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Page 1 United States Office of Chemical Environmental Protection Safety and Pollution Agency Preven tion



# TSCA Workplan Chemical Risk Assessment

HHCB 1,3,4,6,7,8

- Hexahydro

- 4,6,6,7,8,8

- hexamethylcyclopenta

- γ
- 2

- benzopyran

CAS RN: 1222

- 0 5

- 5

C H 3 C H 3 C H 3 C H 3 C H 3

O C H 3



### 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 circul ated for review of its technical accuracy and science policy implications.

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## Table of Contents

AUTHORS / CONTRIBUTO RS / ACKNOWLEDGEMENT S / REVIEWERS

................................ .................... 5

GLOSSARY OF TERMS AN D ABBREVIATIONS

................................ ................................ ....................... 6

EXECUTIVE SUMMARY

................................ ................................ ................................ ....................... 8

CHAPTER 1: BACKGROU ND AND SCOPE

................................ ................................ ........................... 10

INTRODUCTION

................................ ................................ ................................ ................................ ... 10
A. S COPE OF A SSESSMENT FOR HHCB

................................ ................................ ................................ ....... 12
B. C ONCEPTUAL M ODEL FOR E NVIRONMENTAL A SSESSMENT OF HHCB

................................ .......................... 12

CHAPTER 2: SOURCES AND ENVIRONMENTAL FA TE

................................ ................................ .......... 15

INTRODUCTION

................................ ................................ ................................ ................................ ... 15
A. P HYSICAL AND C HEMICAL P ROPERTIES OF HHCB

................................ ................................ ..................... 15
B. E NVIRONMENTAL F ATE

................................ ................................ ................................ ........................ 17
C ONCLUSIONS OF E NVIRONMENTAL F ATE

................................ ................................ ................................ .... 21
C. P RODUCTION AND U SES

................................ ................................ ................................ ....................... 21
C ONCLUSIONS ON P RODUCTION V OLUME AND U SES

................................ ................................ ..................... 26

CHAPTER 3: ENVIRON MENTAL ASSESSMENT

................................ ................................ ................... 27

A. E NVIRONMENTAL E XPOSURE A SSESSMENT

................................ ................................ .............................. 27
B. E NVIRONMENTAL H AZARD A SSESSMENT

................................ ................................ ................................ . 32
C. E NVIRONMENTAL R ISK C HARACTERIZATION

................................ ................................ ............................. 39
D. K EY S OURCES OF U NCERTAINTY AND D ATA L IMITATIONS

................................ ................................ .......... 41
C ONCLUSIONS OF E NVIRONMENTAL A SSESSMENT

................................ ................................ ......................... 42

REFERENCES

................................ ................................ ................................ ................................ .... 43

APPENDIX A: HUMAN H EALTH TOXICITY STUDI ES, BIOMONITORING DA TA, AND RISK ASSESSM ENTS 61

A

- 1. H UMAN H AZARD C HARACTERIZATION

................................ ................................ ................................ . 61
A

- 2. H UMAN B IOMONITORING

................................ ................................ ................................ ................. 67
A

- 3. S UMMARY OF 2008 EU H UMAN H EALTH R ISK A SSESSMENT

................................ ................................ .. 70
A

- 4. K EY S OURCES OF U NCERTAINTY AND D ATA L IMITATIONS ON H UMAN H EALTH

................................ ........... 74
A

- 5. C ONCLUSIONS OF H UMAN H EALTH A SSESSMENT

................................ ................................ .................. 75

APPENDIX B: CHEMICA L SYNTHESIS OF HHCB

................................ ................................ ................... 77

APPENDIX C: HHCB DI ASTEREOISOMERS (#1 T O #6) AND RELATED ST RUCTURAL ANALOGS (#7 TO #15)

................................ ................................ ................................ ................................ ........................ 7 8

APPENDIX D: IUR DAT A FOR HHCB

................................ ................................ ................................ ... 81

APPENDIX E: MODELED RELEASE ESTIMATES AC CORDING TO STAGE OF PRODUCTION AND USE

....... 83

APPENDIX F: ENVIRON MENTAL FATE AND BIOA CCUMULATION STUDY SU MMARIES

........................ 92

APPENDIX G: ENVIRONMENTAL MONITO RING DATA ANALYSIS

................................ ........................ 95

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## 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 V olume 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 HHC B.

................................ ................................ ............................. 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
Ta ble E

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

................................ ................................ ................................ ....................... 88 Table E

- 4. Geograph ic 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 Concent rations 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

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## List of Figures

Fig ure 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 Conce ntrations 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

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## Authors / **C ontributors**  **/ 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
Ev a 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 Review ers 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 influent ial 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 i ndependent expert s

.

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## Glossary of **T erms and**  **A bbreviations** 

AHTN 6

- Acetyl

- 1,1,2,4,4,7

- hexamethyltetraline
BB Benzyl benzoate
BAF Bioaccumulation factor
BCF Bioconcentration factor
CAS RN C hemical A bstract S ervice R egistry N umber
CBI Confidential Business Information ChV Chronic value
COC Concentration of c oncern DEP Diethyl phthalate dw Dry weight
E 2 

### E stradiol
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/ m ass spectrometry GD Gesta tion day
GLP Goo d Laboratory Practice
HERA Human and Environmental Risk Assessment Project
HHCB 1,3,4,6,7,8

- H exahydro

- 4 ,6,6,7,8,8

- hexamethylcyclopenta

- γ

- 2

- benzopyran
HPLC High p erformance l iquid c hromatography
HPV High production volume
IFRA International Fragrance Association
IPM Isopropyl myristate
IUR Inventory Update Reporting Rule k 1 

### Uptake rate constant k 2 

### Elimination rate constant
K d 

### Sorption coefficient kg Kilogram (s)
K OC 

### Organic carbon

- normalized sorption coefficient
K OW 

### 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

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### Log K OC 

### Logarithmic organic carbon partition coefficient Log K OW 

### 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
NOAE C No

- observed

- adverse

- effect concentration
NOAEL No

- o bserved

- a dverse

- e ffect l evel
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, e valuation, a uthorisation and r estriction of c hemicals
RHO earthworm 

### Bulk density of earthworm
RAR Risk assessment report RQ Risk q uotient
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
UF total 

### Total uncertainty factor
US United States
USGS United States Geological Survey
wt% Weight percent

ww W et weight
WWTP W astewater treatment plant yr(s) Year

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## 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 U nited 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 ex h ibits moderate environmental persistence and moderate bioaccumulation potential.

Th is assessment evaluated the environmental risk to the aquatic and terrestrial environment s 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 speci es, 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 conce rn. 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, hum an health risk was not assessed further

.

HHCB is one of the most widely used polycyclic musk fragrance ingredient s in a range of consumer products including perfumes, cosmetics, shampoos, lotions , detergents, fabric softeners, and household clean ers

. HHCB is not produced in the US and is imported by four importers. T he 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 increase d steadily since 2000

, suggest ing 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 release d 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 likel y 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 sa mpling measurements were available for effluent, surface water and sedime nt from the USGS National Water Information System ( NWIS ) database

. These values represent a range of date s and location s across the US. In addition, s oil monitoring data was availa ble from 2 published studies.

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### 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 orga nisms with effects on adult and embryo survival, behavior, respiration, and larval development. Th is assessment considered only direct exposure to aquatic species, sediment

- dwelling species, terrestrial species, and plants via contact with contaminated wa ter, sedim ent, 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 t he 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 95 th percen tile of measured environmental concentrations for surface water and sediment. A vailable soil concentrations from two published studies were used to calculate an RQ for the maximum

measure d soil concentration.

The a cute 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 sediment

dwelling species at the upper range of measur ed environmental concentrations for l ess than 1.5% of the more than 600 sediment measurements from the USGS. Though the data were limited, c hronic 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 t errestrial invert ebrate s and plants. T he 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 avai lable 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 strength s 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

.

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## 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 a n ingredient in a wide range of consumer products including perfumes, cosmetics, shampoos, lotions

, detergents, fabric softeners, air freshen ers, and clean ing agents

. HHCB has moderate hazard potential for developmental toxicity 1



### , high potential exposure, and has been detected in human s. EPA estimated potentially high releases of HHCB to the environment

. HHCB presents mod erate 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 i s 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

- to

Know (EPCRA) Section 313 List of Toxic Chemicals. HHCB is approved as an in ert 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

, ” bu t 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 test ing 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 r isk conce rns were identified for certain highly contaminated sediments in

1 Human health risks from HHCB uses we re 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 is opropyl 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 )

.

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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 ide ntified 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 )

. T his review of HH CB 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 b iotransformed to less toxic, less bioaccumulative

, polar metabolites.

• HHCB has high acute and chronic toxicity for aquatic organisms. Chronic effects of
HHCB in a quatic organisms include e ffects on adult and embryo survival, behavior, respiration, and larval development. M echanistic studies show that HHCB may interfer e 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 co ncern for human health hazards, based on minimal developmental toxicity at relatively high oral exposures and no developmental toxicity at lev els 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 alternativ e, 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.

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## A. S COPE O F 

## A SSESSMENT F OR 

## HHCB

B ased 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. T he use of products containing HHCB

, such as personal care products and cleaning products, result s 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 envi ronment.

Th re e 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 p opulation

- or community

- level ecological endpoints

were beyond the scope of this assessment

.



## B. C ONCEPTUAL 

## M ODEL F OR 

## E NVIRONMENTAL 

## A SSESSMENT 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 a cute 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 Mo del ( Figure 1

- 1 ) and described below

.

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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 conta ining HHCB

, the majority of
HHCB enters the wastewater stream . O ther minor releases to the environment also we re estimated (see Appendix E : Release Estimates )

. Due to the ubiquity of fragranced products,

consumer use s are a continuing source of HHCB to w aste water

. When HHCB enter s wastewater treatment plants ( WWTP s )

, 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 t o soil

.

During the wastewater treatment process , influent is separate d into two components: wastewater effluent and solids, known as either sewage sludge or biosolids. Wastewater effluent is released to surface water at outfall site s 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 ( SSO s ) occur each year in the US , resulting in releases of billions of gallons of untreated wastewater directly to surface water

. When HHCB ha s not been removed during the treatment process, adsor ption to suspended solids in the surface water and partition ing 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 t he 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

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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 a bout 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 plant s and soil invertebrates

, leading to potential exposure and toxicity

.

Th is environmental assessment evaluated HHCB exposure s for the scenarios described above based on available monitoring data for the US an d assess ed the potential ecological hazard s for

aqu atic, benthic, and soil species by calculating concentration of concerns ( COCs )

. The potential risk s to aquatic, benthic, and soil species w ere characteriz ed by calculating risk quotients ( RQ s ) for each route of environmental exposure ( i.e.

, water, sediment, and soil) based on measured concentrations

. Because ex posure 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 pur poses, 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 )

.

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## 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. P HYSICAL AND 

## C HEMICAL 

## P ROPERTIES 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.

C H 3 C H 3 C H 3 C H 3 C H 3

O C H 3

Figure 2

- 1. Chemical S tructure of HHCB ( HSDB, 2007 )

.



### S ynonyms for HHCB include:

• 1,3,4,6,7,8

- hexahydro

- 4,6,6,7,8,8

- hexame t hyl

- cyclopenta

- gamma

- 2

- benzopyran

• 1,3,4,6,7,8

- hexahydro

- 4,6,6,7,8,8

- hexame thyl

- 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 f our diastereoisomers ( i.e.

, a n 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 ;

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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

c ommercial 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 f ormula C 18 H 26 O

Molecular w eight 258.44
Physical f orm Colorless liquid; highly viscous liquid at 20 ° C and 1

, 013 hPa with a musk odor
Melting p oint

- 10 to 0 ° C;

- 20 ° C
Boiling p oint 320 ° F at 760 mmHg (converted from 160 ° C at 4 hPa)
Vapor p ressure 0.0727 Pa (5.47 × 10 - 4 mmHg) @ 25 ° C (measured; Organisation for
Economic Co

- operation and Development [ OECD ] 104)
Logarithmic o ctanol: w ater
p artition c oefficient ( log K OW ) 5.3 ("slow stirring" method); 5.9 @ 25 ° C (measured; OECD 117)

Water s olubility 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 p oint 144 ° C

Source: (HSDB, 2007)

.

4 EPA recently proposed a ban for DEP as an inert ingredient in pesticide products ( US EPA, 2012c )

.

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Page 17 

## B

. E NVIRONMENTAL 

## F ATE

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) an d target organisms of concern. HHCB enters domestic wastewater treatment and is subjected to removal by a variety of processes. HHCB and any degradation products genera ted 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
R adiolabeled HHCB (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

appear ance of more polar metabolites. The half

- life for HHCB in activated sludge was 21 hours (first

- order rate constant = 0.033 h ours - 1 

### )

, and 85 percent of the HHCB initially pre sent 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/ex posure/pubs/episuite.htm and http://www.epa.gov/oppt/exposure/pubs/halflife.htm

.

HHCB is expected to partition strongly to solid phases b ased on its high measured log K OW 

### of 5.9 ( see Rimkus, 1999 for a summary of values for musks ) and estimated soil/sediment o rganic carbon partition coefficient ( K OC 

### ; 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 pri mary and secondary sludge ( EC, 2008 )

. Values for both K d 

### (sorption coefficient) and K OC 

### (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 ) summarize d both theoretical

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Page 18 a nd measured percent removal values for European and US activated s ludge

- 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 w as 33 to 43 hours, depending on the starting concentration

. After 12.5 days

, about 40 percent of the radioactivi ty 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 disappearanc e of HHCB in river water ( Schaefer et al

., 2005; as reported in E C , 2008 )

. A significant fraction (up to 16 percent ) was volatilized after 28 days. V olatilization 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 disappearan ce of 2 to 4 days were observed ( see Table 2

- 2 )

. Degra dation is expected to be slower in streams that do no t 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 routine ly appli e s 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 f our 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 o xide

- rich soil. The initial HHCB concentrations in spiked soil were 6 and 13

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Page 19 mg/kg soil, whereas the levels in unspi ked 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 r anged 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 lowe r 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 sludg e

- amended soils, respectively.

Table 2

- 2

. HHCB D egradation H alf

- L ives and H alf

- D isappearance T imes for E nvironmental M edia

.

Source Degradation H alf

- L ife or H alf

- D isappearance T ime a

Water Sediment Soil Sludge Sludge

- A mended
S oil

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 hours b 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 differ s from degradation half

- li fe in that it include s all mechanisms contributing to disappearance ( e.g. , volatilization in addition to degradation )

. b Water with 10 mg/ L suspended solids.



### O bserved soil and sediment half

- lives consistently exceeded 60 days (Table 2

- 2). Field

measurements on sludge

- amended soil indicate d 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,

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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 

### cm 3 

### 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 impor tant process for HHCB.

BIOACCUMULATION AND BIOCONCENTRATION

Based on available data, HHCB is considered a moderate concern for bioaccumulation.
B ioaccumulation 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.
T hese studies

, together with monitoring data in biota, suggest th at HHCB is not subject to biomagnification . Table 2

- 3 summarizes published bioaccumulation data for aquatic species.

Table 2

- 3

. BCF s and BAF s ( L/kg w w ) of HHCB in V arious A quatic S pecies

. Species Common N ame BCF BAF

Scardinius erythrophthalmus Rudd 20

T

inca tinca Tench 510

Carassius carassius Crucian carp 580
Anguilla anguilla Eel 862 290
Dreissena polymorpha Zebra mussel 620
Lepomis macrochirus Bluegill sunfish 1,584

D anio rerio Zebrafish 620

Source : Dietrich and Hitzfield (2004 )

.



### Jager (1998 ) ( as cited in EC, 2008 ) reported a calculated BCF for earthworms ( Lu m bricus
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 ) :

BCF worm 

### = (0.84 + 0.012 × K ow 

### )/RHO earthworm

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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 2,395 a Jager (1998)

a Calculated.



## C ONCLUSIONS OF 

## E NVIRONMENTAL 

## F ATE

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

m easured BAFs and BCFs for HH CB suggest bioaccumulation potential in the low range, but
HHCB is considered moderately bioaccumulative for some species.



## C. P RODUCTION AND 

## U SES

This section discusses the US production volumes and uses specifically for HHCB and more generally for fragrances. HHCB is a H igh P roduction V olume chemical that is used in a vast number of cleaning and personal care products. D ata on production volume and use s are amenable for determining release estimat es to the environment

. T hese 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

- a cetyl

- 1,1,2,4,4,7

- hexamethyltetraline; CASRN 1506

- 02

- 1 ; also CASRN
2114 5

- 77

- 7) are the most widely used of the polycyclic musks. Together

, HHCB and AHTN represent 90 percent of the U S polycyclic musk market ( HERA, 2004a )

, with HHCB being the most widely consu med of the polycyclic musks ( EC, 2008 )

. Four US firms reported non

confidential 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.

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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 HHC B 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 data a

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 )

. Globa l 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 m usks ( Rimkus, 1999 ; as cited in Luckenbach and Epel, 2005 )

. This shift, reflected in decreasing production rates of nitro musks and increasing production rate s of polycyclic musks ( Hutt er 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 )

. A dditionally, in June 2009 , t he International Fragrance Association (IFRA) voluntarily banned

m usk 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, i n 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 )

.

D ata from IFRA have show n 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

.

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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 refl ection 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 e xample, 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 level s for both HHCB and musk xylene suggests that there ha s 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 T housands, K, or M illions, 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 EP A (2012d ) ; US EPA (2012e )

.



### Uses Musks are considered important compounds to the f ragrance 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 discolo r 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 )

. Synth etic 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 la beled “fragrance free” ( Potera, 2007 )

.

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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 m ixture

, 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 compound s because fragrance mixtures can contain from 50 to 300 ingredients ( H ERA, 2004a ; Bickers et al. , 2003 )

. As shown in T able 2

- 7 , the majority of fragrance oils (25 percent) are used in detergents, followed by fabric softeners (14 percent) and personal care products (13 p ercent), 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 t he 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 c are 13

Bath and s hower 10

Hair care 10
Soaps 9

Industrial and h ousehold c leaner 8

Other 6
Fine fragrances 5

Source: Balk et al. (2001 ) ( as cited in OSPAR, 2004 )

.

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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. , 200 2 )

. 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 HH CB 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, th e HHCB

lactone

, was <5 mg/kg in the same products. This study also showed that products within a single class might c ontain differing levels of HHCB. F or 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. H owever, these studies indicate that levels

var ied significantly across and even within product types. This variation is not unexpected given the complexity of fragrance mixtures which often contain hundreds of constituents.

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Page 26 

## C ONCLUSIONS ON 

## P RODUCTION 

## V OLUME AND 

## U SES

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).

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Page 27 

## Chapter **3 : Environmental A**  **ssessment**

INTRODUCTION

This assessment evaluated the environmental risk to the aquatic and terrestrial environment s 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 ) a n 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 measure d in the environment to the COCs determined from aquatic, benthic, and terrestrial toxicity studies.



## A. E NVIRONMENTAL 

## E XPOSURE 

## A SSESSMENT

F or this assessment, environmental monitoring data on measured levels of HHCB in wastewater, surface water, sedimen t, 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, t he 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 b lending 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 t he entire HHCB content of all of the various types of end

use products was rel eased down

- the

- drain after use

. Therefore, this assessment estimates that
90% of HHCB is release d to municipal wastewater after commercial and consumer use of products contain ing 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 t he 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

. T his database include s monitoring data as well as data collected for stu dies of organic anthropogenic waste indicators, organic wastewater contaminants, and biologically

- active chemical mixtures in urban streams

.

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Page 28 Although ten times more data was available from the USGS as compared to data from published literature studies f or effluent, surface water and sediment

, the measured HHCB
concentrations were within the same order of magnitude. EPA assumed that d ata 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 de tection and quantification of HHCB across all media. C oncentrations 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 provid ed in Appendix G

.

Wastewater In published studies, m easured 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: Ap pendix 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 w ere process dependent and concentrations were seasonally dependent and within an order of magnitude of those measured in the US (Smyth et al ., 2008).

T he 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 (95 th 

### percentile: 3400 ng/L). The mean effluent concentrations were comparable at stream and outfall sites

, but the range of concent rations

was greater at outfall sites.

Surface Water In published studies, s urface water concentrations of HHCB were found to range from non

detect ( ND ) to 1

, 6 00 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

t he 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 site s; the highest concentrations were measured at outfall si tes. The 95 th 

### 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). L ower concen trations were consistently

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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 publishe d 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 impact ed 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 concentration s 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
95 th 

### percentile concentration at lake/reservoir/impoundment sites was 213 μ g/kg, and the
mean concentration was 87 μ g/kg. The 95 th 

### percentile concentration at stream sites was 200

μ g/kg, and the mean concentration was 68 μ g/kg. The HHCB concent ration s (5%ile, mean, and 95%ile) w ere similar at both types of site s

.

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. Becaus e 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 obta ined 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). T he available data indicate that land application of treated waste water effluent or biosolids results in detectable quantities of HHCB in those soil s

.

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 scie ntific literature, and the greate st 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 repo rted 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).

USG S Data Analysis In addition to reported quantitative values, the USGS data set include s values that are reported as less than the USGS laboratory reporting level 7 

### (LRL ); data that are between the LRL and the long

- term method detection level (LT

- MDL); and da ta 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 rang e because of increased measurement uncertainty or values below the LT

MDL 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 di vided 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

a re biased towards the LRL.

H igh 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 aquati c biota in numerous studies. T hese data strongly suggest that HHCB is ubiquitous and pote ntially widespread in the environment

. The data are assumed to be representative.

7 The LRL for water sa mpling 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.

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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 (2007), Osemwengie (2004)

- Effluent >135 10 13,000 21

- 5310

- Effluent discharged to ocean n/a ND 2500

--- d Anderson (2012)

- Effluent (at stream or outfall) 47 90 a 3400 b 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 20 a 2300 b 120

- 1080 USGS NWIS (code 62075 and 62823)

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

Sediment ng/g

- Freshwater >59 ND 388 1.1

- 144 c Tertuliani (2008), Klecka (2010), Sapozhnikova (2010), Chase (2012), Reiner

(2011), Peck (2006)

- Bottom material (Lake, Resevoir, Impoundment, Stream) 606 17.68 a 212.87 b 67.99

- 87.46 USGS NWIS (code 63209)

Biosolids/Sludge >58 μg/kg 13 427,000 13

- 177,000 c Kinney (2006, 2008), Buyuksoonmez (2005), DiFrnacesco (2004)

Soil μg/kg

- Biosolids amended 6 1050 2770

--- d Kinney (2008), Chase (2012)

- Wastewater treated n/a <0.33 5.69

--- 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 5 th percentile value
b 95 th percentile value c Range of averages
d Range of means not available

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Page 32 

## B. E NVIRONMENTAL 

## H AZARD 

## A SSESSMENT

The environmental hazard reviewed and summarized in this section is based on previous
summaries including Balk a nd 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 peer

reviewed 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 acu t e 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 ratin g 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 T oxicity D ata for HHCB

— Acute Toxicity.

Test
O rganism Species Test Type Endpoint Value (mg/L) Hazard
R ating Reference

Bluegill sunfish L. macrochirus OECD TG
204 LC 50 0.452 High Balk and Ford (1999b )

Fathead
minnow Pimphales promelas OECD TG
210 LC 50 >0.140 High Balk and Ford (1999b )

Acute 96

- hour EC 50 0.39 High Dietrich and Hitzfield (2004 )

8 The US participat ed 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) accept ing the SIDS Initial Assessment Report (SIAR) under international agreements (http://webnet.oecd.org/hpv/ui/handler.axd?id=b36caa92

- a554

- 4853

- b98 d

- 744e72cb9d56).

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Page 33 Table 3

- 3. Aquatic T oxicity D ata for HHCB

— Acute Toxicity.

Test
O rganism Species Test Type Endpoint Value (mg/L) Hazard
R ating Reference

Zebrafish D. rerio Acute LC 50 96

- hour LC 50 96

- hour EC 50 0.452
>0.67
>0.67 High Dietric h and Hitzfield (2004 )

Acute 48

- hour NOEC/ LOEC >1.0/NE

P a,b Moderate Carlsson and Norrgren (2004 )

Water flea D aphnia magna OECD TG 202, part II, proposed updated
version

, June 1993 48

- hour EC

R 50 0.282 High Balk and Ford (1999b )

Marine copepod A cartia tonsa Acute

, static
ISO 1997 standard 48

- hour LC

R 10 LC

R 50 0.12

P a
0.47

P a High Wollenberger et al. (2003 )

Estuarine Copepods N orellia spinipes Life cycle 96

- hour LC

R 50 1.9

P a High Breitholtz et al. (2003 )

Freshwater mussel L ampsilis cardium Acute

, static renewal 24

- hour LC

R 50

48

- hour LC

R 50 1.0

- >1.75

P a

0.999

- >1.75

P a Moderate Gooding et al. (2006 )

Acute

, static renewal 96

- hour EC

R 50 0.153

- 0.831

P a High Gooding et al. (2006 ) Green algae Pseudo

kirchneriella subcapitata Acute
OECD TG
201 72

- hour EC

R 50 0.854 (growth) 0.72 (biomass) High Balk and Ford (1999b )

P a

P Nominal 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

P Not established.

LOEC = l owest

- observed

- effect concentration; NOEC = n o

- 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. T he
48

- hour LC

R 50

R 

### 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

.

Chroni c 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.

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Page 34 Table 3

- 4. Aquatic T oxicity D ata for HHCB

— Chronic Toxicity

. Test
O rganism Species Test Type Endpoint Value (mg/L) Hazard
R ating 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 C hronic 21

- day EC 50 0.282 Moderate Dietrich and Hitzfield (2004 )

Water flea D. magna C hronic 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
EC 10 EC 50 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 EC 10 5

- day EC 50 0.30/NE a
0.037
0.059 Moderate High c High c Wollenberger et al. (2003 ) d

Estuarine Copepods N. spinipes Life cycle 22

- day NOEC/LOEC
ChV 0.007/0.02 b

0.012 High c 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 well

documented studies were not considered for this risk assessment because the actual concentration of HHCB is uncertain

. For example, the measur ed 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. V olatilization or adsorption to organic material likely accounts for the disappear ance of HHCB from the water phase.

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Page 35 The principal study, a partial life cycle test (OECD D raft Guideline) with the marine copepod

, Acartia tonsa

, ( Bjornestad, 2007 ) was designed to prevent the form ation 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/m L 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 w ere measured.

In this study, t he actual test concentrations were >80 percent of nominal during the entire test. Larval mortality was 27.1 and 29.1 perc ent 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 dose

response 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 EC 10 

### (concentration affecting 10% of organisms) for LDR was 43.8 μg/ L (95 percent confidence interval 30.1

- 55.3 μg /L ), and the EC 50 

### was 115 μg/ L (131

- 153 μg/ L ).

Of the available studies, t he 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 Sedimen t

- 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 eff ects of sediment

- associated HHCB on growth, feeding rate, survival, and reproduction in the gastropod

, Potamopyrgus antipodarum

, were measured over 12 weeks i n 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. J uvenile 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 d w of sedi ment; a no observed effect concentration ( NOEC ) of 1.0 mg HHCB/kg d w of sediment w as reported , along with a calculated chronic value of 3.2 mg HHCB/kg d w of sediment ( Pedersen et al. , 2009 )

.

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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 T oxicity D ata for HHCB.

Test
O rganism Species Test Type Endpoint Value (mg/kg dw) Reference

Amphipod Hyalella

azteca OECD TG 218 LC 50 NOEC/LOEC
EC 50 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
EC 50 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 LC 50 

### was reported as 489 μg HHCB/g and the 14

- day LC 50 

### 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

.

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Page 37 Table 3

- 6. Soil T oxicity D ata for HHCB.

Test
O rganism Species Test Type Endpoint Value Reference

Springtail Folsomia candida ISO/CD 11267 NOEC/LOEC
ChV 45/105 mg/kg dw a 68.7 mg/kg dw a Balk and Ford (1999b )

Earthworm Eisenia fetida OECD TG 207 4

- week NOEC ChV 45/105 mg/kg dw a 68.7 mg/kg dw a

OECD TG 207

- Filter paper

contact screening test 24

- hour LC 50 48

- hour LC 50 72

- hour LC 50 32.60 μg/cm 2 11.87 μg/cm 2 6.14 μg/cm 2 Chen et al. (2011b )

OECD TG 207 7

- day LC 50 14

- day LC 50 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 LC 50 >194.6 mg/L Mori et al. (2006 )

Chronic (growth, maturation, reproduction) NOEC 9.8 mg/L

Wheat T riticum
aestivum Seed germination EC 50 EC 50 143.4 mg/L 422.3 mg/L An et al. (2009 )

Seedling development NOEC/LOEC NE b /0.2 mg/L

NOAEC/LOAEC (seed germination) ChV (seed germination)
LC 50 (seed germination)
EC 50 (root elongation)
EC 50 (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

.

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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 COC s for earthworms and wheat were 3.87 and 12.2 mg/kg dw soil

.

Add itional Studies
S tudies 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. W eak estrogenic activity of HHCB (purity unknown) was observed in competitive estrogen receptor bin ding 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 vitel logenin 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 with in 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 metab olic 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 ec d ysteroid agonist and antagonist activity of HHCB was assessed in an assay with the
Droso phila 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 tox icants 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 aquati c 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 Environment al Toxicity

. Acute

, a quatic s pecies 56.4 μ g/ L (56.4 ppm)

Chronic

, a quatic s pecies 5.24 μ g/ L (5.24 ppm)

Chronic

, s ediment

- d welling s pecies 320 μ g/kg d w
Chronic

, t errestrial i nvertebrates 3.87 mg/kg dw

Chronic

, t errestrial p lants 12.2 mg/kg dw



## C. E NVIRONMENTAL 

## R ISK 

## C HARACTERIZATION

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 addre ss 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?

T o answer th es e key questions, envir o nmental risk was characterized by ca lculating 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) wa s determined based on measured concentration s of HHCB (Table 3

- 2)

. The COCs reported in Table 3

- 7 served as the effect levels for aquatic, benthic, and terrestrial specie s

. 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.

F requency , timing, and duration of exposure affect s the potential for adverse effects in aquatic or terrestrial organisms

. A quatic exposure is assume d 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 e ven 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
95 th 

### 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 HH CB. Monitoring data was limited for HHCB in US marine

waters. T he RQ for soil was calculated only for the maximum concentrations in biosolid

amended 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 Risk s of HHCB

.

RQ Maximum Published
Concentration RQ Range of Mean Published
Concentrations RQ
95 th

P ercent ile
USGS
NWIS RQ
Mean
USGS
NWIS RQ Range

Acute risk

, a quatic 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

, a quatic

s pecies 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

, s ediment

d welling s pecies 1.2 1 a 0. 003 to 0.45 0.67 0.21 to
0.27 0. 003 to 1.2 1

Chronic risk

, t errestrial
i nvertebrates

- biosolids

- amended soil 0.72 No data b No data No data 0.72

- wastewater

- treated soil 0.0015 No data No data No data 0.0015

Chronic risk

, t errestrial

p lants

- 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 RQ s 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 RQ s 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 3 8 8 ng/g dw (Table 3

- 2)

. This HHCB sediment concentration is from a single sample taken from the upper Hudson Riv er near Troy, NY (Reiner, 2011) and a second sample (from the same study) taken near Alba ny, NY was 351 ng/g. B oth of these values are above the 95 th 

### 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 95 th 

### 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 c ontaminated areas

.

L imited 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 R Q s for this site w ere 0.72 for terrestrial invertebrates and 0.23 for terrestrial plants , which indicate no risk concern s for chronic effects of HHCB on terrestrial invertebrate species and plants using this limited data

.



## D. K EY 

## S OURCES OF 

## U NCERTAINTY AND 

## D ATA 

## L IMITATIONS

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, uncertaintie s need to be adequately characterized to understand the limitations of the approach

. This section discusses the qualitative l imitations of the monitoring data, ecological toxicity data, and risk characterization

.

Available monitoring data for HHCB in wa ter, sediment, and soil were used in the assessment. A robust dataset of o ver 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. T he 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. Calculat ion of the RQ based on the maximum value, the 95%ile

, and the highest mean value was performed in an ef fort to capture the full range of concentrations in the environment. Unless actual e nvironmental
concentrations of HHCB in surface water or sediment are greater than one or two orders of magnitude higher than expected, EPA assumes that additional measurem ents or estimates will not affect the risk calculation

.

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Page 42 Limited soil monitoring data for HHCB prevent ed 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 biosolid s or wastewater effluent. The locations are limited to two sites and may not represent the US given differences in sludge application practices , WWT P contaminant removal, or seasona l 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 valu es 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. S ediment 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.



## C ONCLUSIONS OF 

## E NVIRONMENTAL 

## A SSESSMENT

Th is assessment evaluated the environmental risk to the aquatic and terrestrial environment s 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 biosolid s 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, sediment

dwelling species, terrestrial species, and plants via contact with contaminated water, sediment, or soil.

The a cute and chronic risk for aquatic species did not exceed the COCs at m aximum measured surface water concentrations. The chronic COC was exceeded for sediment

- dwelling species at the upper range of measur ed environmental concentrations for l ess than 1.5% of the more than 600 sediment measurements from the USGS. T he data for c hronic risk for terrestrial invertebrates and plants did not exceed the COC at maximum measured soil concentrations ; this conclusion was based on limited data

.

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Page 43 

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- 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

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. Accessed June 15, 2012.

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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),
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- 990.

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# Appendi **ces**

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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 docume nts for more detailed information.



## A

- 1. H UMAN 

## H AZARD 

## C HARACTERIZATION

Toxicokinetics (Absorption, Distribution, Metabolism, Excretion) Available toxicokinetics data indicate that HHCB has poor dermal absorption and is extensively metabolized and excret ed, with no evidence of significant bioaccumulation. An in vivo human study reported approximately 0.1 percent dermal absorption following exposure to HHCB ( Hawk ins 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, kidn eys, 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 th e 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 to xicity 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 LD 50 

### values >3,000 mg/kg

- bw.

Subchronic/Repeated

- Dose Toxici ty 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] ( Ap i 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

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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 reporte d 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 the y used fragrance mixtures in which HHCB was only present at low levels

.

Repro ductive 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 a ny 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 nursin g. 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 breas t 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. F 1 

### offspri ng were exposed to
HHCB in utero from GD 14 through lactation via the dam’s milk, and exposures ceased after weaning in the F 1 

### offspring and in the F 2 

### 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 per iod, 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 F 1 

### generation offspring reached 84 days of age, they were mated (avoiding sibling pairings) and the reproductive capacity was assessed by examining the f emales before and after mating to determine time of pregnancy, estrous cyclicity, pre

- coital time, pregnancy rates, and duration of gestation. F 2 

### generation offspring were examined for abnormalities at parturition and then periodically until study termina tion 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

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Page 63 NOAEL for maternal, reproductive, and developmental toxici ty 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 Sprague

Dawley 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 lute a, implantation sites, resorption sites, live and dead fetuses, and live pup body weights were recorded. F etuses were examined for sex and underwent a gross examination. One half of the fetuses was examined for visceral anomalies and the other half was e xamined 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 significan ce 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 perio d (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 reduction s 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 wer e 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 fo r 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 LOAE L for developmental toxicity of 500 mg/kg

- bw/day was concluded and the NOAEL was 150 mg/kg

bw/day.

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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 sens itization 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 possib le photo

irritant 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, 1 987 ; 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 activi ty 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 E C, 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 (E 2 

### ). 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.

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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
LD 50 (mg/kg

- bw) >3,000 Minne r and Foster (1977 ); Moreno (1975 )

Acute dermal toxicity
LD 50 (mg/kg

- bw) >3,250 Minner and Foster (1977 ); Moreno (1975 )

Acute inhalation toxicity
LC 50 (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 ) ; Mersch

Sundermann 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 )

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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

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Page 67 

## A

- 2. H UMAN 

## B IOMONITORING

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.

Seve ral 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 r eported 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 concentrati on 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 adi pose 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) ( Schia vone 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 indi cated 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 i n 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 population s . 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.

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- 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 samp les, Siena, Italy, 3 females,
9 males 2005

- 2006 HHCB dete cted in
92 percent of adipose tissue samples Mean: 361 ± 46 7 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 μ g/kg fat Range: 38

- 422 μ g/kg fat
Mean: 179 ± 111 μ g/kg Large variability in individual samples Duedahl

- Olesen e t 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; bre ast 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

- 0.51 5 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 )

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- 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 fr om 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, Shangha i, 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 m aternal 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 s amples (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 = 2 04 (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 )

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## A

- 3. S UMMARY OF 

## 2008 EU H UMAN 

## H EALTH 

## R ISK 

## A SSESSMENT

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 well

documented, 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 t he
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 percen t 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 inhalat ion exposures were used.

For general systemic effects following repeated exposures, the oral NOAEL of 150 mg/kg

bw/day from the 13

- week subchronic toxicity study in rats was used as a point of departure. Since NOAELs are not available following dermal o r 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).

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### 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 LD 50 

### 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, co mbined 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 expec ted.

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). C ombined 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. A lthough 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 o f 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 wor kers 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.

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### 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 w as no concern for local effects in consumers following repeated exposures. S ince there is a lack of skin and eye irritation potential, no significant respiratory tract irritation potential was expected.

For general systemic effects following repeate d 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). Ther efore, 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 consum ers, 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 meas ures 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 fis h 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 g eneral systemic effects following repeated exposures, an internal NOAEL of ≥75 mg/kg

bw/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) w ere 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

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### regional exposure scenarios, there was no concern for genera l 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
in formation 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 mutagenici ty 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 valu es 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 vi a 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 consumpt ion 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 exposur es for consumers; and (3) oral and

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### 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 differenc es, and a factor of 2 for semichronic to chronic exposure extrapolation). The calculated MOS was well above 100 (798). F or the developmental/ reproductive toxicity endpoint, a MOS of 50 (based on 4*2.5 for interspecies species differences and 5 for intra species 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 r isk reduction measures beyond those already being applied.



## A

- 4. K EY 

## S OURCES OF 

## U NCERTAINTY AND 

## D ATA 

## L IMITATIONS ON 

## H UMAN 

## H EALTH

Overall, adequate screening

- level animal toxicity data are available to characterize the human health hazard for HHCB. Toxico kinetics (by the dermal route), acute, repeated

- dose, and developmental toxicity data are available to characterize the human health hazard of HHCB. Although n o multigenerational reproductive toxicity studies on HHCB are available, information on developm ental and reproductive toxicity was obtained from both the repeated

- dose dietary toxicity study (reproductive organ data) and the p eri/post

- natal reproductive toxicity study with modified exposures. Several assays testing for endocrine disruption, genotox icity, and irritation/sensitization (including several studies in humans) are also available. The database is incomplete for t oxicokinetic data in animals by the oral and inhalation routes; acute toxicity data by the inhalation route; repeated

- dose toxici ty 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 st udy 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 stud y authors to be

treatment

- related. Additionally, fetal malformations observed were reported in on ly 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. Exposu res in the F 1 

### offspring occurred in utero from GD 14 through lactation; there were no

e xposures to HHCB beginning from weaning in the F 1 

### offspring throughout the F 2 

### 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

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### postnatal death) and reproductive toxicity (reproductive performance) as well as a battery of developmental neurotoxicity tests. Other rel iable entities such as the 2008 EU RAR on HHCB ( EC, 2008 ) and the 2009 SIDS Initial Assessment Profile ( OECD, 2009 ) als o 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 represe ntative. 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 va ry

.



## A

- 5. C ONCLUSIONS OF 

## H UMAN 

## H EALTH 

## A SSESSMENT

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 ey e 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. Howeve r, 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/kg

bw/day) lead to the conclusion that 500 mg/kg

- bw/day is a conservative LOAEL. N o effects were reported in fetuses in the perinatal toxicity study with lactational exposures at doses

s imilar to those reported in human milk and several orders of magnitude higher (20 mg/kg

bw/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.

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### The review of human health hazard studies, biomonitoring studies in the US and elsewhere, and the EU RAR showed an overall l ow risk concern for human health, including the risk for developmental toxicity, from current use of HHCB.

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## 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 t hen 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

- (

- hydroxy isopropyl)indane (2) to 1,3,4,6,7,8

- hexahydro

- 4,6,6,7,8,8

- hexamethylcy clopenta

- 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: Wiley

VCH 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.

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## Appendix **C :**  **HHCB Di a**  **stereoisomers (#1 to**  **#6) and**  **Related S**  **tructural Analogs (#7 to #15)** 

Table B

- 1. HHCB Diastereoisomers and Related Structural Analogs.

CASRN Chemicals Structure Chemical Index Name On TSCA Inventory

1 1723 39

- 62

- 7 Cyclopenta [ g]

- 2

- benzopyran,1,3,4,6,7,8

- hexahydro

- 4,6,6,7,8,8

- hexamethyl

- , (4S,7S)

- N

2 1723 39

- 63

- 8 Cyclopenta [ g]

- 2

- benzopyran,1,3,4,6,7,8

- hexahydro

- 4,6,6,7,8,8

- hexamethyl

- , (4R,7S)

- N

3 2523 32

- 95

- 9 Cyclopenta [ g]

- 2

- benzopyran,1,3,4,6,7,8

- hexahydro

- 4,6,6,7,8,8

- hexamethyl

- , (4S,7R)

- N

4 2523 32

- 96

- 0 Cyclopenta [ g]

- 2

- benzopyran,1,3,4,6,7,8

- hexahydro

- 4,6,6,7,8,8

- hexamethyl

- , (4R,7R)

- N

5 2529 33

- 48

- 5 Cyclopenta [ g]

- 2

- benzopyran,1,3,4,6,7,8

- hexahydro

- 4,6,6,7,8,8

- hexamethyl

- , (4R,7R)

- rel

N

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- 1. HHCB Diastereoisomers and Related Structural Analogs.

CASRN Chemicals Structure Chemical Index Name On TSCA Inventory

6 2529 33

- 49

- 6 Cyclopenta [ g]

- 2

- benzopyran,1,3,4,6,7,8

- hexahydro

- 4,6,6,7,8,8

- hexamethyl

- , (4R,7S) rel

N

7 1222

- 06

- 6 Cyclopenta [ g]

- 2

- benzopyran, 1,3,4,6,7,8

- hexahydro

- 4,4,6,6,8,8

- hexamethyl

N

8 857091

- 61

- 3 Cyclopenta [ g]

- 2

- benzopyran, 1,3,4,6,7,8

- hexahydro

- 3,6,6,7,8,8

- hexamethyl

N

9 1022 96

- 64

- 0 Cyclopenta [ g]

- 1

- benzopyran, 2,3,4,6,7,8

- hexahydro

- 4,4,6,6,8,8

- hexamethyl

N

10 1355 46

- 43

- 9 Cyclopenta [ g]

- 1

- benzopyran,2,3,4,6,7,8hexahydro

4,6,6,7,8,8

- hexamethyl

N

11 1355 46

- 42

- 8 Cyclopenta [ g]

- 2

- benzopyran, 1,3,4,6,7,8

- hexahydro

- 1,6,6,7,8,8

- hexamethyl

N

12 1141 09

- 63

- 6 Cyclopenta [ h]

- 2

- benzopyran, 1,3,4,7,8,9

- hexahydro

- 4,7,7,8,9,9

- hexamethyl

N

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- 1. HHCB Diastereoisomers and Related Structural Analogs.

CASRN Chemicals Structure Chemical Index Name On TSCA Inventory

13 1141 09

- 62

- 5 Cyclopenta [ f] [ 2]benzopyran, 1,2,4,7,8,9

- hexahydro

- 1,7,7,8,9,9

- hexamethyl

N

14 78448

- 48

- 3 Cyclopenta [ g]

- 2

- benzopyran, 6

- ethyl

1,3,4,6,7,8

- hexahydro

- 4,6,8,8

- tetramethyl

N

15 78448

- 49

- 4 Cyclopenta [ g]

- 2

- benzopyran, 8

- ethyl

1,3,4,6,7,8

- hexahydro

- 4,6,6,8

- tetramethyl

N

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## 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

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- 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

.

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## 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 c ycle of a fragrance ingredient such as HHCB includes :

• Manufactu ring 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. The se 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 result ed in the highest per

- site release due to a relatively low number of domestic compounding sites. A discussion of the release estimates acc ording to stage s 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 a ssessment, 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

.

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### 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/Site

Day) 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

- 18 a 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

- 251 a 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

- 0.2 b 5

- 49 2

- 18 a 13.2

- 37.8 c 1,190

- 3,400 d EPA Bulk Transport Residual Model ( US EPA, 1988 ; 1992 )
Cleaning compounding equipment Water 0.07

- 1 b 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 spray

drying 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

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### 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/Site

Day) 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 throug hput per site by the container volume, assuming that one cont ainer is unloaded per site per day. Bulk containers are assumed to be 5,000 gallons and totes are assumed to be 550 gallons. The cal culation for the cleaning of bulk containers is presented as a sample calculation of this variable:
Low

- end of range of n umber of unloading days (or release days) per site: HHCB throughput per site = (1,700,000 kg/year ÷ 49 sites) ÷ (1 g/cm 3 × kg/1,000 g × cm 3 /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. Upper High

- 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/cm 3 × kg/1,000 g × cm 3 /0.0002642 gallons) = 89,828 gallons per site.
Numbe r 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 equipm ent. 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/cm 3 × kg/1,000 g × cm 3 /0.0002642 gallons) = 13.2 kg/site

- day Upper High

- end of range of amount released per site

- day = 5,000 gallons/container

- site

- day × 0.002 residual gallons/container gallon × (1 g/cm 3 × kg/1 ,000 g × cm 3 /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 variabl e: 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 (high

end) and 0.3 percent (central tendency).

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### Estimated Release from Compounding The number of compounding sites in which HHCB is pro cessed 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. Compoun ding sites are classified under North American Industry Classification System ( NAICS ) code 3 2562

, 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, wh ich represent the vast majority of fragrance volume produced in the US. IFRA North America currently has 49 member companies ( IFR A, 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 3 2562 by subcategory, but shows that the total number of companies within NAICS 3 2562 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, c ompounding 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 compound ed 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 3 2562 are small companies with no more than one establishment, the 34 companies are assumed to represent 34 establishments (sites).

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Page 87 Table E

- 2

. Geographic Distribution for Facilities under NAICS 3 2562 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

.

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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 HHC B

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 maxim um 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 blendi ng 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 3 2562

, 3 2561 1, and 3 2561 2. A geographic distribu tion 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 3 2561 1 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

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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 3 2561 2 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

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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 container s 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 fragr ance oils ; therefore, these releases we re 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 powder ed products is not known, EPA conservatively assumes that the entire volume is formulated into powder products, wh ich 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

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### 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 Rele ase 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 rele ase 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 “fin e 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. Th ese 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 t reatment 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 re leases from all sites on an annual basis.

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## Appendix **F :**  **Environmental Fate and Bioaccumulation Study**  **Summaries**

ENVIRONMENTAL PERSISTENCE Fate in Wastewater Treatment

In the Langworthy et al

. , 2000 study ( as report ed in EC , 2008)

, HHCB disappeared with the subsequent appear ance 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 o riginal radioactivity

. Graduall y 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 fo r HHCB hydroxycarboxylic acid. A third, highly polar metabolite constit uted 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),

m ore 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 me tabolites were formed at different rates and with increasing polarity over time ( Schaefer et al

., 2005; as reported in E C , 2008 )

. In the river die

- away test the total radioactivity of metabolites rose from 25 to 56 percent between day s 2 and 7 and t o nearly 62 percent at 28 days. When the results we re 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 t he radioactivity was lost by volatilization at 300 h ou r s , but some of the volatilized radioactivity could have been in the form of degradation products, since degradation in water wa s also happening during the test

.

Fate in Soil and Sediment
In Envirogen

, 1998 ( as reported in EC , 2008)

, s ealed flasks with HHCB

- spiked soil at 10 μg/g soil (10 mg/kg) were incubated at laboratory amb ient 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 a gricultural soil, respectively. The TLC results showed that HHCB was degraded to vario us more polar metabolites. It was hypothesized that the majority of unrecovered radiolabel became covalently bound to soil organ ic compounds ( i.e.

, was immobilized by humification).

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### As reported by DiFrancesco et al

. , 2004 ( dissipation of fragrance materials in sludge

- amended soils ), a naerobically 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
d w in the year s 200 0 and 2002, respectively. For Wilmington the concentrations were 43 a nd 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 /m 2 

### of sludge. Each of the four soils was incubated with digested sludge, both unspiked a nd spiked with a mix ture of fragrance ingredients. The soil and sludge were mixed over 30 minutes in a cement mixer to e nsure 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 ext r acted 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 initi al amount in the spiked soils. HHCB was not detected in lea chate from unspiked soils (EC, 2008)

. The low percentage of initial HHCB present in leachate wa s probably a result of both degradation and sorption.

BIO ACCUMULATION 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 bioa ccumulation potential (EPA, 1999a; 1999b; Boethling et al ., 2009) . Measured BCFs of 620, 862 and 1,584 were reported for zebra fish, eel and bluegi lls, respectively (Table 2

- 10), whereas BAF values of 290, 510, 580

, and 620 were reported for eel, tench, crucian carp and zebra mussels, respective ly. 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 organi sms than in their food (prey). Metabolism may account for the observation that the relatively low BCFs and BAFs a ppear in consistent with the high measured log K OW 

### of HHCB (5.9).

The bluegill study was performed using G ood L aboratory P ractice (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 concen trations 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 T LC/HPLC. Nominal exposure concentrations were 1 and 10 μg/ L . The fish were exposed for 28 days; the elimin ation period was also 28 days. The BCF was derived from actual concent rations of parent compound in exposure water and t he plateau level in fish. The elimination rate constant (k 2 

### ) was estimated from the elimination curve (first

- order kinetics), and determination of an uptake rate constant (k 1 

### ) was attempted using the follo wing formula : k 1 

### = BCF × k 2 

### . However, this

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### 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 d ays of exposure. Elimination followed first

- order k inetics with a half

- life of 2 to 3 days for HHCB, allowing calculation of k 2 

### . 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 relativel y short time and the n excreted to the water phase. E ven if HHCB is metabolized rapidly , bioaccumulation can still be significant. In BCF tests

, metabolism can occ ur during exposure of fish to the test substance; therefore the measured BCF reflects a bal ance 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 t he substance from reaching a concentration in the fish that is 1,584 times greater than in the exposure medium.

Bioaccumulation has been studied with tw o benthic species (Table 2

- 11). The aqueous
concentrations in the test with non

- biting midges ( Chiro nomus 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 day s 1 and 3

, and then the level rapidly d ecreased 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 we re likely caused by relatively fast biotransformation of HHCB in this organism

. In a flo w

- 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 d ays , 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

BCF worm 

### = (0.84 + 0.012 × K ow 

### )/RHO earthworm



### For RHO earthworm 

### (bulk density of worm) by default a value of 1 (kg wet weight / L ) may be assumed. This leads to a predicted BCF worm 

### of 2,395 L /kg.

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## 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

m aximum 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

sam pling. These results

, in part, reflect the high production volume of HHCB ( US EPA, 2012d )

. The i nfluent 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 ef fl uent was published in 2012 ( Hope et al. , 2012 ) . Oregon's Senate Bill 737, enacted in 2007 , required the state's 52 largest municipal WWTP s and water pollution control facilities to collect ef fl uent 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 fl ow 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 val ue 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 betwee n 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

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### of moderate population s ize (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, Minnesot a 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 popula tions. 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 conf luence 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 musk s. 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 one

year 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

, al though the authors note d that process temperature ma y have influence d 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 les s than the influent HHCB concentrations, as summarized in Table G

- 1.

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- 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 med ian ; 13,000 max imum (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 Taylor s Falls WWTP St. Croix Falls WWTP USGS (2011 )

Ontario,
C anada 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 L agoon E xtended aeration
O xidation ditch
A ctivated sludge
A ctivated sludge
A ctivated sludge Smyth et al. (2008 )

a Not reported.

b Not measured.

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### 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 inheren t to the WWTP

, such as WWTP size and type (municipal or industrial), processes utilized to treat waste, and populations served (rural v er s us 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 )

. Alt hough 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 eme rging concern, evaluate the ir 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 effluent

dominated 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 m aximum 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 Nor th 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 differ ences in concentrations of HHCB in these two locations

, indicating minimal in

- stream dilution or transformation. The mean concentration w as reported as 1.6 μ g/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 v icinity 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 sur face waters of the Canyon Lakes System and North Fork Brazos River near Lubbock

, Texas in 2010 ( Chase et al. , 2012 ) . Reported concentration s in the lake ranged from ND to trace amounts, w hereas concentrations in the river were appreciably higher

, ranging from 77 to 794 ng/L. The lake concentrations were expected to be low as the

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### groundwater from a land applied site where the source was f rom a lake

, whereas occurrence in the stream water was attributed to direct release of WWTP effluent.

Peck et al. (2006 ) characte rized 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 concentration s for HHCB ranging from 3.95 to 25.8 ng/L ( Rei ner and Kannan, 2011 ) . The highest concentration was found near Albany, N ew Y ork, 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 in dicators in the St. Croix National Scenic Riverway in Minnesota and Wiscon sin ( USGS, 2008 )

. Samples of treated wastewater effluent from two WWTPs , located in Taylor s 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 influe nced 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 conclude d that further investigation is needed.

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Page 100 Table G

- 2 . Measured Concentrations of HHCB in Surface Water

.

Location Date n Concentration (ng/L) Comments Reference

CA 2008 970 max imum 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 ) Taylor s Falls, MN St. Croix Falls, WI Franconia, MN Sunrise, MN 2007

- 200 8 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 95 th percentile Great lakes watershed, summary Klecka et al. (201 0 )

Lake Ontario,
C anada June 2008

— a 0.12

- 0.34

est imated
0.21

- 10.1

est imated Port Hope Pickering

- Ajax Helm et al. (2012 )

Lake Ontario,
C anada 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.

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### Chase et al. (2012 ) measured the concentration of HHCB in the sediment of a playa lake, a non

effluent impacted study site located in proximity to an area treated by a WWT P. 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 concentration s of HHCB w ere 2.13 and 1.43 ng/g dw in the summer and fall of 2010

, respectively, as shown in Table G

- 3

.

The US GS 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 μ g/kg dw with an estimated median value of 60 μ g/kg dw . The reported upstream concen trations were < 60 μ g/kg dw

, whereas the downstream concentrations ranged from <30 to 390 μ g/kg dw with three of the seven downstream samples having measured values > 200 μ g/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

, N ew Y ork ( Reiner and Kannan, 2011 )

. WWTPs in each of the se New York cities discharge effluents into the Hudson R iver. 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 note d 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 200 7 ( 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 d etails regarding the number of samples for each location were not provided.

Sediment samples from Lake Erie and Lake Ontario were measured in Aug ust 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 concentra tion 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

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### 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 P laya Chase et al. (2012 )

Tinkers Creek, OH Cuyahoga River, OH 2006 20 a

- <60, 37 av erage 20 a

- 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 200 7 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 K OW 

### values of 5.3 to 5.9 ( s ee Table 2

- 1 ) and the estimated K OC 

### values of 4.1 to 4.3 (E PI Suite v4.10 , see Chapter 2, S ection B

, Environmental Fate ). These solids, produced through the process of w astewater 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 uptak e.

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 seve n 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. HHC B 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

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### sludge and chlorin e 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 μ g/kg dw . All other measured HHCB concentrations were below 4,000 μ g/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 va lue for HHCB was 1,461 μ g/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 concentration s measured, with no one factor being clearly delineated in this study.

A second study by Kinney et al. (2008 ) was perfor med 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 μ g/kg dw, although details regarding this sou rce 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 fr om the return activated sludge line after purging the stagnant sludge from the sampling line

. T he presence of HHCB was determined using
GC / MS following extraction with a Soxhlet apparatus for 24 hours. The average concentration of HHCB in biosolids w as r eported to be 7.84 mg/kg dw (7,840 μ g/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 W ilmington 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 μ g/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 ex tended 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 usi ng supercritical fluid or microwave assisted techniques, measurement was performed using GC / MS. HHCB concentrations in digested biosolid ranged

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### from 9,430 to 55,500 μ g/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 ( Stev ens et al. , 2003 )

.

Yang and Metcalfe (2006 ) also studied biosolids from WWTPs in Ontario, Canada. Samples were collected from the Pete rborough WWTP over a one

- year period

, and the HHCB
concentration s 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

, the re 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
concentration s of HHCB w ere 3,309.9 μ g/kg dw in the return activated sludge and 6,788.4 μ g/kg dw i n the digested biosolid.

Table G

- 4 . Measured Concentrations of HHCB in Biosolids

.

Location Collection
Year n Concentration

μ g/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

U nknown 2008 3 427,000 b 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 43,000 b
21,800 (± 4,300) Anaerobically digested sludge DiFrancesco et al. (2004 )

Georgetown,
DE 7/ 2000 3/ 2002 — a

3 86,000 b
37,600 (± 4,500) Anaerobically digested sludge

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Page 105 Table G

- 4 . Measured Concentrations of HHCB in Biosolids

.

Location Collection
Year n Concentration

μ g/kg dw Preparation/Comments Reference

Ontario,
C anada (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 )

C anada (site 1/site 2) 2003

- 2005 — a 36,500/20,100 25,600/ 17

, 50 0 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 )

C anada Peterborough,
ON 2/ 2003

4/ 2003

7/ 2003

10/ 2003

Annual

a v era g e 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 N ot 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
HH CB reported concentrations in the range of < 100 to > 100,000 μ g/kg dw . T he 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.

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### Measured Concentrations in Soil Limited dat a 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 amendm ent of soils with biosolids resulted in measurable concentrations of HHCB. The high frequency of HHCB detection in biosolids and wastewater suggest s that they may be an important organic waste contaminant source to terrestrial environments, particularly w hen 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, T exas 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 thr ough 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 difference s in measured HHCB concentration s 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 loca tions 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

, excep ting 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 agricultur al soil amended with biosolid s . 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 rec eiving 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 μ g/kg dw. HHCB concentrations were highest at site 2 and ranged from 1,050 to 2,770 μ g/kg dw. Table G

- 5 summarizes the measured concentrations in HHCB in soil.

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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 in ches 12

- 18 in ches 24

- 30 in ches

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, C anada 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, degradati on mechanisms (such as volatilization, transformation

, and leaching)

, and organic matter content on long

- term concentrations. Soil studies, al though clearly limited, do indicate that land application of treated wastewater effluent or biosolids results in detectable quantities of HHCB in soil and hence represent s a potential route through which HHCB may enter terrestrial ecosystem s

.

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 e ach 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 rangin g from 300 to 2,100 ng/g tissue.

C oncentrations 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

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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 concentra tions

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 syntheti c musks. HHCB was detected in all samples

, with concentrations ranging from 48 to 683 ng/g lipid (av era g e 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, ind icates the widespread exposure of aquatic biota to HHCB.

Kannan et al. (2005 ) reported the presence of HHCB in the t issues of higher trophic level aquatic organisms and humans. Among the wildlife species analyzed, tissue concentrations ranged from <1 to 25 ng/g ww . B lubber tissue obtained from spinner and bottlenose dolphins from coastal waters off the state of Florid a contained the highest levels of HHCB.

Seven to eight carp were collected 200 meters from the drinking water intake area on Lake Mead, N evada 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 concentration s in earthworm tissue obtained from amended soils where biosolids or liquid swine man ure 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 μ g/kg ww ) were an order of magnitude higher than the levels measured after 156 days (131 μ g/kg ww ). Tissue concentrations in worms from the manure amended site were 49 μ g/kg after 139 days.

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- 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 p erch (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 c atfish (liver) 3
1 Troy, NY Catskill, NY 11.1

- 39
<1

Smallmouth b ass (liver) 3
3
3 Troy, NY Albany, NY Catskill, NY <1

- 11.1
1

- 31.9

<1 Largemouth b ass (liver) 1
1 Albany, NY Catskill, NY 10.9
8.22

White c atfish (liver) 1
1 Albany, NY Catskill, NY 6.56
5.79

Brown b ullhead (liver) 3 Catskill, NY <1

- 51.1

Zebra m ussel 4 Catskill, NY 10.3

- 19.3

American e el (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

conc entration)

Earthworm
3
3
3 Midwestern US

- minimally affected
site

- biosolid amended site

- manure amended site 2005
61 a ; ND b
3,340 c ; 131 d

ND e ; 49 f Kinney et al. (2008 )

Largemouth b ass 6 Chicago, IL North Shore Channel 2006 1,300 mean; 1,800

max Ramirez et al. (2009 )

Smallmouth b uffalo 6 Dallas, TX Trinity River 800 mean; 1,800 max

Bowfin 6 Orlando, FL Little Econlockhatchee
River 100 mean; 300 max

Common c arp 6 Phoenix, AZ Salt River 1,800 mean; 2,100

max
White s ucker 6 West Chester, PA Taylor Run 1,800 mean; 2,000

max

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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 o tter (liver) 8 Monterey Bay, CA 1993

- 1999 <1

- 3. 2 g Kannan et al. (2005 ) Harbor s eal (liver) 3 Central CA coast 1996

- 1997 4.4

- 5.5; 4.8 mean
California s ea l ion (liver) 3 Central CA coast 1993

- 1996 1.5

- 4.4; 2.8 mean
River o tter (liver) 3 Grand River, MI 1997 2.4

- 3; 2.8 mean
Bottlenose d olphin (blubber) 4 FL coast 1994

- 2000 4.2

- 20.5; 12 mean

Striped d olphin (blubber) 3 FL coast 1995

- 1997 8.1

- 25; 14 mean Pygmy s perm w hale (blubber) 1 FL coast 2000 6.6

Atlantic s harpnose s hark (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 m erganser (liver) 2 Buffalo Harbor, NY 1999 3.7

- 4.2; 4 mean

Greater and l esser s caup (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 s almon (skin on fillet) 6 Farmed and wild, local market NY 2003 <1

- 3. 2 h

Smallmouth b ass(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.

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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 ( US GS, 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.

T he 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 Childr ess 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., 6 2075 , 6 2823

, and 63209) for HHCB were included. The LRL for water sampling was 0.5 μ g/L for sampling dated 7/16/ 2001 to 9/30/ 2009 . The LRL was updated to 0.05 μ g/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, 1 st 

### (Q1) and 3 rd 

### (Q3) quartile, and 5 th 

### and 95 th 

### percentile values as shown in Figures G

- 1

to G

- 3. A summary of these calculated values is presented in Table G

- 7

.

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Page 112 Table G

- 7 . Summary of Box Plots for USGS HHCB Data

.

USGS Parameter Code Medium/Site HHCB Concentration

5 th

P ercentile Q1 Median Mean Q3 95 th

P ercent ile

6 2075 Water, f iltered, r ecoverable ; HHCB ( μ g/L)
Effluent/ s tream 0.35 0.35 1.40 1.18 1.82 2.16 Effluent/ o utfall 0.09 0.35 0.76 0.98 1.2 3.40

Surface w ater/ o utfall 0.35 0.35 0.65 1.08 1.70 2.30

Surface w ater/ s tream 0.10 0.35 0.35 0.35 0.35 0.35

Surface w ater/ l ake, reservoir, impoundment 0.04 0.35 0.35 0.33 0.35 0.35 Groundwater/ w ell 0.04 0.35 0.35 0.32 0.35 0.35

6 2823 Water, u nfiltered, r ecoverable ; HHCB ( μ g/L)
Effluent/ o utfall 0.14 0.14 1.02 1.01 1.85 2.26

Surface w ater/ s tream 0.02 0.05 0.14 0.12 0.14 0.15

Surface w ater/ l ake, reservoir, impoundment 0.03 0.14 0.14 0.12 0.14 0.14

63209 Solids, recoverable, dry weight ; HHCB ( μ g/kg)

Bottom m aterial/ l ake, reservoir, impoundment 27.93 41.02 65.42 87.46 106.08 212.87

Bottom m aterial/ s tream 17.68 24.75 40.00 67.99 67.19 200.00

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Page 113 Table G

- 8. USGS Medium and Site Codes.

Medium Codes

Surface w ater 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 m ater ial 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 o il 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 o f 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, al though the range of values v aried from ND to > 3 μ g/L. The mean effluent concentrations were comparable

, but the range of concentrations was greate r 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 μ g/L for all medium and site types. Groundwater

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Page 114 

### concentrations at well sites and surface water concentration s at streams and lakes, reservoirs

, and impoundment sites ranged from ND to < 0.5 μ g/L. For effluent/outfall filtered/non

- filtered samples, the HHCB concentration s were higher in unfiltered samples (Figure G

- 2)

. This may be attributed to the presence of H HCB in suspended solids. Again, significantly lower concentrations were consistently recorded in surface water samples collected from streams and lake s , reservoir s, and impoundment sites

— areas that are unaffected or nominally affected by wastewater treatm ent outfall.

Figure G

- 1. Monitoring D ata S ummary from USGS NWIS for HHCB C oncentrations in W ater, F iltered.

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Page 115 Figure G

- 2. Monitoring D ata S ummary from USGS NWIS for HHCB C oncentrations in W ater, U nfiltered.

Figure G

- 3. Monitoring D ata S ummary from USGS NWIS for HHCB C oncentrations in S olids.



### 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/impou ndment and streams (Figure G

- 3)

. The 95 th 

### percentile concentration range at lake /reservoir/impoundment sites was 213 μ g/kg

, and the mean concentration was
87 μ g/kg. The 95 th 

### percentile concentration range at stream sites was 200 μ g/kg

, and the mean
concentration was 68 μ g/kg. The range of data was similar for both site types, with the Q3

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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.