose oral (dietary) toxicity study showed no adverse, 

treatment-related effects on any parameters measured. The no-observed-adverse-effect level 
DRAFT DO NOT CITE OR QUOTE

Page 62

(NOAEL) for systemic toxicity was 150 mg/kg-bw/day, the highest dose tested (Api and Ford, 

1999; Hopkins et al., 1996; as cited in EC, 2008). In that same study, no effects on organ weight 

or histopathology were reported in the reproductive organs at any dose level. Repeated-dose 

subchronic toxicity studies by the inhalation and dermal routes were available, but were 

considered to be of limited value because they were not conducted according to guidelines or 

GLPs or they used fragrance mixtures in which HHCB was only present at low levels. 

Reproductive Toxicity and Fertility

No combined repeated-dose/reproductive/developmental or multigenerational reproductive 

toxicity studies on HHCB are available. Information on reproductive organs is available from 

the repeated-dose dietary toxicity study in rats; this study did not show any signs of toxicity to 

the reproductive organs examined at doses as high as ~150 mg/kg-bw/day. 

A modification to the ICH Guideline on Detection of Toxicity to Reproduction for Medicinal 

Products (http://www.ich.org/products/guidelines/safety/article/safety-guidelines.html) was 

conducted in rats in order to assess the potential for adverse effects to the neonate following 

exposure to HHCB through nursing. In this comprehensive study, the basis for the dose 

selection for this study was a pharmacokinetic analysis by Hawkins et al. (1996b) (as cited in EC, 

2008) that determined the oral doses required to produce levels in the milk of lactating rats 

similar to those reported in human milk and several orders of magnitude higher. At oral doses 

of 2 and 20 mg/kg-bw/day, HHCB was found in the milk of lactating dams at 2.28 and 32.8 

mg/L, respectively (see Human Biomonitoring below for comparison to levels in human breast 

milk). Five groups of pregnant Sprague-Dawley rats (28/dose) were administered 0, 2, 6, or 20 

mg/kg-bw/day HHCB (>95 percent purity) by gavage from gestation day (GD) 14 through 

weaning on lactation day 21 (Ford and Bottomley, 1997; Jones et al., 1996; as cited in EC, 2008). 

Dams were allowed to litter and rear their offspring to weaning. F1 offspring were exposed to 

HHCB in utero from GD 14 through lactation via the dam’s milk, and exposures ceased after 

weaning in the F1 offspring and in the F2 generation. Since milk consumption was not 

measured in the lactation study, actual intake of HHCB could not be determined. 

After parturition, the offspring were counted, sexed, weighed, and examined for external 

abnormalities. On postnatal day 4, the pups were weighed and the litters were culled to four 

males and four females. During the pre-weaning period, pups were examined for several 

developmental milestones, including surface righting reflex, startle reflex, air righting reflex, 

and pupil reflex. Selected offspring from these litters (24/sex/dose) were allowed to reach 

sexual maturity and were then assessed for behavioral changes in motor coordination and 

balance, activity, and avoidance. When the F1 generation offspring reached 84 days of age, 

they were mated (avoiding sibling pairings) and the reproductive capacity was assessed by 

examining the females before and after mating to determine time of pregnancy, estrous 

cyclicity, pre-coital time, pregnancy rates, and duration of gestation. F2 generation offspring 

were examined for abnormalities at parturition and then periodically until study termination on 

postnatal day 21. No adverse, treatment-related effects were reported in the dams or in either 

generation of offspring for any parameters assessed up to and including 20 mg/kg-bw/day. The 
DRAFT DO NOT CITE OR QUOTE

Page 63

NOAEL for maternal, reproductive, and developmental toxicity was 20 mg/kg-bw/day, based on 

no effects observed at the highest dose tested. 

Developmental Toxicity

In a pilot range-finding study of prenatal developmental toxicity (Christian et al., 1997; Christian

et al., 1999; as cited in EC, 2008), four groups of 25 pregnant Sprague-Dawley rats were 

exposed to 0, 100, 250, 500, and 1,000 mg/kg-bw/day HHCB (purity not specified) by gavage on 

GDs 7 to 17. Three dams were found dead at 1,000 mg/kg-bw/day. Maternal body weights 

were reduced (approximately 10 percent) in the three highest dosage levels. No other effects 

were reported in the dams. No signs of gross fetal malformations were reported; however, 

fetal body weights were reduced at the two highest doses. 

Based on the results of the dose range-finding study described above, pregnant SpragueDawley rats (25/dose) were administered 0, 50, 150, and 500 mg/kg-bw/day HHCB (>95 percent 

purity) by gavage on GDs 7 to 17 (Christian et al., 1997; Christian et al., 1999; as cited in EC, 

2008). The dams were observed for clinical signs of toxicity. Food and water consumption and 

body weights were recorded at various intervals. On GD 20, the dams were sacrificed and a 

gross necropsy was performed. The numbers of corpora lutea, implantation sites, resorption 

sites, live and dead fetuses, and live pup body weights were recorded. Fetuses were examined 

for sex and underwent a gross examination. One half of the fetuses was examined for visceral 

anomalies and the other half was examined for skeletal abnormalities. Signs of maternal 

toxicity in the dams consisted of significant increases in salivation, urine-stained fur, red or 

brown substance on the forepaws, and alopecia beginning at 150 mg/kg-bw/day, with 

statistical significance at 500 mg/kg-bw/day. At ≥150 mg/kg-bw/day, statistically significant 

reductions in food consumption and maternal body gain were observed, with the most severe 

reductions in body weight occurring during GDs 7 to 10. After completion of the dosing period 

(GDs 18 to 20), a significant rebound in food consumption and body weight gains occurred, with 

values reported as being comparable to controls. Signs of developmental toxicity were 

confined to the highest dose of 500 mg/kg-bw/day. Significant reductions in offspring body 

weight were reported at the high dose; however, according to the study authors, when 

severity, dose relationships, and historical ranges were taken into consideration, these body 

weight reductions were not definitively considered to be treatment-related. The number of 

litters or fetuses with morphological alterations (malformations and variations) reported at the 

high dose was not statistically significantly increased when compared to the control group. 

Malformations that did occur were few in number and were considered spontaneous in origin, 

except for the vertebral/rib malformations reported in only three fetuses that were not litter 

mates. No other parameters were reported to be affected by treatment at any dose level. The 

LOAEL for maternal toxicity was 150 mg/kg-bw/day, based on increases in clinical signs of 

toxicity and decreases in body weight, and the NOAEL was 50 mg/kg-bw/day. Although there is 

uncertainty concerning the effects reported in the offspring, a conservative LOAEL for 

developmental toxicity of 500 mg/kg-bw/day was concluded and the NOAEL was 150 mg/kgbw/day.
DRAFT DO NOT CITE OR QUOTE

Page 64

Genetic Toxicity

Following a wide array of in vitro and in vivo mutagenicity studies, HHCB did not induce gene 

mutations in bacterial cells or chromosome aberrations in mammalian cells (San et al., 1994; 

Api and San, 1999; Mersch-Sundermann et al., 1998a; 1998b; Kevekordes et al., 1997; Curry 

and Putman, 1995; Gudi and Ritter, 1997; San and Sly, 1994; Kevekordes et al., 1998; as cited in 

EC, 2008). 

Carcinogenicity

A 2-year cancer study is not available for HHCB. As an alternative, EPA evaluated the weight of the 

evidence from mutagenicity, subchronic, and endocrine studies, to support the conclusion that cancer 

data are not a critical data need.

Additional Information

Irritation and sensitization studies in humans and animals showed that HHCB was not corrosive, 

irritating, or sensitizing in rabbits and humans. No data on respiratory tract irritation were 

available. HHCB was considered to be a minimal eye irritant in rabbits and a possible photoirritant in rabbits and guinea pigs, but not in humans (Guillaume et al., 1973a; 1973b; 

Rubenkoenig and Ede, 1964; Lindstrum et al., 1978a; 1978b; 1978c; Harrison and Stolman, 

1986; Gabriel and Mark, 1987; Folk and Dammers, 1987; Shanahan and Alworth, 1987; Haynes, 

1984; 1985; 1986; Sauer, 1980; Levenstein, 1973a; 1973b; 1975a; 1975b; Wolven and 

Levenstein, 1963; Basketter, 1996; Parish, 1988; as cited in EC, 2008). 

In a series of in vitro assays with HHCB using different protocols designed to test for potential 

endocrine disruption (i.e., estrogenic and/or anti-estrogenic activity), HHCB showed weak 

estrogenic and anti-estrogenic activity in vitro, dependent on the estrogen receptor type 

(Seinen et al., 1999; Bitsch et al., 2002; Schreurs et al., 2002; Schreurs et al., 2004; Schreurs et 

al., 2005a; Schreurs et al., 2005b; Gomez et al., 2005; as cited in EC, 2008). Marginal repressing 

effects were also found in vitro on the androgen and progesterone receptor. However, no 

estrogenic effects were seen in the in vivo uterotrophic assay. 

Possible anti-estrogenic effects in zebrafish were assessed (Schreurs et al., 2004; as cited in EC, 

2008). Transgenic zebrafish were exposed to 0.01, 0.1, 1, and 10 μM HHCB with and without 

estradiol (E2). The highest concentration of 10 μM HHCB was toxic. Concentrations of 0.01, 

0.1, and 1 μM HHCB were not estrogenic. Dose-dependent antagonistic effects were reported 

at 0.1 and 1 μM levels.
DRAFT DO NOT CITE OR QUOTE

Page 65

Table A-1. Summary of Human Health Hazard Information.

Endpoints HHCB

CASRN (1222-05-5)

References

(as cited in EC, 2008)

Acute oral toxicity

LD50 (mg/kg-bw) >3,000

Minner and Foster (1977); Moreno (1975)

Acute dermal toxicity

LD50 (mg/kg-bw) >3,250

Minner and Foster (1977); Moreno (1975)

Acute inhalation toxicity

LC50 (mg/L)

No data

Repeated-dose toxicity 

Oral (mg/kg-bw/day) 

Dermal (mg/kg-bw/day)

Inhalation (mg/L/day)

NOAEL = 150 (highest dose tested)

No adequate data

No adequate data

Api and Ford (1999); Hopkins et al. (1996)

Reproductive toxicity

Oral (mg/kg-bw/day)

Maternal NOAEL/LOAEL

Reproductive NOAEL/LOAEL

Developmental NOAEL/LOAEL

NOAEL = 20 (highest dose tested)

NOAEL = 20 (highest dose tested)

NOAEL = 20 (highest dose tested)

Ford and Bottomley (1997); Jones et al. (1996)

Developmental toxicity 

Oral (mg/kg-bw/day), rats

Maternal NOAEL/LOAEL

Developmental NOAEL/LOAEL

NOAEL = 50/LOAEL = 150

NOAEL = 150/LOAEL = 500

Christian et al. (1997); Christian et al. (1999)

Genetic toxicity, gene mutation

In vitro

Genetic toxicity, gene mutation

In vivo

Negative

Negative

San et al. (1994); Api and San (1999); MerschSundermann et al. (1998a); Mersch-Sundermann et 

al. (1998b)

Genetic toxicity, chromosomal aberrations

In vitro

Genetic toxicity, chromosomal aberrations

In vivo

Negative 

Negative 

Api and San (1999); Kevekordes et al. (1997); Curry 

and Putman (1995); Gudi and Ritter (1997); San and 

Sly (1994); Kevekordes et al. (1998)
DRAFT DO NOT CITE OR QUOTE

Page 66

Table A-1. Summary of Human Health Hazard Information.

Endpoints HHCB

CASRN (1222-05-5)

References

(as cited in EC, 2008)

Additional information

Corrosivity/skin-eye Irritation/sensitization

Respiratory tract irritation

Carcinogenicity

Endocrine disruption

In vitro

In vivo

Negative

No data 

No data

Weak

Negative

See summary above
DRAFT DO NOT CITE OR QUOTE

Page 67

A-2. HUMAN BIOMONITORING

Biomonitoring data describing human exposure levels of HHCB are summarized in Table A-2. 

Data are preliminary and limited based on small numbers of subjects, limited geographic 

representation, high variability, and incomplete details on methodology.

Several small studies have reported concentrations of HHCB in limited numbers of the general 

population in Asia and Europe. Only two small studies have reported exposure levels in the US. 

To date, HHCB has been measured in adipose tissue, blood, breast milk, and umbilical cord 

blood. Most of these studies reported data for <100 samples and several of them did not 

provide data tables or details on the results of the data. However, HHCB was detected in a high 

majority of the samples collected in the studies reported here. Therefore, it can be assumed 

that exposure is widespread. 

Both US studies took place in 2004 in the northeastern part of the US. A study on adipose 

tissue in 49 residents of New York City undergoing liposuction reported a mean concentration 

of 178 ng/g lipid weight (lw) (range: 12 to 798 ng/g lw) (Kannan et al., 2005), which falls 

between the reported concentrations of HHCB in adipose tissue from two other available 

studies (Moon et al., 2012; Schiavone et al., 2010). A study of 43 Korean women reported a 

mean of 81 ng/g lw in adipose tissue (range: 28 to 211 ng/g lw) (Moon et al., 2012), while a 

small study in Italy reported a mean of 361 ng/g lw (range: 28 to 211 ng/g lw) (Schiavone et al., 

2010).

Several small studies on concentrations of HHCB in breast milk have been conducted. In the US, 

a mean of 227 ng/g lw was reported for 39 samples collected in Massachusetts (Reiner et al., 

2007b). The levels reported in this small study are much higher than those reported in other 

countries to date (see Table A-2). However, in order to compare the results, the timing of the 

collection of the samples must be considered (length of time after birth), and it is not clear from 

the report when the US samples were collected. 

No studies in the literature reported blood concentrations of HHCB in the US. A collection of 

matched maternal and cord blood samples in the Netherlands indicated high detection and 

correlation between maternal and cord blood serum. Concentrations reported in studies in the 

Netherlands, Austria, and China are so variable that they cannot and should not be compared 

(see Table A-2). Little data were provided in these studies, which were primarily done in 2005. 

Given that HHCB is sequestered in the fat, blood may not be an appropriate matrix for 

biomonitoring. 

High variance in measurements may be a result of the analytical methods for measuring HHCB 

in human matrices as they are still being developed and modified Because the sample sizes are 

very small, the results should not be extrapolated to larger populations. Also, data from other 

countries may not represent those of the US since exposure patterns may vary greatly between 

countries, especially for personal care products such as those that contain HHCB. 
DRAFT DO NOT CITE OR QUOTE

Page 68

Table A-2. Human Biomonitoring Data for HHCB.

Population Sampling Year(s) Levels Reference

New York City residents 

undergoing liposuction; 

n = 49 (12 males, 

37 females)

2003-2004

HHCB detected in 

adipose tissue in all 

samples

Mean: 97 ± 88 ng/g ww

Range: 6.1-435 ng/g ww

Mean: 178 ± 166 ng/g lw

Range: 12-798 ng/g lw

Kannan et al. (2005)

n = 43 women 

undergoing laparoscopy

surgery for myoma 

2007-2008

100 percent detection 

in adipose tissue

Mean: 81 ± 44 ng/g lw

Range: 28-211 ng/g lw

Moon et al. (2012)

n = 12 surgical samples, 

Siena, Italy, 3 females, 

9 males

2005-2006

HHCB detected in 

92 percent of adipose 

tissue samples

Mean: 361 ± 467 ng/g lw Schiavone et al. (2010)

n = 39 milk samples from 

Massachusetts women

2004

HHCB detected in 

97 percent of breast 

milk samples; sample 

collection time not 

reported

Mean ± SD: 227 ± 228 ng/g 

lw

Range: <5-917 ng/g lw

Reiner et al. (2007b)

10 primiparous mothers 

(25-29 years old ) 

1999

2003/2004

HHCB detected in all 

breast milk samples; 

collected 14-26 weeks 

after birth

Median: 147 μg/kg fat

Range: 38-422 μg/kg fat

Mean: 179 ± 111 μg/kg

Large variability in individual 

samples

Duedahl-Olesen et al.

(2005)

n = 101 random samples 

of 266 breast milk 

samples collected from 

44 primiparous women, 

Uppsala County

1996-2003, 14-21 days 

postpartum

HHCB detected in all 

breast milk samples

Median: 63.9 ng/g lw

Range: 2.8-268 ng/g lw

Lignell et al. (2008)

40 mothers (24-38 years 

old), Munich

1997-1998; HHCB 

detected in 

35/40 breast milk 

samples

Median: 64 ng/g lw

Mean: 115 ng/g lw

Range: 21-1,316 ng/g lw

Liebl et al. (2000)

Seoul, Korea

n = 20 volunteers, 

>25 years old, 17 breast 

milk samples, 

14 umbilical cord 

samples; 18 maternal 

serum samples

2007; breast milk 

samples collected 

3-10 days after

delivery; HHCB 

detected in 

100 percent breast 

milk samples; 

70 percent cord blood 

samples; 90 percent 

maternal serum 

samples

Breast milk mean: 

0.055-0.515 ng/g lw

Umbilical cord blood: 

0.67-2.7 ng/g lw

Maternal serum: 

0.17-1.4 ng/g lw

Kang et al. (2010)
DRAFT DO NOT CITE OR QUOTE

Page 69

Table A-2. Human Biomonitoring Data for HHCB.

Population Sampling Year(s) Levels Reference

n = 110 mothers in 

southwest China taking 

children for vaccines; 

110 breast milk samples

2009 Median: 11.5 ng/g lw

Range: <1.1-456.7 ng/g lw

Yin et al. (2012)

n = 10 breast milk 

samples from Chengdu, 

China; validating 

analytical method

2009

HHCB detected in 

breast milk, all 

samples

Range: 11.7-67.6 ng/g lw Wang et al. (2011)

100 volunteers in 3 cities 

in Yangtze River Delta, 

Shanghai, Shaoxing, and 

Wuxi

2006-2007

Breast milk samples 

collected 1-2 weeks 

after delivery;

99 percent detection

Median: 63 ng/g

Ranges: 

Shanghai: <5-278

Shaoxing: 5-782

Wuxi: 24-281 ng/g lw

Zhang et al. (2011)

Serum samples from 

volunteers in the 

Netherlands

n = 42 maternal serum 

samples, 27 cord blood 

samples

2005 report

HHCB detected in 

38/42 maternal serum 

samples and 

26/27 cord blood 

samples 

Maternal: 0.15-3.2 ng/g 

serum

Cord blood: 0.11-1.6 ng/g 

serum

Peters (2005)

Students at Medical 

University of Vienna

n = 100 plasma samples 

(55 female, 45 male); 

19-43 years old

2005 report

Detection in 

91 percent of plasma 

samples; women had 

higher levels than men 

but lower values in 

26-43 year olds

Plasma concentrations:

Males, median: 260 ng/L

 Range: 98-540 ng/L

Females, median: 580 ng/L

 Range: 290-885 ng/L

Hutter et al. (2005)

Dept of Angiology at 

Hanusch-Krankenhaus, 

Vienna;

n = 53 women >50 years 

old (51-71 years)

Detection in

89 percent of plasma 

samples; older women 

had higher 

concentrations of 

HHCB

Maximum plasma 

concentration: 6,900 ng/L

No other data provided

Hutter et al. (2010)

11 cities in China

n = 204 (94 female, 

110 male) 17-75 years 

old (median 25 years)

98 percent detection 

in blood samples 

(almost all above LOQ)

Whole blood

Median: 0.85 μg/L

Maximum: 1.63 μg/L

No data tables provided

Hu et al. (2010)
DRAFT DO NOT CITE OR QUOTE

Page 70

A-3. SUMMARY OF 2008 EU HUMAN HEALTH RISK ASSESSMENT

The human health hazards and risks of HHCB have been extensively assessed by several other 

reliable entities (EC, 2008; OECD, 2009; US EPA, 2003; HERA, 2004a; 2004b). Each of these welldocumented, independent reviews concluded that there was no significant risk to human 

health from exposure to HHCB as used in household cleaning products. Moreover, the overall 

conclusions of the EU RAR (EC, 2008) were that there was no need for further information 

and/or testing and no need for risk reduction measures beyond those already being applied. 

The EPA has thoroughly reviewed these other assessments and presents a summary of the 

most recent EU RAR. 

Assumptions and Points of Departures Used in the EU RAR

The EU RAR (EC, 2008) assessed risk in three populations for relevant endpoints: workers

(exposures from handling during production and dilution, compounding, formulation, and 

professional cleaning); consumers (exposures from a wide variety of consumer products such as 

perfumes, creams, toiletries, soaps, shampoos, and household and laundry cleaning products); 

and humans exposed via the environment (exposures via food and water including fish, root 

crops, and mother’s milk). 

Dermal absorption of 16 and 5.2 percent in rats and humans, respectively, was used. Since 

toxicokinetic data of HHCB following oral and inhalation absorption are not available in humans 

or animals, intermediate defaults of 50 percent for oral absorption and 100 percent absorption 

from inhalation exposures were used. 

For general systemic effects following repeated exposures, the oral NOAEL of 150 mg/kgbw/day from the 13-week subchronic toxicity study in rats was used as a point of departure. 

Since NOAELs are not available following dermal or inhalation exposures, the administered 

dose from the oral 13-week toxicity study was converted into an internal dose (body burden) by 

taking into account different absorption factors. Based on an oral absorption value of 

50 percent in humans and animals, an internal NOAEL of ≥75 mg/kg-bw/day was used, with 

100 percent absorption assumed for inhalation exposures and 5.2 percent absorption for 

dermal exposures.

For developmental/reproductive toxicity, the NOAEL of 20 mg/kg-bw/day (the highest dose 

tested) from the peri/post-natal oral (gavage) toxicity study was used as a point of departure 

(Ford and Bottomley, 1997; Jones et al., 1996; as cited in EC, 2008). Based on an oral 

absorption rate of 50 percent, an internal NOAEL of ≥10 mg/kg-bw/day was used, with 

100 percent absorption assumed for inhalation exposures and 5.2 percent absorption assumed 

for dermal exposures. 

Worst-case estimates were based upon combined (simultaneous) dermal and inhalation 

exposures for compounding workers only (the only scenario relevant for combined exposures). 
DRAFT DO NOT CITE OR QUOTE

Page 71

Risk to Workers

Oral exposures in workers were assumed to be mitigated by personal hygiene measures, and 

therefore, only risks following dermal and inhalation exposures in the workplace were assessed. 

For acute toxicity following dermal exposures, high LD50 values (>3,000 mg/kg-bw) indicated no 

concern for workers. Although acute toxicity studies by the inhalation route are not available, 

low-level, short-term exposures in workers, combined with the low acute toxicity by the oral 

route, suggested no concern for workers for acute toxicity by the inhalation route. 

Given that HHCB showed no potential for skin/eye irritation, corrosivity, or sensitization/ 

photosensitization, it was concluded that there was no concern for local effects in workers 

following repeated exposures. Likewise, since there is a lack of skin and eye irritation potential, 

no significant respiratory tract irritation potential was expected. 

For both the dermal and inhalation exposure scenarios for the general systemic toxicity 

endpoint, a minimal margin of safety (MOS) of 100 was used for comparison. The MOS was 

based on an interspecies factor of 10 (4 for metabolic size differences and 2.5 for remaining 

differences), an intraspecies factor of 5, and a factor of 2 for extrapolation from semi-chronic to 

chronic exposure. For each inhalation/dermal exposure scenario, the calculated MOS was well 

above the minimal MOS of 100 (≥2,600). Combined exposure routes for total body burdens for 

skin contact and inhalation for the compounding scenario resulted in MOS values well above 

100 (≥2,000). As a result, it was concluded that for workers under these exposure scenarios for 

general systemic toxicity, there is, at present, no need for further information and/or testing 

and no need for risk reduction measures beyond those already being applied. 

Given that HHCB was not found to be mutagenic in a wide array of studies, it was concluded 

that for mutagenicity for workers under these exposure scenarios, there is, at present, no need 

for further information and/or testing and no need for risk reduction measures beyond those 

already being applied. Although there are no carcinogenicity data for HHCB, it was concluded 

that there was no concern for workers under these exposure scenarios for carcinogenicity 

based on negative mutagenicity data and the lack of any apparent structural alerts that would 

raise a concern. 

For both the dermal and inhalation exposure scenarios for the developmental/reproductive 

toxicity endpoint, a minimal MOS of 50 was used for comparison. The MOS was based on an 

interspecies factor of 10 (4 for metabolic size differences and 2.5 for remaining differences) and 

an intraspecies factor of 5. For dermal, inhalation, and combined exposure scenarios, the 

calculated MOS was above 50 (≥261). As a result, it was concluded that for workers under 

these exposure scenarios for developmental/reproductive toxicity, there is, at present, no need 

for further information and/or testing and no need for risk reduction measures beyond those 

already being applied. 

 
DRAFT DO NOT CITE OR QUOTE

Page 72

Risk to Consumers

The main route of exposure for consumers was assumed to be dermal, with some inhalation 

exposures and no oral exposures. 

Given that HHCB showed no potential for skin/eye irritation, corrosivity, or sensitization/ 

photosensitization, it was concluded that there was no concern for local effects in consumers 

following repeated exposures. Since there is a lack of skin and eye irritation potential, no 

significant respiratory tract irritation potential was expected. 

 

For general systemic effects following repeated exposures, a MOS of 200 (based on a factor of 

10 for intraspecies differences, 4*2.5 for interspecies differences, 2 for duration extrapolation, 

and 1 for dose-response) was used. The calculated MOS for dermal exposure was well above 

200 (≥1,400). Therefore, it was concluded that there was no concern for general systemic 

toxicity in consumers following repeated dermal exposures. 

Given that HHCB was not found to be mutagenic in a wide array of studies, it was concluded 

that for mutagenicity for consumers, there is, at present, no need for further information 

and/or testing and no need for risk reduction measures beyond those already being applied. 

Likewise, although there are no carcinogenicity data for HHCB, it was concluded that there was 

no concern for consumers under these exposure scenarios for carcinogenicity based on 

negative mutagenicity data and the lack of any apparent structural alerts that would raise a 

concern. 

For the dermal exposure scenario for the developmental/reproductive toxicity endpoint, a 

minimal MOS of 100 was used for comparison. The MOS was based on an intraspecies factor of 

10, an interspecies species factor of 10 (4 for metabolic size differences and 2.5 for remaining 

differences), and factor of 1 for dose-response. The calculated MOS was above 100 (MOS ≥189). 

It was concluded that for consumers with a dermal exposure scenario for developmental/ 

reproductive toxicity, there is, at present, no need for further information and/or testing and no 

need for risk reduction measures beyond those already being applied. 

Risk to Humans Exposed Indirectly via the Environment

Exposure to HHCB by the inhalation route (via air) was considered to be negligible. The main 

route of exposure to humans in the environment was oral (via fish and root crops). Exposures 

were based on local and regional daily intake estimates following production. A separate risk 

characterization for breast-fed babies was conducted for exposure via mother’s milk (see next 

section for more information).

For general systemic effects following repeated exposures, an internal NOAEL of ≥75 mg/kgbw/day and a MOS of 200 (based on a factor of 10 for intraspecies differences, 4*2.5 for 

interspecies differences, 2 for duration extrapolation, and 1 for dose-response) were used. The 

calculated MOS for exposure via food was >3E+4 for the local production scenario and 

>8E+5 for the regional production scenario. Therefore, it was concluded that for the local and 
DRAFT DO NOT CITE OR QUOTE

Page 73

regional exposure scenarios, there was no concern for general systemic toxicity in humans 

exposed indirectly via the environment. 

As with the other populations, since HHCB was not found to be mutagenic in a wide array of 

studies, it was concluded that for mutagenicity, there is, at present, no need for further 

information and/or testing and no need for risk reduction measures beyond those already 

being applied. Likewise, although there are no carcinogenicity data for HHCB, it was concluded 

that there was no concern for carcinogenicity based on negative mutagenicity data and the lack 

of any apparent structural alerts that would raise a concern. 

For the oral exposure scenario for the developmental/reproductive toxicity endpoint, an internal 

NOAEL of ≥10 mg/kg-bw/day and a MOS of 100 were used. The MOS was based on an 

intraspecies factor of 10, an interspecies species factor of 10 (4 for metabolic size differences 

and 2.5 for remaining differences), and factor of 1 for dose-response (based on the lack of effect 

at the highest dose tested). The calculated MOS values for local and regional oral exposure 

scenarios were 3,846 and 10,000, respectively, indicating no concern for developmental/ 

reproductive toxicity in humans exposed indirectly via the environment. 

Assessment of Risk for Breast-Fed Babies Exposed via Mother’s Milk

The presence of HHCB in human milk was considered to be the result of indirect environmental 

exposures from a variety of sources, including maternal exposure to consumer products; intake 

via food, water, or air; and occupational exposures. 

The concentrations of HHCB in milk samples for humans and rats were compared. By using the 

dose level in dams (20 mg/kg-bw/day, the highest dose tested from the oral peri/postnatal 

developmental toxicity study), the measurement of levels of HHCB in the milk of dams 

(17.6 and 5.0 μg/mL at four or eight hours post dosing, respectively), with the maximum level of 

HHCB found in human milk samples (1,316 μg/kg fat), the EU RAR (EC, 2008) concluded that 

pups in the high-dose group were estimated to be exposed to levels approximately 100 to 

360 times the maximum level found in human milk samples. 

The amount of milk consumed by infants and rat pups was compared. Assuming 50 percent 

absorption of the ingested HHCB, the average daily uptakes (ADUs) for the breast-feeding infant

for 0 to 3 and 3 to 12 months, as well as the average daily milk consumption for the rat pup, 

were estimated. A comparison of these two estimates of uptake showed a difference of 

approximately three orders of magnitude between the levels of HHCB exposure in the rat (in 

which no adverse effects were found) and the human infant exposure. 

Total internal (worst-case) combined exposures were estimated and compared to the internal 

NOAELs for general systemic toxicity following repeated exposures and for developmental/ 

reproductive toxicity (≥75 and ≥10 mg/kg-bw/day, respectively) in order to calculate MOS 

values. The worst-case combined exposure was estimated from the sum of the worst-case 

estimates from three populations: (1) dermal and inhalation exposures for scenario 2 for 

compounding workers; (2) dermal and inhalation exposures for consumers; and (3) oral and 
DRAFT DO NOT CITE OR QUOTE

Page 74

inhalation, locally via the environment. For general systemic effects following repeated 

exposures, a MOS of 100 was used (based on 4*2.5 for metabolic size and other differences, a 

factor of 5 for intraspecies differences, and a factor of 2 for semichronic to chronic exposure 

extrapolation). The calculated MOS was well above 100 (798). For the developmental/

reproductive toxicity endpoint, a MOS of 50 (based on 4*2.5 for interspecies species differences 

and 5 for intraspecies differences) was used. The calculated MOS was above 50 (≥106). 

As a result, the EU concluded that, overall, there was no concern for breast-fed babies exposed 

indirectly via the environment and no need for further information and/or testing or risk 

reduction measures beyond those already being applied.

A-4. KEY SOURCES OF UNCERTAINTY AND DATA LIMITATIONS ON HUMAN HEALTH

Overall, adequate screening-level animal toxicity data are available to characterize the human 

health hazard for HHCB. Toxicokinetics (by the dermal route), acute, repeated-dose, and 

developmental toxicity data are available to characterize the human health hazard of HHCB. 

Although no multigenerational reproductive toxicity studies on HHCB are available, information 

on developmental and reproductive toxicity was obtained from both the repeated-dose dietary 

toxicity study (reproductive organ data) and the peri/post-natal reproductive toxicity study with 

modified exposures. Several assays testing for endocrine disruption, genotoxicity, and 

irritation/sensitization (including several studies in humans) are also available. The database is 

incomplete for toxicokinetic data in animals by the oral and inhalation routes; acute toxicity 

data by the inhalation route; repeated-dose toxicity data by the inhalation and dermal routes; 

and chronic toxicity/carcinogenicity. Data in humans on most toxicity endpoints are not 

available. 

One area of uncertainty concerns the effects reported in offspring in the prenatal 

developmental toxicity study in rats (Christian et al., 1997; Christian et al., 1999; as cited in EC, 

2008). When severity, dose relationships, and historical ranges were taken into consideration, 

the reductions in pup body weight were not definitively considered by the study authors to be 

treatment-related. Additionally, fetal malformations observed were reported in only three 

fetuses from separate litters. Even though there is uncertainty surrounding these endpoints, 

the study authors, as well as other reliable entities, have concluded a conservative LOAEL for 

developmental toxicity of 500 mg/kg-bw/day from this study.

Another area of uncertainty concerns the period of dosing in the peri/post-natal reproductive 

toxicity study (Ford and Bottomley, 1997; Jones et al., 1996; as cited in EC, 2008). This study 

was designed to determine effects of HHCB on the neonate when exposed through nursing. 

Exposures in the F1 offspring occurred in utero from GD 14 through lactation; there were no 

exposures to HHCB beginning from weaning in the F1 offspring throughout the F2 generation. 

Dosing of pregnant animals should typically include the entire period of organogenesis, which in 

the case of the rat is GDs 6 to 15. Although dosing in this study began towards the end of 

organogenesis, it was considered to be adequate for screening-level purposes in order to 

characterize developmental toxicity (including endpoints such as pup weight, pup survival, and 
DRAFT DO NOT CITE OR QUOTE

Page 75

postnatal death) and reproductive toxicity (reproductive performance) as well as a battery of 

developmental neurotoxicity tests. Other reliable entities such as the 2008 EU RAR on HHCB 

(EC, 2008) and the 2009 SIDS Initial Assessment Profile (OECD, 2009) also concluded that this 

study was adequate for assessing hazard for these endpoints. 

Human biomonitoring data indicate that HHCB is present in milk, fat, and blood, but the studies 

in the US are preliminary and of limited value for characterizing exposures. Specifically for the 

US, there are few biomonitoring studies in limited locations with small numbers of samples. 

HHCB has not been measured as part of the National Health and Nutrition Environmental 

Survey (NHANES), so the data are not representative. Exposures in the US are assumed to be 

similar to those in Europe, but it is unknown how manufacture and use of products containing 

HHCB may differ between regions or cultures since use of fragrances, personal care, and 

cleaning conventions may vary.

A-5. CONCLUSIONS OF HUMAN HEALTH ASSESSMENT

The available toxicokinetic data indicate that HHCB is poorly absorbed through the skin; 

toxicokinetic data for the oral and inhalation routes are not available. Limited pharmacokinetic 

data have reported HHCB metabolites in the milk of pregnant and lactating rats. Human 

biomonitoring data have reported HHCB in milk, fat, and blood, but the studies are preliminary 

and of limited value for characterizing exposures. The acute toxicity by the oral and dermal

routes is low. Signs of systemic toxicity following repeated exposures have not been reported. 

None to weak estrogenic and anti-estrogenic activity has been demonstrated. HHCB is not 

mutagenic, corrosive, irritating, or sensitizing; however, minimal eye irritation and possible 

signs of photoirritation have been reported. Carcinogenicity data are not available.

HHCB was initially selected for review based on a moderate hazard concern for developmental 

toxicity and a high potential for exposure. However, following further review of the 

developmental and reproductive toxicity data, and taking into account several lines of 

evidence, the conclusion is an overall low concern for developmental toxicity. The uncertainty 

surrounding the pup body weights seen at 500 mg/kg-bw/day from the prenatal developmental 

toxicity study, the occurrence of only three fetuses from separate litters exhibiting fetal 

malformations at 500 mg/kg-bw/day from this same study, and the lack of fetal morphological 

changes observed in the pilot studies, even at maternally lethal dosages (up to 1,000 mg/kgbw/day) lead to the conclusion that 500 mg/kg-bw/day is a conservative LOAEL. No effects 

were reported in fetuses in the perinatal toxicity study with lactational exposures at doses 

similar to those reported in human milk and several orders of magnitude higher (20 mg/kgbw/day). Overall, a low concern for developmental toxicity is indicated by the data. 

Additionally, the human health hazards and risks of HHCB have been extensively reviewed and 

assessed by several other reliable entities, most recently the EU, and it has been concluded that 

the overall concern for human health hazards, including that for developmental toxicity, is low. 
DRAFT DO NOT CITE OR QUOTE

Page 76

The review of human health hazard studies, biomonitoring studies in the US and elsewhere, 

and the EU RAR showed an overall low risk concern for human health, including the risk for 

developmental toxicity, from current use of HHCB. 
DRAFT DO NOT CITE OR QUOTE

Page 77

Appendix B: Chemical Synthesis of HHCB

O

O

O

+

pentamethylindane (1)

H

1,1,2,3,3-pentamethyl-5-

(-hydroxyisopropyl)indane (2)

HCHO/

Galaxolide

H

.alpha.-methyl styrene 2-methyl-2-butene

propylene oxide

Figure 2-2. Chemical Synthesis of HHCB.

HHCB is manufactured by a three-step reaction (Ullmann, 2003; Zviely, 2002). First, a cycloaddition reaction of 

.alpha.-methyl styrene and 2-methyl-2-butene (i.e., amylene) is performed under acidic conditions to obtain 

1,1,2,3,3-pentamethylindane (1). Second, the pentamethylindane (1) is then hydroxyalkylated with propylene 

oxide in a Friedel-Crafts reaction using aluminum chloride as a catalyst. Third, the ring closure of the resulting 

1,1,2,3,3-pentamethyl-5-(-hydroxyisopropyl)indane (2) to 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-

hexamethylcyclopenta-γ-benzopyran (HHCB; Galaxolide) is accomplished with paraformaldehyde and a lower 

aliphatic alcohol via the acetal or with paraformaldehyde and a carboxylic acid anhydride via the acylate.

1) Ullmann's Encyclopedia of Industrial Chemistry (2003). 6th ed.Vol 1: Federal Republic of Germany: WileyVCH Verlag GmbH & Co. 2003 to Present, p. V14 145.

2) Zviely M; Kirk-Othmer Encyclopedia of Chemical Technology. (2005). NY, NY: John Wiley & Sons; Aroma 

Chemicals. Online Posting Date: Jan 25, 2002.
DRAFT DO NOT CITE OR QUOTE

Page 78

Appendix C: HHCB Diastereoisomers (#1 to #6) and Related 

Structural Analogs (#7 to #15)

Table B-1. HHCB Diastereoisomers and Related Structural Analogs.

CASRN Chemicals Structure Chemical Index Name On TSCA 

Inventory

1 172339-62-7 Cyclopenta[g]-2-benzopyran,1,3,4,6,7,8-

hexahydro-4,6,6,7,8,8-hexamethyl-, 

(4S,7S)-

N

2 172339-63-8 Cyclopenta[g]-2-benzopyran,1,3,4,6,7,8-

hexahydro-4,6,6,7,8,8-hexamethyl-, 

(4R,7S)-

N

3 252332-95-9 Cyclopenta[g]-2-benzopyran,1,3,4,6,7,8-

hexahydro-4,6,6,7,8,8-hexamethyl-, 

(4S,7R)-

N

4 252332-96-0 Cyclopenta[g]-2-benzopyran,1,3,4,6,7,8-

hexahydro-4,6,6,7,8,8-hexamethyl-, 

(4R,7R)-

N

5 252933-48-5 Cyclopenta[g]-2-benzopyran,1,3,4,6,7,8-

hexahydro-4,6,6,7,8,8-hexamethyl-, 

(4R,7R)-relN
DRAFT DO NOT CITE OR QUOTE

Page 79

Table B-1. HHCB Diastereoisomers and Related Structural Analogs.

CASRN Chemicals Structure Chemical Index Name On TSCA 

Inventory

6 252933-49-6 Cyclopenta[g]-2-benzopyran,1,3,4,6,7,8-

hexahydro-4,6,6,7,8,8-

hexamethyl-, (4R,7S) relN

7 1222-06-6 Cyclopenta[g]-2-benzopyran, 

1,3,4,6,7,8-hexahydro-4,4,6,6,8,8-

hexamethylN

8 857091-61-3 Cyclopenta[g]-2-benzopyran, 

1,3,4,6,7,8-hexahydro-3,6,6,7,8,8-

hexamethylN

9 102296-64-0 Cyclopenta[g]-1-benzopyran, 

2,3,4,6,7,8-hexahydro-4,4,6,6,8,8-

hexamethylN

10 135546-43-9 Cyclopenta[g]-1-

benzopyran,2,3,4,6,7,8hexahydro4,6,6,7,8,8-hexamethylN

11 135546-42-8 Cyclopenta[g]-2-benzopyran, 

1,3,4,6,7,8-hexahydro-1,6,6,7,8,8-

hexamethylN

12 114109-63-6 Cyclopenta[h]-2-benzopyran, 

1,3,4,7,8,9-hexahydro-4,7,7,8,9,9-

hexamethylN
DRAFT DO NOT CITE OR QUOTE

Page 80

Table B-1. HHCB Diastereoisomers and Related Structural Analogs.

CASRN Chemicals Structure Chemical Index Name On TSCA 

Inventory

13 114109-62-5 Cyclopenta[f][2]benzopyran, 1,2,4,7,8,9-

hexahydro-1,7,7,8,9,9-hexamethylN

14 78448-48-3 Cyclopenta[g]-2-benzopyran, 6-ethyl1,3,4,6,7,8-hexahydro-4,6,8,8-

tetramethylN

15 78448-49-4 Cyclopenta[g]-2-benzopyran, 8-ethyl1,3,4,6,7,8-hexahydro-4,6,6,8-

tetramethylN
DRAFT DO NOT CITE OR QUOTE

Page 81

Appendix D: IUR Data for HHCB

The information in Tables D-1 through D-4 is from EPA’s non-CBI IUR database (EPA, 2010) for 

the 2006 reporting cycle, for HHCB, CASRN 1222-05-5. The chemical name used in the IUR 

database is cyclopenta[g]-2-benzopyran, 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethyl-. 

Table D-1. IUR National HHCB Information.a

Production volume (aggregated) 1-<10 million lbs

Maximum concentration (at manufacture or import site) >90 percent

Physical form(s): Liquid

Number of manufacturing, processing, and use sites 

(aggregated)b

100-999

Number of reasonably likely to be exposed industrial 

manufacturing, processing and use workers (aggregated)b

100-999

Was industrial processing or use information reported? Yes

Was commercial or consumer use information reported? Yes

a

More detailed information was not publically available on the IUR website because it was considered to be CBI. b

The IUR website states: “Data may be greatly underestimated. Please see Overview of 2006 TSCA Inventory 

Update Reporting (PDF).”

Table D-2. HHCB IUR Process and Industry Information.

Type of Processing Industrial Sector (Based on NAICS) Industrial Function

Processing—incorporation into 

formulation, mixture, or reaction 

product

All Other Chemical Product and 

Preparation Manufacturing

Not readily obtainable

Processing—incorporation into 

formulation, mixture, or reaction 

product

All Other Chemical Product and 

Preparation Manufacturing

Odor agents

Processing—incorporation into 

formulation, mixture, or reaction 

product

All Other Food Manufacturing Not readily obtainable

Processing—incorporation into 

formulation, mixture, or reaction 

product

Soap and Cleaning Compound 

Manufacturing

Odor agents
DRAFT DO NOT CITE OR QUOTE

Page 82

Table D-3. HHCB IUR Consumer Information.

Commercial/Consumer Product 

Category

Maximum Concentration in Related 

Consumer/Commercial Product 

Category

Intended for Use in 

Children’s Products in 

Related Product Category

Soaps and detergents Not readily obtainable Yes

Other Not readily obtainable Not readily obtainable

Other Not readily obtainable No

Table D-4. IUR Company Site Information.

Company Site City State Zip Code Manufacture Import Site Limited

Berje, Inc. Berje, Inc. Bloomfield NJ 07003 No Yes N/A

International 

Flavors & 

Fragrances, 

Inc.

Ashland, Inc. Morrisville PA 19067-3701 No Yes N/A

International 

Flavors & 

Fragrances, 

Inc.

International 

Flavors & 

Fragrances, 

Inc. - Hazlet

Hazlet NJ 07730-1797 No Yes N/A

Quest 

International 

Flavors and 

Fragrances

Quest 

International 

Flavors and 

Fragrances

Mt. Olive NJ 07828 No Yes N/A

S C Johnson 

& Son, Inc.

S C Johnson 

& Son, Inc.

Sturtevant WI 53177 — — —

Source: From public data resulting from the IUR Reporting Cycle (2006) 

http://www.epa.gov/iur/tools/data/index.htm. 
DRAFT DO NOT CITE OR QUOTE

Page 83

Appendix E: Modeled Release Estimates According to Stage of 

Production and Use

This Appendix contains modeling estimates of releases of HHCB to the environment from 

industrial sites of all stages of the HHCB life cycle and estimates of releases from consumer and 

commercial use of end-use products. 

In general, the life cycle of a fragrance ingredient such as HHCB includes:

• Manufacturing of the chemical substance,

• Compounding of fragrance oils containing the chemical substance,

• Blending of fragrance oils containing the chemical substance into commercial and 

consumer products (e.g., perfumes, cosmetics, soaps, detergents), 

• Use of these commercial and consumer products.

All release estimates are based on the estimated 2011 use volume of 1,700 metric tons 

(3.74 million lbs) for HHCB. This volume represents the vast majority of the quantity of HHCB 

used to prepare fragrance oils or fragrance compounds within the United States, based on data 

collected from IFRA’s Volume of Use Survey (IFRA, 2012d). 

The release estimates are summarized in Table E-1. These results indicate that the majority of 

the use volume of HHCB (>90%) is released to WWTP influent as a result of consumer and 

commercial use of HHCB-containing products. For industrial releases, compounding resulted in 

the highest per-site release due to a relatively low number of domestic compounding sites. A 

discussion of the release estimates according to stages of the life cycle follows. 

Estimated Release from Manufacture and Import

As discussed in Chapter 2, HHCB is not currently manufactured in the US (IFRA, 2012a). For this 

release assessment, four companies are assumed to import HHCB to the US at five sites (see 

Appendix D, Table D-4). Releases are not expected to result from import activities, but may 

occur at import sites if HHCB is also diluted and compounded onsite after import, as further 

discussed below.
DRAFT DO NOT CITE OR QUOTE

Page 84

Table E-1. Summary of Estimated Environmental Releases.

Release Activity Release Media

Release Factor 

(Amount 

Released per 

Amount Used, 

Percent)

Number 

of Sites

Number of 

Release Days 

at Each Site 

(Days/Year)

Daily 

Release

(kg/SiteDay)

Combined Annual 

Releases of All Sites

(kg/Year)

EPA Models and Assumptions

Import and Compounding of Fragrance Oils

Unloading 100 percent of HHCB 

at 75 °C from bulk containers into 

dilution tank

Air NA 5-49 2-18a 9.02 × 10-3 0.812-0.884 EPA AP-42 Model (US EPA, 1991; 

Fehrenbacher and Hummel, 1996)

Transferring 65 percent of 

(diluted) HHCB at 75 °C from 

dilution tank into totes for 

distribution to customers or 

indoor vessels for compounding

Air NA 5-49 26-251a 3.22 × 10-4-

6.45 × 10-4

0.405-0.821 EPA AP-42 Model (US EPA, 1991; 

Fehrenbacher and Hummel, 1996)

Cleaning of bulk containers Water 0.07-0.2b 5-49 2-18a 13.2-37.8c 1,190-3,400d EPA Bulk Transport Residual Model 

(US EPA, 1988; 1992)

Cleaning compounding 

equipment

Water 0.07-1b 5-49 250 0.097-13.6 1,190-17,000 EPA Single Process Vessel Residual 

Model (US EPA, 1988; 1992)

Blending of Fragrance Oils into Products

Cleaning of transport containers Water, 

incineration, 

land

0.3-3

e 950 250 0.021-0.215 5,100-51,000 EPA/Office of Pollution Prevention 

and Toxics (OPPT) Drum Residual 

Model (US EPA, 1988; 1992)

Cleaning of process vessels Water, 

incineration, 

land

2 950 250 0.143 34,000 EPA/OPPT Multiple Vessel Residual 

Model (US EPA, 1988; 1992)

Conveying, mixing, and packaging 

powder products

Air, water, 

incineration, 

land

3.83 950 250 0.274 65,110 4.5 percent dust losses from spraydrying unit with 85 percent air 

pollution control device efficiency 

(4.5 percent × 0.85 = 3.825 

percent) (OECD, 2010)

Use of End Products

Disposal of products Water 90.0-93.7 Not 250 Not known 1,530,000-1,592,900 Engineering judgment assuming a 
DRAFT DO NOT CITE OR QUOTE

Page 85

Table E-1. Summary of Estimated Environmental Releases.

Release Activity Release Media

Release Factor 

(Amount 

Released per 

Amount Used, 

Percent)

Number 

of Sites

Number of 

Release Days 

at Each Site 

(Days/Year)

Daily 

Release

(kg/SiteDay)

Combined Annual 

Releases of All Sites

(kg/Year)

EPA Models and Assumptions

known 100 percent release scenario

Total Releases Water 90.1-94.9 NA 2-250 NA 1,532,380-1,613,300

Air NA NA 2-251 NA 1.22-1.71

Water, 

incineration, 

land (uncertain)

2.3-5 NA 250 NA 39,100-85,000

Air, water, 

incineration, 

land (uncertain)

3.83 NA 250 NA 65,110

NA – not applicable. All release calculations assume an annual HHCB volume of 1,700,000 kg/year (IFRA, 2012b). 

Note: containers may be cleaned by a third-party (i.e., contractors); therefore, container cleaning releases may or may not occur on-site. a

The number of unloading days, which is equal to the number of release days, is calculated by dividing the average HHCB throughput per site by the container volume, assuming that one container is 

unloaded per site per day. Bulk containers are assumed to be 5,000 gallons and totes are assumed to be 550 gallons. The calculation for the cleaning of bulk containers is presented as a sample 

calculation of this variable:

Low-end of range of number of unloading days (or release days) per site:

HHCB throughput per site = (1,700,000 kg/year ÷ 49 sites) ÷ (1 g/cm3 × kg/1,000 g × cm3

/0.0002642 gallons) = 9,166 gallons per site.

Number of release days per site = 9,166 gallons/site-year ÷ 5,000 gallons/site-day = 1.83 days ~2 days.

UpperHigh-end of range of number of unloading days (or release days) per site:

HHCB throughput per site = (1,700,000 kg/year ÷ 5 sites) ÷ (1 g/cm3 × kg/1,000 g × cm3

/0.0002642 gallons) = 89,828 gallons per site.

Number of release days per site = 89,828 gallons/site-year ÷ 5,000 gallons/site-day = 18 days. b

The release factor, or the amount released per unit amount used, is equal for this release activity to the residual amount in a transport container or process equipment. This residual amount is 

expressed as a fraction of container or equipment volume and has the following range of values: 

(1) container unloading and equipment cleaning by gravity drain: 0.07 percent (central tendency) and 0.2 percent (high-end).

(2) equipment cleaning by pumping: 1 percent (conservative). c

The calculation of this range of releases per site is presented as a sample calculation of this variable:

Low-end of range of amount released per site-day:

5,000 gallons/container-site-day × 0.0007 residual gallons/container gallon × (1 g/cm3 × kg/1,000 g × cm3

/0.0002642 gallons) = 13.2 kg/site-day

UpperHigh-end of range of amount released per site-day =

5,000 gallons/container-site-day × 0.002 residual gallons/container gallon × (1 g/cm3 × kg/1,000 g × cm3

/0.0002642 gallons) = 37.8 kg/site-day d

The calculation of this range of combined releases of all sites per year is presented as a sample calculation of this variable:

Lower-end of range of amount released per year = 13.2 kg released /site-day × 5 sites × 18 days of release = 1,190 kg/year

Upper-end of range of amount released per year = 37.8 kg released/site-day × 5 sites × 18 days of release = 3,400 kg/year e

Residual amount as a fraction of drum volume: (1) containers unloaded by pumping: 3 percent (high-end) and 2.5 percent (central tendency); (2) containers unloaded by pouring: 0.6 percent (highend) and 0.3 percent (central tendency). 
DRAFT DO NOT CITE OR QUOTE

Page 86

Estimated Release from Compounding

The number of compounding sites in which HHCB is processed is not known. Therefore, the 

number of sites was estimated using two alternative approaches, resulting in an estimated 

range of per-site release from compounding sites. EPA expects that imported HHCB may be 

delivered to compounding sites. Compounding sites are classified under North American 

Industry Classification System (NAICS) code 32562, Toilet Preparation Manufacturing, and 

specifically under the Perfume Oil Mixtures and Blends Manufacturing subcategory. This NAICS 

code is not mentioned in the 2006 non-CBI IUR data on HHCB; therefore, EPA assumed that 

HHCB compounding may occur only at the five import sites. This approach resulted in a high 

estimate of per-site release. Alternatively, the number of sites was estimated to be equal to 

the number of IFRA member companies. The estimated HHCB volume used to compound 

fragrance oils was collected through surveys of IFRA member companies, which represent the 

vast majority of fragrance volume produced in the US. IFRA North America currently has 49 

member companies (IFRA, 2012d). If each member company owns at least one site in the US, 

then each site may compound HHCB for fragrance oils. Alternately, the 2002 Economic Census 

estimates 33 companies under this NAICS subcategory (USCB, 2004). The 2007 Economic 

Census does not provide a detailed breakdown of NAICS 32562 by subcategory, but shows that 

the total number of companies within NAICS 32562 increased by four percent from 2002 to 

2007. EPA infers that the number of companies under the Perfume Oil Mixtures and Blends 

Manufacturing subcategory also increased by four percent from 2002 to 2007, and estimates a 

total of 34 companies under this subcategory in 2007.

9 Therefore, the number of compounding 

sites is estimated to include 5 to 49 sites. The geographic distributions of all sites under NAICS 

32562 are shown in Table E-2.

In summary, compounding could occur at these sites under two scenarios: (1) after import, 

HHCB is diluted and compounded at the five import sites; or (2) HHCB is delivered to and 

compounded at the 49 compounding sites identified in the Census. During compounding, EPA

assumes that imported HHCB is unloaded into large outdoor storage tanks for dilution to a 65

percent solution and is subsequently transferred to an indoor tank in the compounding facility 

for further processing. Typically, HHCB is present in the compounded fragrance oils at two to 

four percent (HERA, 2004a).

At room temperature (25 °C), HHCB is a non-volatile liquid with a vapor pressure of 

5.47 × 10-4 mmHg (see Table 2-1). Because HHCB is a highly viscous liquid at concentrations 

≥65 percent, concentrated HHCB may be heated to become less viscous during transfer 

activities, during which HHCB may have sufficient vapor pressure to volatilize (EC, 2008). The 

unloading and transfer are expected to occur at temperatures between 25 and 75 °C (EC, 2008). 

Therefore, there is potential for releases to air due to volatilization from unloading and transfer 

of HHCB at elevated temperatures. Releases to other environmental media may also occur 

 9 Because the number of establishments under this subcategory is not known, and most companies under 

NAICS 32562 are small companies with no more than one establishment, the 34 companies are assumed to 

represent 34 establishments (sites). 
DRAFT DO NOT CITE OR QUOTE

Page 87

Table E-2. Geographic Distribution for Facilities under NAICS 32562 Toilet Preparation 

Manufacturing.

State Number of 

Establishments

Establishments with 20 

Employees or More

Number of 

Employees

California 188 59 7,572

New Jersey 97 56 10,227

New York 62 24 6,243

Texas 60 20 3,134

Florida 59 16 1,531

Illinois 41 22 3,860

Pennsylvania 30 10 1,575

Georgia 24 6 652

Connecticut 22 11 2,582

Minnesota 18 10 1,469

Missouri 17 6 1,091

North Carolina 17 9 3,984

Ohio 17 11 2,370

Colorado 14 3 283

Tennessee 14 9 2,033

Washington 14 1 106

Wisconsin 12 3 190

Arizona 10 6 —a

Massachusetts 9 4 249

Arkansas 8 4 1,545

Maryland 8 3 500

Idaho 7 3 318

Indiana 7 3 378

Michigan 7 5 2,795

Utah 7 3 —b

Virginia 7 4 1,208

Kentucky 5 2 —a

Iowa 4 2 —b

Delaware 2 1 —b

Vermont 2 1 —c

a

250 to 499 employees. b

500 to 999 employees. c

100 to 249 employees.
DRAFT DO NOT CITE OR QUOTE

Page 88

Source: USCB (2007a).

from cleaning of transport containers and compounding vessels. The environmental release 

media may include water or incineration, depending on the method of cleaning (IFRA, 2012c).

No quantitative release information was found for the compounding scenarios. Compounding 

releases were assessed using US EPA release models related to cleaning and transfer activities

(Fehrenbacher and Hummel, 1996; US EPA, 1988; 1991; 1992). In the absence of information 

to indicate otherwise, EPA assumed that container and equipment residues containing HHCB 

were released to water. EPA estimated the number of release days per year for container 

unloading and transfer activities at compounding sites, assuming that one container is 

unloaded per site per day and using a default unloading rate of approximately 200 gallons per 

minute. EPA estimated the number of release days per year for equipment cleaning by 

assuming that cleaning occurs after each batch, there is one batch per site per day, and there 

are 250 days per year operation at each site.

Estimated Release from Blending of Fragrance Oils 

After compounding, fragrance oils containing HHCB are blended with other materials to 

formulate commercial and consumer products. EPA assumed that HHCB may be present in the 

formulated products at a maximum of 0.9 percent (HERA, 2004b). The 2006 non-CBI IUR 

reported a range of 100 to 999 industrial processing and use sites specifically associated with 

HHCB. The estimate of 100 to 999 sites includes both compounding and blending sites. 

Because there are approximately 49 compounding sites, EPA estimated that there may be up to 

950 blending sites (999 sites – 49 sites = 950 sites). Blending sites are classified under NAICS 

codes 32562, 325611, and 325612. A geographic distribution of these sites, many of which are 

located in California, New York, and Texas, is provided in Tables E-3 and E-4. Note that the data 

include sites within the NAICS industry sectors, and may include sites that do not specifically 

handle HHCB. 

Table E-3. Geographic Distribution for Facilities under NAICS 325611 Soap and Other Detergent 

Manufacturing.

State Number of 

Establishments

Establishments with ≥20 

Employees

Number of 

Employees

Texas 72 11 1,111

California 69 11 851

Illinois 43 14 2,123

Ohio 39 15 3,075

New Jersey 34 16 1,787

Pennsylvania 32 8 900

Florida 31 5 —a

Missouri 31 8 1,553

Georgia 28 10 862
DRAFT DO NOT CITE OR QUOTE

Page 89

New York 28 5 356

Michigan 27 2 255

North Carolina 23 6 1,341

Wisconsin 21 3 182

Indiana 20 6 903

Tennessee 15 3 392

Louisiana 14 4 —b

Minnesota 13 1 150

Oregon 11 1 —c

Arizona 10 2 —b

Massachusetts 10 3 196

Colorado 9 1 —c

Kansas 8 4 —a

Utah 8 1 —d

Connecticut 7 3 270

Kentucky 6 2 —b

Maryland 6 2 —a

Rhode Island 6 3 —a

Mississippi 4 1 —c

Vermont 2 1 —c

Maine 1 1 —c

West Virginia 1 1 —c

Wyoming 1 1 —c

a

250 to 499 employees. b

500 to 999 employees. c

100 to 249 employees. d

1,000 to 2,499 employees.

Source: USCB (2007b).

Table E-4. Geographic Distribution for Facilities under NAICS 325612 Polish and Other Sanitation 

Good Manufacturing.

State Number of 

Establishments

Establishments with ≥20 

Employees

Number of 

Employees

California 68 21 1,492

New York 36 11 1,077

Texas 36 11 876

Ohio 34 15 1,243

Florida 31 8 459

Illinois 28 9 629
DRAFT DO NOT CITE OR QUOTE

Page 90

Pennsylvania 25 10 564

Wisconsin 25 11 4,459

Georgia 21 7 1,082

Missouri 18 7 —a

Indiana 17 6 497

Minnesota 17 1 —b

New Jersey 16 5 454

North Carolina 16 7 264

Michigan 14 4 255

Tennessee 14 4 281

Colorado 13 2 151

Washington 12 1 133

Massachusetts 11 4 209

Mississippi 9 3 —b

Oregon 9 2 195

Maryland 6 3 439

Virginia 5 2 128

Iowa 2 2 —b

West Virginia 2 1 —b

Delaware 1 1 —c

a

1,000 to 2,499 employees. b

100 to 249 employees. c

500 to 999 employees.

Source: USCB (2007c).

During the blending of fragrance oils into commercial and consumer products, releases may 

result from the cleaning of transport containers and process vessels. Additionally, releases may 

occur as a result of conveying, mixing, and packaging of powder products containing HHCB. No 

quantitative release information was found for the scenario of blending of fragrance oils; 

therefore, these releases were estimated using the OECD Emission Scenario Document (ESD) on 

the Blending of Fragrance Oils into Commercial and Consumer Products (OECD, 2010). The ESD 

assumes that fragrance oils may be transported to blending sites in drums. Further, the ESD 

indicates a potential for dust losses if the operation involves the blending of solid commercial 

and consumer products (e.g., powder detergent) (OECD, 2010). Because the volume of HHCB 

formulated into solid and powdered products is not known, EPA conservatively assumes that 

the entire volume is formulated into powder products, which results in an overestimate of the 

release amount. EPA estimated 250 release days per year (OECD, 2010), assuming that 

container cleaning occurs on a daily basis. In some cases, containers may be sent to contractors 
DRAFT DO NOT CITE OR QUOTE

Page 91

for cleaning, and actual releases may not occur onsite. EPA also assumes that one batch of 

fragrance product is blended per site per day and that equipment is cleaned after each batch. 

Estimated Release from Use of Commercial and Consumer Products

No quantitative release information was found for the use of commercial and consumer 

products containing HHCB. EPA assumed that the use of all end-use products containing HHCB 

results in down-the-drain release of HHCB to water. The basis for this assumption is the 

information on the type of end-use products that fragrance oils are used in that is given in Table 

2-7. All of the types of end-use products listed with the possible exception of “other” and “fine 

fragrances” are products whose use is likely to result in down-the-drain release of these 

products, and these products together represent 89% of the use volume of HHCB. It is noted 

that the data in Table 2-7 is neither specific to HHCB nor to the use of HHCB in the US. 

However, a 100% release to water from the use of end-use products is also assumed in other 

HHCB risk assessments (EC, 2008 and HERA, 2004b). This assumption is deemed to be 

adequate for the purpose of this release assessment as well. These releases, combined with 

upstream losses from compounding and blending, account for 100 percent of the HHCB use 

volume. Wastewater containing detergent and cleaning products is discharged to the sewer 

and routed to industrial and municipal wastewater treatment facilities. The release estimates 

represent the quantity released prior to treatment (e.g., wastewater influent). Because HHCB 

has uses in many different types of products and an accurate estimate of the number of 

commercial sites that may use these products is not available, release calculations for use of 

commercial and consumer products are presented as combined releases from all sites on an 

annual basis. 
DRAFT DO NOT CITE OR QUOTE

Page 92

Appendix F: Environmental Fate and Bioaccumulation Study 

Summaries

ENVIRONMENTAL PERSISTENCE

Fate in Wastewater Treatment 

In the Langworthy et al., 2000 study (as reported in EC, 2008), HHCB disappeared with the 

subsequent appearance of more polar metabolites. The first metabolite appeared between 

day 1 and 8 and had an elution time by thin-layer chromatography (TLC) similar to that of 

HHCB-lactone and accounted for about 40 percent of the original radioactivity. Gradually a 

second metabolite increased and accounted for up to 45 percent of the radioactivity after 27 

days. This metabolite had a similar elution time as for HHCB hydroxycarboxylic acid. A third, 

highly polar metabolite constituted up to 15 percent of the radioactivity after 150 hours (6.25 

days). The mass balance accounted for 99 percent of the radioactivity, and no significant 

volatilization was detected.

Fate in Water

In batch experiments (Langworthy et al., 2000 and Federle et al., 2002 as reported by EC, 2008), 

more polar metabolites made up 60 percent of the original radioactivity after 29 days. The 

mass balance accounted for 92 percent of the radioactivity. In a series of die-away tests, a

variety of metabolites were formed at different rates and with increasing polarity over time 

(Schaefer et al., 2005; as reported in EC, 2008). In the river die-away test the total radioactivity 

of metabolites rose from 25 to 56 percent between days 2 and 7 and to nearly 62 percent at 28 

days. When the results were corrected for the amount volatilized and the non-recovered 

fraction as well as for abiotic transformation, the primary biodegradation extent on day 28 was 

approximately 60 percent. About 40 percent of the radioactivity was lost by volatilization at 

300 hours, but some of the volatilized radioactivity could have been in the form of degradation 

products, since degradation in water was also happening during the test. 

Fate in Soil and Sediment

In Envirogen, 1998 (as reported in EC, 2008), sealed flasks with HHCB-spiked soil at 10 μg/g soil 

(10 mg/kg) were incubated at laboratory ambient temperature for one year. Periodically the 

headspace was flushed for oxygen replenishment and the effluent gas was drawn through a 

train of scintillation fluids to capture volatile organics and carbon dioxide. After incubation, the 

flasks were sacrificed and exhaustively extracted with ethyl acetate and/or acetone/hexane. An 

aliquot of the solvent was used for TLC. After one year, significant amounts of polar 

metabolites were found, and HHCB was present at 4, 7, 9 and 35 percent of the initial 

concentration in the river sediment, forest soil, sludge-amended soil, and agricultural soil, 

respectively. The TLC results showed that HHCB was degraded to various more polar 

metabolites. It was hypothesized that the majority of unrecovered radiolabel became 

covalently bound to soil organic compounds (i.e., was immobilized by humification).
DRAFT DO NOT CITE OR QUOTE

Page 93

As reported by DiFrancesco et al., 2004 (dissipation of fragrance materials in sludge-amended 

soils), anaerobically digested and dewatered sludge was obtained from activated sludge 

WWTPs in Georgetown, Delaware (100 percent domestic sewage; 10 percent solids) and 

Wilmington, Delaware (70 percent domestic sewage; 17 percent solids). The unamended 

concentration of HHCB in the digested sludge of Georgetown, Delaware was 86 and 38 mg/kg 

dw in the years 2000 and 2002, respectively. For Wilmington the concentrations were 43 and 

22 mg/kg dw, respectively. The test was carried out using trays containing 24 liters of soil 

mixed with 1 liter of sludge, simulating sludge applications of 7,000 wet gallons per acre and a 

15-cm plow depth or 0.6 to 1.1 kg /m2 of sludge. Each of the four soils was incubated with 

digested sludge, both unspiked and spiked with a mixture of fragrance ingredients. The soil and 

sludge were mixed over 30 minutes in a cement mixer to ensure uniform mixing. Four 

combinations were duplicated, so a total of 20 trays were set up outdoors. Leachate samples 

were collected at the end of each rain event. Soil samples were extracted by accelerated 

solvent extraction, and the leachate was extracted using C-18 speed disks. All extracts were 

analyzed by gas chromatography/mass spectrometry (GC/MS). The leachate was collected over 

3 to 5 months, and the HHCB found amounted to 0.03 to 0.18 percent of the initial amount in 

the spiked soils. HHCB was not detected in leachate from unspiked soils (EC, 2008). The low 

percentage of initial HHCB present in leachate was probably a result of both degradation and 

sorption.

BIOACCUMULATION AND BIOCONCENTRATION

Dietrich and Hitzfield (2004) summarized information on the bioconcentration and 

bioaccumulation potential of HHCB. BCFs and BAFs varied with species but were generally 

lower than the commonly used criterion of 1,000 (log = 3) for moderate bioaccumulation 

potential (EPA, 1999a; 1999b; Boethling et al., 2009). Measured BCFs of 620, 862 and 1,584 

were reported for zebra fish, eel and bluegills, respectively (Table 2-10), whereas BAF values of 

290, 510, 580, and 620 were reported for eel, tench, crucian carp and zebra mussels, 

respectively. At the same time, these and other (chiefly monitoring) studies suggest that HHCB 

is not subject to biomagnification, which occurs when a substance achieves higher 

concentrations on a lipid-adjusted basis in organisms than in their food (prey). Metabolism 

may account for the observation that the relatively low BCFs and BAFs appear inconsistent 

with the high measured log KOW of HHCB (5.9).

The bluegill study was performed using Good Laboratory Practice (GLP) standards and 

according to OECD Test Guideline 305 (EC, 2008; Rimkus, 1999). Bluegill sunfish (Lepomis 

macrochirus) were exposed in a flow-through system to two concentrations of radiolabeled 

HHCB. A solubilizer was used to prepare a solution (concentration 0.001 percent (w/v)). 

Identification of the parent compound in water and fish was performed using TLC/HPLC. 

Nominal exposure concentrations were 1 and 10 μg/L. The fish were exposed for 28 days; the 

elimination period was also 28 days. The BCF was derived from actual concentrations of parent 

compound in exposure water and the plateau level in fish. The elimination rate constant (k2) 

was estimated from the elimination curve (first-order kinetics), and determination of an uptake 

rate constant (k1) was attempted using the following formula: k1 = BCF × k2. However, this 
DRAFT DO NOT CITE OR QUOTE

Page 94

value could not be directly calculated from the increase of concentrations in fish due to rapid 

attainment of the final plateau level. 

The concentration of HHCB in the fish reached plateau levels after 3-7 days of exposure. 

Elimination followed first-order kinetics with a half-life of 2 to 3 days for HHCB, allowing 

calculation of k2. Based on the concentration of parent material, the BCF for whole fish was 

calculated to be 1,584 L/kg wet wt.

In the bluegill study, HHCB was metabolized to one or more polar metabolites in fish in a 

relatively short time and then excreted to the water phase. Even if HHCB is metabolized 

rapidly, bioaccumulation can still be significant. In BCF tests, metabolism can occur during 

exposure of fish to the test substance; therefore the measured BCF reflects a balance of uptake 

and elimination. If for example a value of 1,584 L/kg wet wt is observed for fish, it means that 

metabolism is not occurring fast enough to prevent the substance from reaching a 

concentration in the fish that is 1,584 times greater than in the exposure medium. 

Bioaccumulation has been studied with two benthic species (Table 2-11). The aqueous 

concentrations in the test with non-biting midges (Chironomus riparius) were stable at 9.8 ± 1.4 

μg/L (Artola-Garciano et al., 2003). The concentrations in the organisms increased to a 

maximum level between days 1 and 3, and then the level rapidly decreased to a new steady 

state with BCF values ranging from 85 to 138 L/kg wet wt. With the addition of piperonyl 

butoxide (PBO), a cytochrome P450 inhibitor, the BCF was 525, and it was concluded that the 

lower BCF values were likely caused by relatively fast biotransformation of HHCB in this 

organism. In a flow-through experiment with blackworms (Lumbriculus variegatus), the 

concentration in water was 4.6 ± 0.6 μg/L. The uptake of HHCB in these worms reached a 

plateau after 3 days, and the measured BCF was 2,692 L/kg wet wt. This level of 

bioaccumulation suggests that in this organism biotransformation and elimination does not 

take place at an appreciable rate.

Jager (1998; as cited in EC, 2008) the equation used to calculate BCF for worms was given as

BCFworm = (0.84 + 0.012 × Kow)/RHOearthworm

For RHOearthworm (bulk density of worm) by default a value of 1 (kg wet weight/L) may be 

assumed. This leads to a predicted BCFworm of 2,395 L/kg.
DRAFT DO NOT CITE OR QUOTE

Page 95

Appendix G: Environmental Monitoring Data Analysis

Measured Concentrations in Wastewater

The majority of synthetic musk fragrances are used in consumer products that enter WWTPs 

through down-the-drain disposal. Therefore, the concentration of HHCB in influent is 

dependent on the source of waste received (municipal, commercial, industrial) and the 

population served by the WWTP. The concentrations of fragrance materials in effluent, 

however, are dependent upon the ability of these compounds to be eliminated during 

treatment as well as WWTP size, type, and waste treatment process. 

In the draft report of the Science Advisory Panel for Chemicals of Emerging Concern in 

California’s Aquatic Ecosystems, the maximum concentration of HHCB in wastewater effluent 

reported from within the state of California was 2,780 ng/L (Anderson et al., 2012). The 

maximum aqueous concentration of HHCB in treated municipal wastewater effluent discharged 

to the ocean was 2,500 ng/L.

Quarterly samples of wastewater influent and effluent were taken at two WWTPs in Texas, and 

synthetic musk fragrance concentration was measured (Chase et al., 2012). HHCB and AHTN 

were consistently the most abundant synthetic musk fragrances present throughout quarterly 

sampling. These results, in part, reflect the high production volume of HHCB (US EPA, 2012d). 

The influent concentration of HHCB was as high as 5,735 ng/L and the effluent concentration 

was as high as 6,136 ng/L as shown in Table G-1.

A state-wide survey in Oregon of trace metals and organic chemicals in municipal effluent was 

published in 2012 (Hope et al., 2012). Oregon's Senate Bill 737, enacted in 2007, required the 

state's 52 largest municipal WWTPs and water pollution control facilities to collect effluent 

samples in 2010 and analyze them for persistent organic pollutants. These facilities are located 

state-wide and represent a variety of treatment types, service population sizes, geographic 

areas, and flow conditions. HHCB was detected in 2/102 (the reported limit of quantitation was 

10,000 ng/L) samples with a median value of 12,500 ng/L and a maximum value of 13,500 ng/L. 

Wastewater influent and effluent concentrations were determined at two WWTPs, one located 

in a rural area in Kentucky and the other from an urban site in Georgia (Horii et al., 2007). 

These two WWTPs treat primarily domestic and commercial wastewater. Concentrations of 

HHCB in influent from both WWTPs were 3 to 60 times higher than those in effluent. Mean 

concentrations of HHCB in influent of the Kentucky plant (2,499 ng/L) were up to six times 

higher than the mean concentrations for the Georgia plant (420 ng/L). However, mean 

concentrations of HHCB in the effluents did not differ significantly between the two sites 

(44 and 55 ng/L, respectively).

Wastewater and sludge samples were collected over a five-day period in October 2005 from 

two WWTPs in the state of New York that employ identical treatment processes and serve cities 
DRAFT DO NOT CITE OR QUOTE

Page 96

of moderate population size (approximately 100,000 people) (Reiner et al., 2007a). Both 

WWTPs receive domestic and commercial discharge, and although one also receives industrial 

discharge, the measured concentration of HHCB did not differ. HHCB was found in wastewater 

samples with influent and effluent concentrations in the ranges of 4,760 to 12,700 and 10 to 98 

ng/L, respectively. 

WWTP effluent was sampled at two sites over a two-year period: Taylors Falls WWTP at Taylors 

Falls, Minnesota and St. Croix Falls WWTP at St. Croix Falls, Wisconsin (USGS, 2011). The sites 

were chosen because the WWTPs discharge into the St. Croix River upstream from endangered 

mussel populations. Mean concentrations of HHCB over the sampling period were 48 ng/L at 

Taylors Falls and 1,330 ng/L at St. Croix Falls.

Osemwengie and Gerstenberger (2004) analyzed surface water from the confluence of three 

municipal sewage treatment effluent streams in the state of Nevada, for a period of 7 to 

12 months. HHCB and AHTN were consistently detected in higher monthly average 

concentrations (32.6 to 97.9 ng/L) relative to the other polycyclic musks. The sewage effluents 

were sampled from a dedicated effluent receiving stream (one that receives only sewage 

effluent and rarely runoff). 

Smyth et al. (2008) measured the HHCB concentration in the influent and effluent over a oneyear period at six different WWTPs, located in Ontario Canada, employing four different 

treatment processes. Influent concentrations were as high as 40,300 ng/L and effluent 

concentrations were as high as 3,730 ng/L. Lagoon treatment produced the lowest effluent 

concentration, although the authors noted that process temperature may have influenced the 

removal efficiency. Three WWTPs that utilize conventional activated sludge processes were 

studied, with only small differences in effluent concentrations observed. Overall, effluent HHCB 

concentrations were an order of magnitude less than the influent HHCB concentrations, as 

summarized in Table G-1.
DRAFT DO NOT CITE OR QUOTE

Page 97

Table G-1. Measured Concentrations of HHCB in Wastewater.

Location Year n

Concentration 

(ng/L)

Influent

Concentration (ng/L)

Effluent Comments Reference

CA

2007

—a

—b

2,780 maximum Effluent dominated 

inland waterway

Anderson et al.

(2012)

ND to 2,500 

maximum

Treated municipal 

wastewater 

discharged to ocean

TX 2009-

2010

—a 4,772-5,735 2,928-6,136 Effluent from two 

WWTPs

Chase et al.

(2012)

OR 2010 102, 2 

detects

—b 12,500 median; 

13,000 maximum

(2 detects from 102 

samples)

52 municipal WWTPs 

sampled

Hope et al.

(2012)

KY 2005 —a 43-7,032, 

2,499 mean

17-24; 21 mean, 

before chlorination

10-37; 44 mean

Horii et al.

(2007)

GA 284-522, 420 

mean

51-418; 189 mean, 

primary

28-98; 55 mean

NY(a) 2005 —a 4,760-12,700

9,030 mean

4,120-8,150; 3,730 

mean, primary

2,810-3,730; 3,350 

mean

Reiner et al.

(2007a)

NY(b) 2005 1,80-11,500

5,940 mean

2,950-6,820; 5,310 

mean, primary

2,360-3,310; 2,700 

mean

NV 2000-

2001

24 —b 32.6-97.9; 56.9 mean Osemwengie 

and 

Gerstenberger 

(2004)

MN

WI

2007 7 —b 48 mean

1,330 mean

Taylors Falls WWTP

St. Croix Falls WWTP

USGS (2011)

Ontario, 

Canada

2003-

2004

3 7,750-14,600

8,380-14,000

1,810-7,210

5,030-40,300

6,640-10,800

3,780-19,000

54.6-673

2,477-3,730

497-1,130

2,290-3,240

2,230-3,340

914-1,680

Lagoon

Extended aeration

Oxidation ditch

Activated sludge

Activated sludge

Activated sludge

Smyth et al.

(2008)

a

Not reported.

b

Not measured.
DRAFT DO NOT CITE OR QUOTE

Page 98

Measured Concentrations in Surface Water

Because synthetic musk fragrances, such as HHCB, are used in many consumer products, these 

compounds enter WWTPs through down-the-drain practices. The ability of fragrance materials 

to be eliminated during treatment depends on a variety of factors inherent to the WWTP, such 

as WWTP size and type (municipal or industrial), processes utilized to treat waste, and 

populations served (rural versus urban). Consequently, effluents from WWTPs are a primary 

source of fragrance materials into the environment. These compounds are found in surface 

water and groundwater located near wastewater discharge areas, with peak environmental 

concentrations typically occurring near effluent discharge points (Heberer et al., 1999; Peck et 

al., 2006). Although the concentration of HHCB in wastewater and surface water may vary, 

HHCB is continuously present and constitutes a constant exposure source.

A science advisory panel convened by the Southern California Coastal Water Research Project 

was charged to identify potential sources of chemicals of emerging concern, evaluate their fate 

and effects, and ultimately provide guidance for developing monitoring programs that assess 

those chemicals with the highest potential to cause effects in the state’s receiving waters 

(Anderson et al., 2012). Three scenarios were evaluated: Scenario 1, a WWTP effluentdominated inland (freshwater) waterway; Scenario 2, a coastal embayment that receives both 

WWTP effluent and stormwater discharge; and Scenario 3, offshore ocean discharge of WWTP 

effluent. The maximum values reported for each of these scenarios are recorded in Tables G-1 

and G-2. 

Water samples were taken from the North Side water reclamation plant effluent and the North 

Shore Channel of the North Branch of the Chicago River, both located in the Chicago, Illinois 

metropolitan area (Barber et al., 2011). This plant collects sewage from residential (96 percent) 

and commercial/industrial (3.4 percent) sources. Following treatment, the water reclamation 

plant effluent is discharged to the North Shore Channel. There were no statistical differences in 

concentrations of HHCB in these two locations, indicating minimal in-stream dilution or 

transformation. The mean concentration was reported as 1.6 μg/L (1,600 ng/L). 

Passive sampling devices were deployed within two nearshore study zones along the north 

shore of Lake Ontario, Canada near the municipalities of Pickering and Ajax, Ontario, and Port 

Hope, Ontario (Helm et al., 2012). Time-weighted average dissolved water concentrations were 

estimated based on chemical amounts sequestered in the device. Organic wastewater 

chemicals, like HHCB, were greatest in the vicinity of the Pickering-Ajax WWTP, and then 

dramatically declined. Calculated sampling device-based concentrations of HHCB ranged from 

approximately 0.2 to 10 ng/L near Pickering-Ajax and from 0.1 to 0.3 ng/L near Port Hope.

HHCB was detected in the surface waters of the Canyon Lakes System and North Fork Brazos 

River near Lubbock, Texas in 2010 (Chase et al., 2012). Reported concentrations in the lake 

ranged from ND to trace amounts, whereas concentrations in the river were appreciably higher,

ranging from 77 to 794 ng/L. The lake concentrations were expected to be low as the 
DRAFT DO NOT CITE OR QUOTE

Page 99

groundwater from a land applied site where the source was from a lake, whereas occurrence in 

the stream water was attributed to direct release of WWTP effluent.

Peck et al. (2006) characterized the concentration of several synthetic musk fragrances in 

western Lake Michigan with water samples that were collected from 1999 to 2001. A lake-wide 

annual mass balance analysis showed that WWTP discharge was the major source of the 

synthetic musks, whereas major loss mechanisms were outflow and volatilization. The average 

measured concentration of HHCB over this time period was 4.7 ng/L. 

Water concentrations were measured in 2006 for HHCB and AHTN from three locations along 

the upper Hudson River, with mean concentrations for HHCB ranging from 3.95 to 25.8 ng/L

(Reiner and Kannan, 2011). The highest concentration was found near Albany, New York, which

corresponded to the location with the highest population.

The USGS and the National Park Service cooperated on a study to determine the occurrence of 

wastewater indicators in the St. Croix National Scenic Riverway in Minnesota and Wisconsin

(USGS, 2008). Samples of treated wastewater effluent from two WWTPs, located in Taylors

Falls, Minnesota and St. Croix Falls, Wisconsin were collected from 2007 to 2008, with 

concentrations ranging from ND to 1,330 ng/L.

Concentrations of HHCB were detected in low concentrations in Lake Mead, with a mean value 

of 0.36 ± 0.3 ng/L in 2001, although this waterbody is downstream of the confluence of three 

WWTPs in Las Vegas, Nevada (Osemwengie and Gerstenberger, 2004). The low concentrations 

of HHCB, compared to the concentrations found in the municipal sewage effluent, may be due 

in part to the dilution of the discharge of the combined sewage treatment plant effluents, as 

the effluent constitutes only 1.37 percent of the total lake volume.

HHCB concentration was measured in a limited set of samples from Hamilton Harbor and the 

open waters of Lake Ontario (Andresen et al., 2007). The measured HHCB concentration within 

the bay of Hamilton (highly influenced by WWTP effluents) was 7.0 ng/L. With increasing 

distance from Hamilton Harbor, the measured concentration decreased and stabilized at 

2.0 ng/L. The authors indicated that the decrease of the measured concentrations might be 

traced back to dilution effects or that musk fragrances are stable under the conditions found in 

Lake Ontario; however, they concluded that further investigation is needed.
DRAFT DO NOT CITE OR QUOTE

Page 100

Table G-2. Measured Concentrations of HHCB in Surface Water.

Location Date n

Concentration 

(ng/L) Comments Reference

CA 2008 970 maximum Resevoir Anderson et 

al. (2012)

Chicago, IL 2006-2007 3 1,600 mean North Shore Channel of the Chicago 

River (Impacted by water reclamation 

plant effluent)

Barber et al.

(2011)

Lubbock, TX Fall 2010

Spring 

2010

—a 45

77-794

Canyon Lakes System

North Fork Brazos River

Chase et al.

(2012)

Michigan 1999-2001 13 4.7 ± 2.5 Lake Michigan, background locations Peck and 

Hornbuckle 

(2004)

Catskill, NY

Troy, NY

Albany, NY

7/2006

5/2006

5/2006

2

2

2

3.95 mean

7.36 mean

25.1 mean

Upper Hudson River Reiner and 

Kannan 

(2011)

Taylors Falls, MN

St. Croix Falls, WI

Franconia, MN

Sunrise, MN

2007-2008 8 48

1,330

ND

ND

Taylors Falls WWTP at Taylors Falls, MN

St. Croix Falls WWTP at St. Croix Falls, WI

St. Croix River above Rock Island

St. Croix River below Sunrise River

USGS (2011)

Las Vegas, NV 2000-2001 14 0.06-1.02, 0.36 

mean

Lake Mead Osemwengie 

and 

Gerstenberg

er (2004)

Great Lakes 50 0.08-180; 0.8 

median

8.44 mean;

27.285 95th

percentile

Great lakes watershed, summary Klecka et al.

(2010)

Lake Ontario, 

Canada

June 2008 —a 0.12-0.34 

estimated

0.21-10.1 

estimated

Port Hope

Pickering-Ajax

Helm et al.

(2012)

Lake Ontario,

Canada

2005 4 7 Hamilton Harbour Andresen et 

al. (2007) a

Not reported.

Measured Concentrations in Sediment

Studies of measured environmental concentrations of fragrance compounds in the sediment in 

North America are limited. Synthetic musk fragrances are known to be present in wastewater 

effluent and the aquatic environment and have been detected in downstream surface water 

sediment of streams and lakes. Aquatic biota may be subsequently exposed to HHCB from 

suspended sediment intake.
DRAFT DO NOT CITE OR QUOTE

Page 101

Chase et al. (2012) measured the concentration of HHCB in the sediment of a playa lake, a noneffluent impacted study site located in proximity to an area treated by a WWTP. The primary 

water sources are rain and urban runoff, as these playas are integral for storm water retention. 

In this study, sampling was performed in the summer and fall timeframes of 2010. The 

measured concentrations of HHCB were 2.13 and 1.43 ng/g dw in the summer and fall of 2010, 

respectively, as shown in Table G-3.

The USGS investigated the occurrence and distribution of organic waste contaminants in the 

Tinkers Creek watershed and two other tributaries to the Cuyahoga River in northeastern Ohio

(USGS, 2008). The effluent from WWTPs constitutes a continuous and sometimes high

proportion of the flow in Tinkers Creek and its tributaries (sometimes ≥80 percent). HHCB was 

detected in the streambed sediment in 29 percent of upstream sites and 86 percent of 

downstream sites. The range of values is 20 to 390 μg/kg dw with an estimated median value 

of 60 μg/kg dw. The reported upstream concentrations were <60 μg/kg dw, whereas the 

downstream concentrations ranged from <30 to 390 μg/kg dw with three of the seven 

downstream samples having measured values >200 μg/kg dw.

Water and sediment samples were collected in May and July 2006 at three locations along the 

upper Hudson River: Troy, Albany, and Catskill, New York (Reiner and Kannan, 2011). WWTPs 

in each of these New York cities discharge effluents into the Hudson River. HHCB and AHTN 

were detected in all sediments sampled. Concentrations of HHCB were 72.8 ng/g dw from 

Catskill, 351 ng/g dw from Albany, and 388 ng/g dw from Troy. The authors noted that these 

sediment concentrations were similar to concentrations that have been measured in Germany 

(Fromme et al., 2001; Dsikowitzky et al., 2002). 

Sediment samples from three tidal tributaries of the Chesapeake Bay, US (Magothy, Corsica,

and Rhode Rivers) were collected in 2007 (Sapozhnikova et al., 2010). The Rhode River is the 

only one that receives effluent from a WWTP. Concentrations of HHCB detected in sediments 

were as high as 9.2 ng/g dw for the Magothy River, 2.3 ng/g dw for the Rhode River, and 

1.6 ng/g dw for the Corsica River. The reported average HHCB concentration was 1.1±2.2 ng/g 

dw. The Magothy River is the most urbanized river among the three tributaries studied and 

was reflected in the concentration of HHCB found in the sampling. Thirty-nine total samples 

were collected, but details regarding the number of samples for each location were not 

provided.

Sediment samples from Lake Erie and Lake Ontario were measured in August of 2003 (Peck et 

al., 2006). The HHCB concentration measured in Lake Ontario (16 ng/g dw) was an order of 

magnitude higher than that of Lake Erie (3.2 ng/g dw). This difference may be attributed to 

multiple factors, including distance from principal sources (e.g., wastewater outfalls) and 

differences in loss mechanisms (e.g., photolysis, outflow, or volatilization). The concentration 

of HHCB over time was extrapolated from the collected sediment core samples and the HHCB 

concentration was correlated to US fragrance consumption from 1980 to 2006. The authors 
DRAFT DO NOT CITE OR QUOTE

Page 102

concluded that HHCB concentration found in the sediment samples reflected the trend of 

fragrance consumption over this time period. 

Table G-3. Measured Sediment Concentrations at Locations in the US.

Location Year Concentration

ng/g dw Comments Reference

Lubbock, TX Summer 

2010

Winter 

2010

2.13

1.43

Playa Chase et al. (2012)

Tinkers Creek, OH

Cuyahoga River, OH

2006 20a-<60, 37 average

20a

-390, 144 mean

Upstream of 7 WWTPs

Downstream of 7 WWTPs

USGS (2011); 

Klecka et al. (2010)

Troy, NY

Albany, NY

Catskill, NY

2006 388

351

72.8

Upper Hudson River, NY Reiner and Kannan 

(2011)

Magothy River, MD

Rhode River, MD

Corsica River, MD

2007 ND - 9.2

ND - 2.3

ND - 1.6

1.1 avg

Tributary, Chesapeake Bay Sapozhnikova et al.

(2010)

Lake Erie

Lake Ontario

2003 3.2

16

Great Lakes Peck et al. (2006)

a

Reported as an estimated value.

ND = not detected

Measured Concentrations in Biosolids

The removal of fragrance compounds from wastewater has been reported in the US, Canada,

and EU in several studies, with overall removal efficiencies ranging from as low as 29.9 percent 

to as high as 99.7 percent (typical values were 70 to 80 percent) depending upon the type of 

wastewater treatment (Simonich et al., 2002; Deblonde et al., 2011; Smyth et al., 2008). A 

fraction of HHCB is expected to preferentially partition to organic-rich biosolids during 

wastewater treatment based on the high measured log KOW values of 5.3 to 5.9 (see Table 2-1) 

and the estimated KOC values of 4.1 to 4.3 (EPI Suite v4.10, see Chapter 2, Section B,

Environmental Fate). These solids, produced through the process of wastewater treatment, are 

subsequently disposed of by either landfill or incineration. They may also be utilized for land 

application to fertilize plants or to improve the quality of soil, and may then become available 

for plant and terrestrial biota uptake.

In a study by Kinney et al. (2006), the presence and concentration of organic waste 

contaminants was determined in solid materials produced during wastewater treatment. Nine 

different biosolid products from municipal WWTPs in seven different states were analyzed. Six 

of these products were intended for commercial, homeowner, and municipal use and are 

available to the public for purchase. Three of the products are used in agriculture. HHCB was 

detected in all nine biosolids. HHCB was extracted from the biosolid using accelerated solvent 

extraction and measured by GC/MS. Each sample was composited from bulk material and 

samples were analyzed in triplicate. All samples had similar treatment (secondary activated 
DRAFT DO NOT CITE OR QUOTE

Page 103

sludge and chlorine disinfection); however, the preparation and applied form of the solid 

varied. The concentration of HHCB varied, with one site sample measuring as high as 

177,000 μg/kg dw. All other measured HHCB concentrations were below 4,000 μg/kg dw. As 

this high value was more than 40 times higher than the next highest measured concentration, 

and more than four orders of magnitude higher than the lowest reported concentration, the 

authors considered it as an “outlier.” Excluding this value, the reported median value for HHCB 

was 1,461 μg/kg dw. Differences in biosolid preparation techniques and WWTP characteristics 

(e.g., location, population served, volumes of wastewater, and retention times) likely account 

for the wide variation in HHCB concentrations measured, with no one factor being clearly 

delineated in this study.

A second study by Kinney et al. (2008) was performed to examine bioaccumulation in 

earthworms from soil obtained from three Midwest agricultural fields, which had been 

amended with biosolids or swine manure. The biosolid source had a concentration of 

427,000 μg/kg dw, although details regarding this source were not provided.

Biosolid samples were collected from the North City Wastewater Reclamation Plant of the City 

of San Diego to study biosolid degradation on a weekly schedule for 19 weeks (Buyuksonmez 

and Sekeroglu, 2005). Samples were taken from the return activated sludge line after purging 

the stagnant sludge from the sampling line. The presence of HHCB was determined using 

GC/MS following extraction with a Soxhlet apparatus for 24 hours. The average concentration 

of HHCB in biosolids was reported to be 7.84 mg/kg dw (7,840 μg/kg dw), although the number 

of samples taken was not provided.

In a study by DiFrancesco et al. (2004), digested sludge was collected from two sites over a 

period of two years. Anaerobically digested and dewatered sludge was obtained from two 

activated sludge plants in Delaware: Georgetown WWTP and Wilmington WWTP. Background 

concentrations of fragrance materials in anaerobically digested sludge from the municipal 

WWTPs were determined in July 2000 and again in March (Georgetown) and October 

(Wilmington) 2002. Sludge was dewatered using a gravity separator at Georgetown WWTP and 

a belt filter press at Wilmington WWTP. Both WWTPs added polymers to assist in sludge 

dewatering, and the Georgetown facility also used aluminum chloride for phosphorus removal. 

Samples were extracted using accelerated solvent extraction and were analyzed using GC/MS. 

The HHCB concentrations ranged from 21,800 to 86,000 μg/kg dw. The overall concentrations 

in this limited sampling were higher in samples collected in 2000 and at the Georgetown site. 

Table G-4 includes a summary of measured concentrations of HHCB in biosolids.

For comparison, studies of biosolids from Canada are also included in Table G-4. HHCB 

concentrations in digested biosolid were studied by Smyth et al. (2007) in Canada. One 

oxidation ditch with tertiary sand filtration, one extended aeration activated sludge plant, and 

three conventional activated sludge plants that discharge into the Grand River, Ontario, Canada 

were used as sampling sites with four sampling events from November 2003 to July 2004. 

Extraction was performed using supercritical fluid or microwave assisted techniques, 

measurement was performed using GC/MS. HHCB concentrations in digested biosolid ranged 
DRAFT DO NOT CITE OR QUOTE

Page 104

from 9,430 to 55,500 μg/kg dw. Smyth et al. (2007) reported that these values were in range 

with previous reported values in Canada (Lee et al., 2003) and the United Kingdom (Stevens et 

al., 2003).

Yang and Metcalfe (2006) also studied biosolids from WWTPs in Ontario, Canada. Samples 

were collected from the Peterborough WWTP over a one-year period, and the HHCB 

concentrations were measured in the raw, return activated sludge, and digested biosolids. 

Concentrations of HHCB did not vary significantly for any one sampling type over the one-year 

study. However, there was a trend for the synthetic musks to accumulate in the return 

activated sludge and to finally concentrate in the digested biosolids. The annual average 

concentrations of HHCB were 3,309.9 μg/kg dw in the return activated sludge and 

6,788.4 μg/kg dw in the digested biosolid. 

Table G-4. Measured Concentrations of HHCB in Biosolids.

Location Collection

Year n

Concentration

μg/kg dw Preparation/Comments Reference

WI 5/2003

11/2003

3

3

1,400 (average)

1,470 (average)

Rotary kiln dried Kinney et al. (2006)

CO 5/2003

11/2003

3

3

3,700 (average)

2,670 (average)

Composted + plant material

TX 8/2003

11/2003

3

3

1,100 (average)

977 (average)

Composted + yard material

WA 11/2003 3 933 (average) Composted with sawdust

WA 5/2004 3 13 (average) Composted with yard waste

KS 9/2003

11/2003

3

3

2,370 (average)

1,970 (average)

Air dried and turned

AZ 6/2003

1/2004

3

3

1,300 (average)

697 (average)

biosolid

AZ 6/2003

1/2004

3

3

1,100 (average)

767 (average)

Dewatered biosolid + polymer

IA 4/2005 3 177,000 (average) Thermophilic and mesophillic 

digestion + polymer

Unknown 2008 3 427,000b No details given Kinney et al. (2008)

San Diego, CA 2005 —a 7,840 (average) Return activated sludge line Buyuksonmez and 

Sekeroglu (2005)

Wilmington, 

DE

7/2000

10/2002

—a

7

43,000b

21,800 (± 4,300)

Anaerobically digested sludge DiFrancesco et al.

(2004)

Georgetown, 

DE

7/2000

3/2002

—a

3

86,000b

37,600 (± 4,500)

Anaerobically digested sludge
DRAFT DO NOT CITE OR QUOTE

Page 105

Table G-4. Measured Concentrations of HHCB in Biosolids.

Location Collection

Year n

Concentration

μg/kg dw Preparation/Comments Reference

Ontario, 

Canada

(5 locations)

2003-

2004

7

10

12

11

11

9,430 (mean)

40,300

42,000

55,500

46,300

Aerobically digested sludge

Aerobically digested sludge

Anaerobically digested sludge

Anaerobically digested sludge

Anaerobically digested sludge

Smyth et al. (2008)

Canada (site

1/site 2)

2003-

2005

—a 36,500/20,100

25,600/17,500

73,700/45,300

23,800/20,500

Primary sludge, wet

Primary sludge, air dried

Waste activated sludge, wet

Waste activated sludge, air 

dried

Smyth et al. (2007)

Canada

Peterborough, 

ON

2/2003

4/2003

7/2003

10/2003

Annual 

average

1

3

3

3

3

3

3

3

3

3

3

3

4,514.4

3,638.8 (± 787.7)

7,896.7 (± 2,130.8)

2,857.4 (± 18.6)

5,103.9 (± 635.0)

7,006.8 (± 931.3)

3,356.1 (± 1,058.6)

2,156.6 (± 504.7)

6,477.6 (± 1,065.3)

2,482.4 (± 340.6)

2,340.4 (± 360.2)

5,772.76 (± 115.2)

3,302.5

3,309.9

6,788.4

469.6 (± 22.4)

Raw sludge

Return activated sludge

Digested sludge

Raw sludge

Return activated sludge

Digested sludge

Raw sludge

Return activated sludge

Digested sludge

Raw sludge

Return activated sludge

Digested sludge

Raw sludge

Return activated sludge

Digested sludge

Trucked biosolid

Yang and Metcalfe 

(2006)

a

Not reported. b

Value was not reported as mean, average, or median.

Studies summarized in Table G-4 reported analytical reliability using accelerated solvent 

extraction or supercritical fluid extraction followed by GC/MS techniques. Studies of HHCB in 

biosolids are somewhat limited, however, because the studies that detected and measured 

HHCB reported concentrations in the range of <100 to >100,000 μg/kg dw. The measured 

concentration depended on a variety of factors, including WWTP operations such as receiving 

waste stream (municipal or industrial), date of collection, location, population served, water 

volume, treatment type, and preparation methodologies prior to land application. Because 

these factors cannot be independently separated from values within the reported data set, it is 

not possible to assign sensitivity to a single factor. Comparison with values reported for studies 

in Ontario, Canada are within the reported range for those from the US and also similarly show 

differences depending upon location, treatment process, and season. 
DRAFT DO NOT CITE OR QUOTE

Page 106

Measured Concentrations in Soil

Limited data were available that described concentrations of HHCB in soil. Though many 

studies are designed toward understanding the biodegradation or fate of HHCB in the soil 

compartment, there are data showing that land application of wastewater effluent or 

amendment of soils with biosolids resulted in measurable concentrations of HHCB. The high 

frequency of HHCB detection in biosolids and wastewater suggests that they may be an 

important organic waste contaminant source to terrestrial environments, particularly when 

applied to soil.

In a study by Chase et al. (2012), soil core samples were collected quarterly over a two-year 

period at sites outside the city limits of Lubbock, Texas to determine the amount of sorption 

after land application of treated wastewater and groundwater recharge. The wastewater at 

this site is treated through primary and secondary treatment processes for removal of 

suspended solids and organic carbon through different operating flow streams at the WWTP. 

Some of the treated effluent is released to a secondary effluent reservoir, which is then applied 

via pivot irrigation at a land application site. Two sites were studied both inside and outside the 

pivot irrigation system; however, small differences in measured HHCB concentrations were 

observed between the two locations. Three sample depths were studied: 0 to 6, 12 to 18, and 

24 to 30 inches below the surface. Over the time period of the study, at locations inside the 

irrigation system, concentrations ranged from <0.33 to 1.98 ng/g dw at 0 to 6 inches, <0.33 to 

1.25 ng/g dw at 12 to 18 inches, and ND to 1.25 ng/g dw at 24 to 30 inches. For locations 

outside the irrigation system, concentrations ranged from <0.33 to 5.69 ng/g dw at 0 to 

6 inches, <0.33 to 0.77 ng/g dw at 12 to 18 inches, and <0.33 to 0.79 ng/g dw at 24 to 30 inches. 

The highest reported value (5.69 ng/g dw) was measured outside the irrigation system in the 

winter of 2009; however, excepting this value, concentrations were below 1.98 ng/g dw over 

the time period of the study. Soil concentrations were found to be lowest during the summer 

months, possibly due to volatilization. Groundwater was also sampled and found to range from 

ND to <5 ng/L. 

In 2008, Kinney et al. (2008) reported HHCB concentrations in agricultural soil amended with 

biosolids. Three sites were studied. Site 1 (minimally affected site) was a non-irrigated soybean 

field that had not been amended with either human or livestock waste for at least the previous 

seven years. Soil and earthworm samples were collected from this field in June and September 

of 2005. Site 2 (biosolid amended site) was a no-till, non-irrigated soybean field receiving 

biosolid as a fertilizer for the first time. Site 3 (manure-amended site) was a non-irrigated 

cornfield receiving liquid swine manure as an organic fertilizer. While HHCB was not detected 

at site 3, HHCB was detected at site 1 in June 2005 at measurable levels, up to 633 μg/kg dw. 

HHCB concentrations were highest at site 2 and ranged from 1,050 to 2,770 μg/kg dw. 

Table G-5 summarizes the measured concentrations in HHCB in soil.
DRAFT DO NOT CITE OR QUOTE

Page 107

Table G-5. Measured Concentrations of HHCB in Soil.

Location Year Site Soil Concentration (μg/kg dw) Reference

Midwest, US

Agricultural fields

2005 Minimally affected ND-633

Kinney et 

al. (2008) Biosolid amended 1,050-2,770

Manure amended ND

Lubbock, TX

Land application 

site

2009-

2010

0-6 inches 12-18 inches 24-30 inches

Chase et al.

(2012)

Inside pivot irrigation <0.33-1.98 <0.33-1.25 ND-0.90

Outside pivot 

irrigation

<0.33-5.69 <0.330-0.77 <0.33-0.79

Ontario, Canada 2003 Biosolid amended 2.0 ± 0.1 (after 1 day)

2.8 ± 0.4 (after 2 weeks)

Yang and 

Metcalfe 

(2006)

ND Not detected

A Canadian study of biosolid-amended agricultural fields similarly showed that HHCB persisted 

in soils for the first two weeks post application, but concentrations declined thereafter (Yang 

and Metcalfe, 2006). More studies are needed, however, to understand the influence and 

integration of time, degradation mechanisms (such as volatilization, transformation, and 

leaching), and organic matter content on long-term concentrations. Soil studies, although 

clearly limited, do indicate that land application of treated wastewater effluent or biosolids 

results in detectable quantities of HHCB in soil and hence represents a potential route through 

which HHCB may enter terrestrial ecosystems. 

Measured Concentrations in Wildlife Biota

The liver, fillet, and blubber of many aquatic species, including fish, shrimp, zebra mussels, and 

aquatic mammals and birds, have been sampled from US waters in a variety of locations

(Table G-6). These data confirm the widespread occurrence of HHCB in aquatic media and 

subsequent exposure to wildlife. 

Six geographical locations in various parts of the US were selected as sampling sites for five 

effluent-dominated rivers receiving discharge from WWTPs located in Chicago, Illinois; Dallas, 

Texas; Orlando, Florida; Phoenix, Arizona; and West Chester, Pennsylvania; and one reference 

site on the Gila River, New Mexico (Ramirez et al., 2009). The reference site was expected to be 

removed from anthropogenic point sources; therefore, no accumulation of HHCB was expected 

or detected in fish collected from this site. A total of 18 to 24 adult fish of the same resident 

species were collected from each sampling location during late summer and fall of 2006. Fish 

sampled at each site were divided into six composites, each containing three or four adult fish 

of similar size. The composites revealed the presence of HHCB at maximum concentrations 

ranging from 300 to 2,100 ng/g tissue.

Concentrations of polycyclic musks, including HHCB, in fish were collected in 2006 from Troy, 

Albany, and Catskill along the upper Hudson River in the eastern region of the state of New 
DRAFT DO NOT CITE OR QUOTE

Page 108

York (Reiner and Kannan, 2011). There are 148 WWTPs that discharge treated wastewater into 

the Hudson River, which flows southward from its headwaters in the Adirondack Mountains

and ultimately empties into the Atlantic Ocean at New York City. The measured levels of HHCB

varied across fish species, with a range of <1 to 39 ng/g lw; the overall highest concentrations 

were measured in white perch liver.

Shrimp samples were collected from a seafood market survey of wild and farmed shrimp from 

the US and other countries (Mexico, India, Ecuador, Thailand, China, and others) in 2006

(Sapozhnikova et al., 2010). The shrimp were analyzed for the presence of synthetic musks. 

HHCB was detected in all samples, with concentrations ranging from 48 to 683 ng/g lipid 

(average 199 ng/g lipid weight) in farmed shrimp and 66 to 762 ng/g lipid (334 ng/g lipid 

weight) in wild shrimp. Farm-raised shrimp from Indonesia (n = 1) showed the highest 

concentration of HHCB, followed by farm-raised shrimp from the US (n = 6). No trends could be 

discerned as the sampling was not statistically significant; however, the presence of HHCB in 

shrimp tissues, both farmed and wild caught, indicates the widespread exposure of aquatic 

biota to HHCB.

Kannan et al. (2005) reported the presence of HHCB in the tissues of higher trophic level 

aquatic organisms and humans. Among the wildlife species analyzed, tissue concentrations 

ranged from <1 to 25 ng/g ww. Blubber tissue obtained from spinner and bottlenose dolphins 

from coastal waters off the state of Florida contained the highest levels of HHCB. 

Seven to eight carp were collected 200 meters from the drinking water intake area on Lake 

Mead, Nevada on a monthly basis over a period of one year (Osemwengie and Steinberg, 2001). 

The measured HHCB concentration in the fish tissue ranged from 1.4 to 4.5 ng/g ww with an 

annual mean concentration of 3.0 ng/g ww.

Kinney et al. (2008) measured HHCB concentrations in earthworm tissue obtained from 

amended soils where biosolids or liquid swine manure was applied. These values were 

compared to values reported for worms collected from a soybean field, which had not been 

amended with human or livestock waste for at least the previous seven years. HHCB 

concentrations in worms 30 days post application (3,340 μg/kg ww) were an order of 

magnitude higher than the levels measured after 156 days (131 μg/kg ww). Tissue 

concentrations in worms from the manure amended site were 49 μg/kg after 139 days.
DRAFT DO NOT CITE OR QUOTE

Page 109

Table G-6. Measured Concentrations of HHCB in Wildlife Biota.

Species n Location Year HHCB Concentration 

(ng/g) Reference

Hudson River HHCB

(ng/g lipid weight)

White perch (liver) 3 Troy, NY

Albany, NY

Catskill, NY

2006 6.27-19.9

7.58-22.5

13.7-27.9

Reiner and 

Kannan (2011)

Channel catfish (liver) 3

1

Troy, NY

Catskill, NY

11.1-39

<1

Smallmouth bass (liver) 3

3

3

Troy, NY

Albany, NY

Catskill, NY

<1-11.1

1-31.9

<1

Largemouth bass (liver) 1

1

Albany, NY

Catskill, NY

10.9

8.22

White catfish (liver) 1

1

Albany, NY

Catskill, NY

6.56

5.79

Brown bullhead (liver) 3 Catskill, NY <1-51.1

Zebra mussel 4 Catskill, NY 10.3-19.3

American eel (whole 

body)

1 Catskill, NY 125

Shrimp 6

3

US 2006 330, wild caught

424, farm raised

Sapozhnikova

et al. (2010)

HHCB

(ng/g tissue 

concentration)

Earthworm

3

3

3

Midwestern US

-minimally affected 

site

-biosolid amended site

-manure amended site

2005

61a

; NDb

3,340c

; 131d

NDe

; 49f

Kinney et al.

(2008)

Largemouth bass 6 Chicago, IL

North Shore Channel

2006 1,300 mean; 1,800 

max

Ramirez et al.

(2009)

Smallmouth buffalo 6 Dallas, TX

Trinity River

800 mean; 1,800 max

Bowfin 6 Orlando, FL

Little Econlockhatchee 

River

100 mean; 300 max

Common carp 6 Phoenix, AZ

Salt River

1,800 mean; 2,100 

max

White sucker 6 West Chester, PA

Taylor Run

1,800 mean; 2,000 

max
DRAFT DO NOT CITE OR QUOTE

Page 110

Table G-6. Measured Concentrations of HHCB in Wildlife Biota.

Species n Location Year HHCB Concentration 

(ng/g) Reference

HHCB (ng/g) ww

Sea otter (liver) 8 Monterey Bay, CA 1993-1999 <1-3.2g Kannan et al.

Harbor (2005) seal (liver) 3 Central CA coast 1996-1997 4.4-5.5; 4.8 mean

California sea lion (liver) 3 Central CA coast 1993-1996 1.5-4.4; 2.8 mean

River otter (liver) 3 Grand River, MI 1997 2.4-3; 2.8 mean

Bottlenose dolphin 

(blubber)

4 FL coast 1994-2000 4.2-20.5; 12 mean

Striped dolphin (blubber) 3 FL coast 1995-1997 8.1-25; 14 mean

Pygmy sperm whale 

(blubber)

1 FL coast 2000 6.6

Atlantic sharpnose shark 

(liver)

3 Indian River Lagoon, FL 

coast

2004 4.6-5.2; 4.8 mean

Mink (liver) 4 Aurora/Plainfield, IL 1997 2.2-5.3; 3.7 mean

Common merganser 

(liver)

2 Buffalo Harbor, NY 1999 3.7-4.2; 4 mean

Greater and lesser scaup 

(liver)

2 Niagara River, NY 1995-1999 1.9-2.7; 2.3 mean

Mallard (liver) 1 North Tonawanda 

Creek, NY

1995 2.7

Atlantic salmon (skin on 

fillet)

6 Farmed and wild, local 

market NY

2003 <1-3.2h

Smallmouth bass(liver) 3 Effley Falls Reservoir 

and Rock Pond, NY

2003 4.3-5.4; 4.8 mean

Carp 84 Lake Mead, NV 2001 1.4-4.5; 3.0 mean Osemwengie 

and 

Gerstenberger 

(2004)

ND Not detected

a

 Measured June 2005 b Measured September 2005 a, c 31 days after application of biosolids. b, d 156 days after application of biosolids. c, e 30 days after application of liquid swine manure. d, f

139 days after application of liquid swine manure. e, g 38 percent of samples contained detectable concentrations. f, h 83 percent of samples contained detectable concentrations.
DRAFT DO NOT CITE OR QUOTE

Page 111

USGS National Water Quality Information System Data

Monitoring data collected by the USGS NWQL and available from the NWIS database up to May 

2012 were obtained for environmental concentrations of HHCB within the US (USGS, 2012). 

The following categories of HHCB data were available from the USGS NWIS: water, filtered 

recoverable; water unfiltered, recoverable; and solids recoverable, dw. These data were 

grouped by medium type and site type, as shown in Table G-7. The Medium and Site Code 

definitions are provided in Table G-8.

The data compiled herein include values that are reported as less than the USGS LRL; values 

that are between the LRL and the LT-MDL; and values that are below the LT-MDL (Oblinger 

Childress et al., 1999). Similar categories of data were available for both types of water 

samples. However, for unfiltered water from effluent/stream, surface water/outfall, and 

groundwater/well sites, significantly fewer than 10 data points were available, and were thus 

not incorporated into the summary plots.

For this assessment, data contained in three NWIS parameter codes (i.e., 62075, 62823, and

63209) for HHCB were included. The LRL for water sampling was 0.5 μg/L for sampling dated 

7/16/2001 to 9/30/2009. The LRL was updated to 0.05 μg/L for samples dating from 10/1/2009

to the present based on a re-evaluation of the LT-MDL by the USGS. Interim reporting levels 

were recorded for data collected from solid samples, based on USGS practices for data 

interpretation. For monitoring data sets where the geometric standard deviation was <3.0, 

values recorded as “less than LRL” or “estimated” were replaced by the LRL divided by the 

square root of two, as per the US EPA OPPT guidance document (US EPA, 1994). Likewise, 

where the geometric standard deviation was >3.0, values recorded as “less than LRL” or 

“estimated” were replaced by the LRL divided by two (US EPA, 1994). It should be noted, 

therefore, that resulting low-end values are biased toward the LRL. This practice presents a 

conservative low-end value, which protects against false negative values. As such, values do 

not necessarily represent quantitative measured concentrations.

For each of the data groupings (Table G-7), box plots were generated with the calculated mean,

median, 1st (Q1) and 3rd (Q3) quartile, and 5th and 95th percentile values as shown in Figures G-1 

to G-3. A summary of these calculated values is presented in Table G-7. 
DRAFT DO NOT CITE OR QUOTE

Page 112

Table G-7. Summary of Box Plots for USGS HHCB Data.

USGS Parameter Code

Medium/Site

HHCB Concentration

5th

Percentile Q1 Median Mean Q3 95thPercentile

62075

Water, filtered, recoverable; HHCB (μg/L)

Effluent/stream 0.35 0.35 1.40 1.18 1.82 2.16

Effluent/outfall 0.09 0.35 0.76 0.98 1.2 3.40

Surface water/outfall 0.35 0.35 0.65 1.08 1.70 2.30

Surface water/stream 0.10 0.35 0.35 0.35 0.35 0.35

Surface water/lake, reservoir, impoundment 0.04 0.35 0.35 0.33 0.35 0.35

Groundwater/well 0.04 0.35 0.35 0.32 0.35 0.35

62823

Water, unfiltered, recoverable; HHCB (μg/L)

Effluent/outfall 0.14 0.14 1.02 1.01 1.85 2.26

Surface water/stream 0.02 0.05 0.14 0.12 0.14 0.15

Surface water/lake, reservoir, impoundment 0.03 0.14 0.14 0.12 0.14 0.14

63209

Solids, recoverable, dry weight; HHCB 

(μg/kg)

Bottom material/lake, reservoir, 

impoundment 27.93 41.02 65.42 87.46 106.08 212.87

Bottom material/stream 17.68 24.75 40.00 67.99 67.19 200.00
DRAFT DO NOT CITE OR QUOTE

Page 113

Table G-8. USGS Medium and Site Codes.

Medium Codes

Surface water Water on the surface of the earth stored or transported in rivers, 

streams, estuaries, lakes, ponds, swamps, glaciers, or other aquatic 

areas. It also may refer to water in urban drains and storm-sewer 

systems.

Effluent Treated or untreated wastewater after use at a facility or wastewater 

treatment plant, or from combined sources, such as combined-sewer 

overflows or tile drainage systems.

Groundwater Water below the surface of the earth contained in the saturated zone.

It does not include soil moisture or interstitial water.

Bottom material A mixture of mineral and organic matter that compose the top bed 

deposits (usually the first few inches) underlying a body of water.

Site Codes

Outfall A site where water or wastewater is returned to a surface-water body 

(e.g., the point where wastewater is returned to a stream). Typically, 

the discharge end of an effluent pipe.

Lake, reservoir, impoundment An inland body of standing fresh or saline water that is generally too 

deep to permit submerged aquatic vegetation to take root across the

entire body (Cf. wetland). This site type includes an expanded part of 

a river, a reservoir behind a dam, and a natural or excavated 

depression containing a water body without surface-water inlet 

and/or outlet.

Well A hole or shaft constructed in the earth intended to be used to locate, 

sample, or develop groundwater, oil, gas, or some other subsurface 

material. The diameter of a well is typically much smaller than the 

depth. Wells are also used to artificially recharge groundwater or to 

pressurize oil and gas production zones. Additional information about 

specific kinds of wells should be recorded under the secondary site 

types or the Use of Site field. Underground waste-disposal wells 

should be classified as waste-injection wells.

Stream A body of running water moving under gravity flow in a defined 

channel. The channel may be entirely natural, or altered by 

engineering practices through straightening, dredging, and (or) lining. 

An entirely artificial channel should be qualified with the "canal" or 

"ditch" secondary site type.

Figure G-1 shows box plots for filtered water from three different mediums and associated 

sites. HHCB concentrations are clearly highest from effluent with higher values from outfall 

sites, although the range of values varied from ND to >3 μg/L. The mean effluent 

concentrations were comparable, but the range of concentrations was greater at outfall sites. 

This may be due to the limited data that were available for effluent/stream sites, such that data 

as summarized here encompasses only 10 data points collected from four individual sites. 

Mean values for filtered water were <1.5 μg/L for all medium and site types. Groundwater
DRAFT DO NOT CITE OR QUOTE

Page 114

concentrations at well sites and surface water concentrations at streams and lakes, reservoirs,

and impoundment sites ranged from ND to <0.5 μg/L. For effluent/outfall filtered/non-filtered 

samples, the HHCB concentrations were higher in unfiltered samples (Figure G-2). This may be 

attributed to the presence of HHCB in suspended solids. Again, significantly lower 

concentrations were consistently recorded in surface water samples collected from streams 

and lakes, reservoirs, and impoundment sites—areas that are unaffected or nominally affected 

by wastewater treatment outfall. 

Figure G-1. Monitoring Data Summary from USGS NWIS for HHCB Concentrations in Water, Filtered.
DRAFT DO NOT CITE OR QUOTE

Page 115

Figure G-2. Monitoring Data Summary from USGS NWIS for HHCB Concentrations in Water, 

Unfiltered.

Figure G-3. Monitoring Data Summary from USGS NWIS for HHCB Concentrations in Solids.

The majority of the data available from the USGS NWIS data set for solids was from bottom 

material. Bottom material data were predominantly divided between two types of sites: 

lake/reservoir/impoundment and streams (Figure G-3). The 95th percentile concentration range 

at lake/reservoir/impoundment sites was 213 μg/kg, and the mean concentration was 

87 μg/kg. The 95th percentile concentration range at stream sites was 200 μg/kg, and the mean 

concentration was 68 μg/kg. The range of data was similar for both site types, with the Q3 
DRAFT DO NOT CITE OR QUOTE

Page 116

value for lake/reservoir/impoundment sites slightly higher than that observed for stream sites. 

Though it is unclear what this may be attributed to, the difference may be reflective of

differences in sampling locations or loss mechanisms.