EPA Document# 746-R1-4001 August 2014 Office of Chemical Safety and
Pollution Prevention



## TSCA Work Plan Chemical Risk Assessment

HHCB

1,3,4,6,7,8-Hexahydro-4,6,6,7,8,8-hexamethylcyclope nta- γ -2- benzopyran

CASRN: 1222-05-5



# C H 3 

# CH 3



# CH 3 

# CH 3 

# C H 3



# O CH 3

August 2014

## TABLE OF CONTENTS

TABLE OF CONTENTS .................................

................................................... ......................... 2

LIST OF TABLES ....................................

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

LIST OF APPENDIX TABLES ...........................

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

LIST OF FIGURES ...................................

................................................... .............................. 7

LIST OF APPENDIX FIGURES ..........................

................................................... ....................... 7

AUTHORS/CONTRIBUTORS/ACKNOWLEDGEMENTS/REVIEWERS.... ........................................ 8

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

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

EXECUTIVE SUMMARY .................................

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

1 BACKGROUND AND SCOPE ..............................

................................................... .......... 14

1.1 INTRODUCTION ......................................

................................................... ................................................. 1 4

1.2 PROBLEM FORMULATION .......................................

................................................... ................................ 16

1.3 CONCEPTUAL MODEL FOR ENVIRONMENTAL ASSESSMENT ........................................ .............................. 17

1.4 ANALYSIS PLAN FOR ENVIRONMENTAL ASSESSMENT ........................................ ........................................ 19

2 SOURCES AND ENVIRONMENTAL FATE ....................

................................................... .. 20

2.1 INTRODUCTION ......................................

................................................... ................................................. 2 0

**2.1.1 Physical and Chemical Properties ..................** 

**................................................... ................................. 20** 

2.2 PRODUCTION AND USES ..............................................

................................................... ............................ 21

**2.2.1 Production ........................................** 

**................................................... ............................................... 22** 

**2.2.2 Uses ..............................................** 

**................................................... ...................................................**  **24**

**2.2.3 Conclusions of Production and Use .................** 

**................................................... ................................ 26** 

2.3 ENVIRONMENTAL FATE ..............................................

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

**2.3.1 Environmental Persistence .........................** 

**................................................... ..................................... 26** 

2.3.1.1 Fate in Wastewater Treatment ......................

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

2.3.1.2 Fate in Water .....................................

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

2.3.1.3 Fate in Soil and Sediment .........................

................................................... ................................................... .. 28

2.3.1.4 Fate in Air .......................................

................................................... ................................................... ............ 30
**2.3.2 Bioaccumulation and Bioconcentration ..............** 

**................................................... ............................ 30** 

**2.3.3 Conclusions of Environmental Fate .................** 

**................................................... ................................ 33** 

3 ENVIRONMENTAL ASSESSMENT ..........................

................................................... ....... 34

3.1 INTRODUCTION ......................................

................................................... ................................................. 3 4

3.2 ENVIRONMENTAL EXPOSURE ASSESSMENT ........................................

................................................... .... 34

**3.2.1 Estimated Environmental Releases ..................** 

**................................................... ............................... 34** 

**3.2.2 Measured Levels in the Environment ................** 

**................................................... .............................. 34** 

3.2.2.1 Wastewater ........................................

................................................... ................................................... ........ 37

3.2.2.2 Surface Water .....................................

................................................... ................................................... ........ 37

3.2.2.3 Sediment ..........................................

................................................... ................................................... .......... 38

3.2.2.4 Biosolids and Soil ................................

................................................... ................................................... ........ 38

3.2.2.5 Biota .............................................

................................................... ................................................... .............. 39

3.2.2.6 USGS Data Analysis ................................

................................................... ................................................... ..... 39
**3.2.3 Conclusions of Environmental Exposure .............** 

**................................................... ............................ 40** 

3.3 ECOLOGICAL HAZARD ASSESSMENT ........................................

................................................... ................ 40**3.3.1 Acute Toxicity to Aquatic Organisms ...............** 

**................................................... ............................... 41** 

**3.3.2 Chronic Toxicity to Aquatic Organisms .............** 

**................................................... .............................. 42** 

**3.3.3 Toxicity to Sediment-Dwelling Organisms ...........** 

**................................................... ........................... 45** 

**3.3.4 Toxicity to Terrestrial Organisms .................** 

**................................................... ................................... 46** 

**3.3.5 Conclusions of Environmental Hazard Assessment ....** 

**................................................... ..................... 49** 

3.4 ENVIRONMENTAL RISK CHARACTERIZATION ..................................

................................................... ........ 49

**3.4.1 Calculation of Risk Quotient (RQ) Values ...........** 

**................................................... ............................. 49** 

**3.4.2 Key Sources of Uncertainty and Data Limitations ...** 

**................................................... ........................ 52** 

3.4.2.1 Representativeness of Exposure Concentrations .....

................................................... ..................................... 52

3.4.2.2 Variability in Environmental Concentrations .......

................................................... .......................................... 52

3.4.2.3 Anaerobic Degradation..............................

................................................... ................................................... . 53

3.4.2.4 Volatilization ....................................

................................................... ................................................... ........... 53

3.4.2.5 Isomers and Metabolites ...........................

................................................... ................................................... . 53

3.4.2.6 Deriving Concentrations of Concern from Single Spec ies Tests ......................................... .............................. 53

3.4.2.7 Assessment of Risk to Terrestrial Invertebrates or Plants ............................................ .................................... 54
**3.4.3 Conclusions of Risk Characterization ..............** 

**................................................... ................................ 54** 

3.5 CONCLUSIONS OF ENVIRONMENTAL ASSESSMENT ........................................ ........................................... 54

REFERENCES ........................................

................................................... ............................. 55

APPENDICES ........................................

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

Appendix A HUMAN HEALTH TOXICITY STUDIES, BIOMONITORING DATA, AND RISK

ASSESSMENTS .......................................

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

A-1 H UMAN H AZARD C HARACTERIZATION

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

**A-1-1 Toxicokinetics (Absorption, Distribution, Metabolis**  **m, Excretion) ..................................... ................ 78** 

**A-1-2 Acute Toxicity ....................................** 

**................................................... .............................................. 78** 

**A-1-3 Subchronic/Repeated-Dose Toxicity .................** 

**................................................... ............................... 79** 

**A-1-4 Reproductive Toxicity and Fertility ...............** 

**................................................... ................................... 79** 

**A-1-5 Developmental Toxicity ............................** 

**................................................... ....................................... 80** 

**A-1-6 Genetic Toxicity ..................................** 

**................................................... ............................................. 81** 

**A-1-7 Carcinogenicity ...................................** 

**................................................... ............................................. 81** 

**A-1-8 Additional Information ............................** 

**................................................... ........................................ 81** 

A-2 H UMAN B IOMONITORING

.................................................. ................................................... .......................... 84

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

.................................................. ............................... 87

**A-3-1 Assumptions and Points of Departures Used in the EU**  **RAR .............................................. ................ 87** 

**A-3-2 Risk to Workers ...................................** 

**................................................... ............................................ 88** 

**A-3-3 Risk to Consumers .................................** 

**................................................... .......................................... 89** 

**A-3-4 Risk to Humans Exposed Indirectly via the Environme**  **nt ................................................ ................... 90** 

**A-3-5 Assessment of Risk for Breast-Fed Babies Exposed vi**  **a Mother’s Milk ................................... ............ 90** 

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

.................................................. ......... 91

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

.................................................. ............................................... 92
Appendix B CHEMICAL SYNTHESIS OF HHCB ........................ ............................................... 94

Appendix C HHCB (*), HHCB DIASTEREOISOMERS (#1 TO #6), AND REL ATED STRUCTURAL

ANALOGS (#7 TO #15) ...............................

................................................... .... 95Appendix D CDR DATA FOR HHCB .................................

................................................... ... 98

Appendix E MODELED RELEASE ESTIMATES ACCORDING TO STAGE OF PRO DUCTION AND

USE ...............................................

................................................... ............. 100

E-1 E STIMATED R ELEASE FROM M ANUFACTURE AND I MPORT

.................................................. .................................. 100

E-2 E STIMATED R ELEASE FROM C OMPOUNDING

.................................................. ................................................... 104

E-3 E STIMATED R ELEASE FROM B LENDING OF F RAGRANCE O ILS

.................................................. ............................... 106

E-4 E STIMATED R ELEASE FROM U SE OF C OMMERCIAL AND C ONSUMER P RODUCTS

.................................................. ...... 109
Appendix F ADDITIONAL STUDIES ................................

................................................... . 110

F-1 E NDOCRINE M ECHANISMS AND M OLECULAR P ATHWAYS

.................................................. .................................. 110
Appendix G ENVIRONMENTAL MONITORING DATA ANALYSIS ............ .............................. 111

G-1 M EASURED C ONCENTRATIONS IN W ASTEWATER

.................................................. ............................................. 111

G-2 M EASURED C ONCENTRATIONS IN S URFACE W ATER

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

G-3 M EASURED C ONCENTRATIONS IN S EDIMENT

.................................................. ................................................... 118

G-4 M EASURED C ONCENTRATIONS IN B IOSOLIDS AND S LUDGE

.................................................. ................................. 120

G-5 M EASURED C ONCENTRATIONS IN S OIL

.................................................. ................................................... ........ 124

G-6 M EASURED C ONCENTRATIONS IN B IOTA

.................................................. ................................................... ..... 126

G-7 USGS N ATIONAL W ATER Q UALITY I NFORMATION S YSTEM D ATA

.................................................. ....................... 131

## LIST OF TABLES

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

.................................................. .................... 21 Table 2-2. Major US Manufacturers or Importers of HHCB .............................................. ........... 22 Table 2-3. US Production/Import Volume of HHCB ...

................................................... ............... 24 Table 2-4. Estimated Distribution of Fragrance Oil s by Use .......................................... .............. 25 Table 2-5. Cosmetic Product Types and Upper Levels of Fragrance Incorporation .................... 25 Table 2-6. HHCB Degradation Half-Lives and Half-Di sappearance Times for Environmental

Media .............................................

................................................... .......................... 30 Table 2-7. BCFs and BAFs (L/kg ww) of HHCB in Aqua tic Vertebrates a ....................................... 32 Table 2-8. BCFs (L/kg ww) of HHCB in Benthic and T errestrial Invertebrates .......................... ... 33
Table 3-1. Measured Concentrations of HHCB in Biot a ................................................. .............. 36 Table 3-2. Aquatic Toxicity Data for HHCB - Acute Toxicity .......................................... ............... 42 Table 3-3. Aquatic Toxicity Data for HHCB - Chroni c Toxicity ........................................ .............. 44 Table 3-4. Sediment Toxicity Data for HHCB .......

................................................... ..................... 46 Table 3-5. Soil Toxicity Data for HHCB ...........

................................................... ........................... 48 Table 3-6. Concentrations of Concern (COCs) for En vironmental Toxicity .............................. .... 49
Table 3-7. Environmental Concentrations Used to Ca lculate RQs ....................................... ....... 50 Table 3-8. Calculated Risk Quotients (RQs) for HHC B ................................................. ................. 51



## LIST OF APPENDIX TABLES

Table_Apx A-1. Summary of Human Health Hazard Info rmation ........................................... .... 82 Table_Apx A-2. Human Biomonitoring Data for HHCB .

................................................... ............ 85 Table_Apx C-1. HHCB, HHCB Diastereoisomers, and Re lated Structural Analogs ....................... 95 Table_Apx D-1. CDR National HHCB Information a

.................................................. ..................... 98 Table_Apx D-2. CDR HHCB Industrial Use Information

.................................................. ............. 98 Table_Apx D-3. HHCB CDR Consumer Information .....

................................................... ............. 99 Table_Apx D-4. CDR Company Site Information (2012)

.................................................. ............ 99 Table_Apx E-1. Summary of Estimated Environmental Releases .......................................... .... 101 Table_Apx E-2. Geographic Distribution for Facilit ies under NAICS 32562 Toilet Preparation

Manufacturing .....................................

................................................... ......... 105 Table_Apx E-3. Geographic Distribution for Facilit ies under NAICS 325611 Soap and Other

Detergent Manufacturing ...........................

................................................... .. 107 Table_Apx E-4. Geographic Distribution for Facilit ies under NAICS 325612 Polish and Other

Sanitization Goods Manufacturing .................. ................................................ 10 8 Table_Apx G-1. Measured Concentrations of HHCB in Wastewater ........................................ . 114 Table_Apx G-2. Measured Concentrations of HHCB in Surface Water ..................................... 118 Table_Apx G-3. Measured Sediment Concentrations at Locations in the US ........................... 12 0 Table_Apx G-4. Measured Concentrations of HHCB in Biosolids and Sludge ........................... 12 3 Table_Apx G-5. Measured Concentrations of HHCB in Soil .............................................. ......... 126Table_Apx G-6. Measured Concentrations of HHCB in Biota *

.................................................. . 129

Table_Apx G-7. Summary of Substituted Values (μg/L ) for Water Samples ............................. 1 31 Table_Apx G-8. Summary of Box Plots for USGS HHCB Data .............................................. ...... 132 Table_Apx G-9. USGS Medium and Site Codes ........

................................................... ............... 133

## LIST OF FIGURES

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

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

.................................................. ............... 20



## LIST OF APPENDIX FIGURES

Figure Apx B-1. Chemical Synthesis of HHCB ......

................................................... ..................... 94 Figure Apx G-1. Monitoring Data Summary from USGS NWIS for HHCB Concentrations in

Water, Filtered ...................................

................................................... ......... 134 Figure Apx G-2. Monitoring Data Summary from USGS NWIS for HHCB Concentrations in

Water, Unfiltered .................................

................................................... ....... 135 Figure Apx G-3. Monitoring Data Summary from USGS NWIS for HHCB Concentrations in Solids

.................................................. ................................................... ................... 135

## AUTHORS/CONTRIBUTORS/ACKNOWLEDGEMENTS/REVIEWERS

This report was developed by the United States Envi ronmental Protection Agency (US EPA), Office of Chemical Safety and Pollution Prevention (OCSPP), Office of Pollution Prevention and Toxics (OPPT). The Workplan Chemical Risk Assessmen t for 1,3,4,6,7,8-Hexahydro-4,6,6,7,8,8- hexamethylcyclopenta- γ

-2-benzopyran (HHCB) was prepared based on currentl y available data and any additional information received during the public comment period and peer review process. Mention of trade names does not constitut e endorsement by the EPA.

EPA Assessment Team
**Leads:** Susan A. Laessig, OPPT/Risk Assessment Division ( RAD) Eva M. Wong, OPPT/RAD

**Team Members:**

Katherine Anitole, OPPT/RAD Fred Arnold, OPPT/RAD Kay Austin, OPPT/RAD Lynne Blake-Hedges, OPPT/Chemistry, Economics & Sus tainable Strategies Division (CESSD) Robert Boethling (retired), OPPT/RAD Ellie Clark, OPPT/Chemical Control Division Majd El-Zoobi, OPPT/RAD Amuel Kennedy, OPPT/RAD Chuantung Lin, OPPT/CESSD Andrea Pfahles-Hutchens, OPPT/RAD Jamie Suski, OPPT/RAD

**Management Lead:** Todd Stedeford, OPPT/RAD

Portions of this document were developed with suppo rt from Abt Associates, Eastern Research Group, Versar and SRC, Inc.

Acknowledgements

The following individuals contributed to portions o f the draft document:

Jonghee Koh (visiting scientist, Korean Environment al Protection Agency) Elizabeth Sommer (OPPT/CESSD)External Peer Review

EPA/OPPT released the peer review plan in August of 2012 and draft risk assessment and charge questions for peer review for public comment in January 2013. EPA/OPPT contracted with The Scientific Consulting Group, Inc. (SCG) to convene a panel of **ad hoc** reviewers to conduct an independent external peer review for the EPA’s draft work plan risk assessment for

HHCB. As an influential scientific product, the draft ris k assessment was peer reviewed in accordance with EPA’s peer review guidance. The peer review panel performed its functions by
web conference and teleconference between December 4, 2013 and February 6, 2014. The panel consisted of the following individuals:

**Daniel Schlenk, Ph.D. (chair)** University of California, Riverside

**Peter M. Chapman, Ph.D.** Golder Associates, Ltd.

**Bill Doucette, Ph.D.** Utah State University

**Robert W. Gensemer, Ph.D.**
GEI Consultants **Duane B. Huggett, Ph.D.** University of North Texas

**Shane Snyder, Ph.D.** University of Arizona

**Lawrence W. Whitehead, Ph.D.** University of Texas, School of Public Health

Please visit the EPA/OPPT’s Work Plan Chemicals web page for additional information on the peer review process ( http://www.epa.gov/oppt/existingchemicals/pubs/risk assess.html ), and the public docket ( Docket: EPA-HQ-OPPT-2012-0722 ) for the independent external peer review report and the response to comments document.

## GLOSSARY OF TERMS AND ABBREVIATIONS

BB Benzyl benzoate
BAF Bioaccumulation factor
BCF Bioconcentration factor CASRN Chemical Abstract Service Registry Number
CBI Confidential Business Information CDR Chemical Data Reporting
ChV Chronic value
COC Concentration of concern DEP Diethyl phthalate dw Dry weight

E 2 Estradiol EC European Commission ECHA European Chemicals Agency EPA Environmental Protection Agency EPCRA Emergency Planning and Community Right-to-Kn ow 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 Rodentici de Act GC/MS Gas chromatography/mass spectrometry GD Gestation day GLP Good Laboratory Practice HERA Human and Environmental Risk Assessment Proje ct HHCB 1,3,4,6,7,8-Hexahydro-4,6,6,7,8,8-hexamethylc yclopenta- γ -2-benzopyran HPLC High performance liquid chromatography 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 Pounds LDR Larval development ratio
LOAEC Lowest-observed-adverse-effect concentration
LOAEL Lowest-observed-adverse-effect level
LOEC Lowest-observed-effect concentrationLog 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 MATC Maximum Acceptable Toxicant Concentration MOS Margin of safety NAICS North American Industry Classification Syste m ND Not detected
NHANES National Health and Nutrition Examination Su rvey NOAEC No-observed-adverse-effect concentration
NOAEL No-observed-adverse-effect level
NOEC No-observed-effect concentration NWIS National Water Information System NWQL National Water Quality Laboratory OCSPP Office of Chemical Safety and Pollution Prev ention OECD Organisation for Economic Co-operation and De velopment
OPPT Office of Pollution Prevention and Toxics
OSPAR Oslo Paris
PNEC Predicted no-effect concentration REACH Registration, Evaluation, Authorisation and Restriction of Chemicals

RHO earthworm Bulk density of earthworm RAR Risk assessment report RQ Risk quotient SCAS Semi-continuous activated sludge SIAM SIDS Initial Assessment Meeting SIAR SIDS Initial Assessment Report SIDS Screening Information Data Set SSO Sanitary sewer overflow
TG Test Guideline TLC Thin-layer chromatography
TSCA Toxic Substances Control Act UF Uncertainty factor

UF total Total uncertainty factor
US United States USGS United States Geological Survey wt% Weight percent ww Wet weight WWTP Wastewater treatment plant yr(s) Year

## EXECUTIVE SUMMARY

As a part of EPA’s comprehensive approach to enhanc e the Agency’s existing chemicals management, in March 2012 EPA identified a work pla n of chemicals for further assessment under the Toxic Substances Control Act (TSCA). 1 The Agency is performing risk assessments on chemicals in the work plan. If an assessment identi fies unacceptable risks to humans or the environment, EPA will pursue risk management. 1,3,4 ,6,7,8-Hexahydro-4,6,6,7,8,8- hexamethylcyclopenta- γ -2-benzopyran (HHCB) was assessed as part of the work plan.

HHCB is one of the most widely used polycyclic musk fragrance ingredients in a range of consumer products including perfumes, cosmetics, sh ampoos, lotions, detergents, fabric softeners, and household cleaners. HHCB is not prod uced in the US, but is imported. The volume of HHCB imported into the US has been report ed in the range of 1 to 10 million pounds (lbs) per year from 1990 to 2012. Within this range , the volume of use in the US has increased steadily since 2000, suggesting that polycyclic mus ks are not being replaced with other synthetic musks, as may be occurring in Europe.

In the US, imported HHCB is compounded into fragran ce oils, which are blended into end-use products and sold for both commercial and consumer use. An estimated 90 to 94 percent of the HHCB use-volume is released to municipal wastewater . HHCB is moderately persistent in soil and sediment, has low mobility in soil, suboptimal removal in wastewater treatment plants (WWTP) (with removal primarily through sorption to sludge), and bioconcentration and bioaccumulation into aquatic organisms is low to mo derate, depending on the species in which
it is measured.

Problem formulation resulted in the assessment focu sing on environmental risk to the aquatic environments from the use of HHCB as a fragrance in gredient in consumer and commercial products. Exposure pathways of interest included di scharge of wastewater to surface water or application of biosolids to land. The assessment e ndpoints of interest were acute and chronic toxicity to aquatic organisms, chronic toxicity to sediment organisms, and chronic toxicity to terrestrial invertebrates and plants. In addition, EPA reviewed the available human health data, including a risk assessment performed by the Europe an Union, evaluated the weight of the
evidence and determined that further assessment of human health risk was not currently needed. Appendix A summarizes the human health info rmation for HHCB.

US environmental monitoring data from the literatur e and from the US Geological Survey (USGS) National Water Information System (NWIS) dat abase were used in the risk assessment. Levels of HHCB measured in wastewater, surface wate r, sediment, and soil were identified from these sources. More than 6800 data points for efflu ent, surface water and sediment were
available for HHCB in the NWIS database. Soil moni toring data was limited to two published
studies.

1 http://www.epa.gov/oppt/existingchemicals/pubs/work plans.htmlEcotoxicity studies for HHCB have been conducted in fish, aquatic invertebrates, aquatic plants,

sediment invertebrates, soil invertebrates, and ter restrial plants. Concentrations of concern (COCs) were derived from these studies according to established EPA/OPPT methods (EPA 2012f; 2013). This assessment considered only direc t exposure to aquatic and sedimentdwelling organisms via contact with contaminated wa ter or sediment. Based on half-lives in water, sediment and continuous wastewater inputs, e xposure to organisms was considered

continuous.

Potential risks were calculated using the risk quot ient (RQ) approach. In this deterministic approach, the RQ is calculated by dividing a point estimate of exposure concentration by a point estimate of the effects concentration of conc ern (the COC). A RQ greater than one indicates there may be risk of concern. To incorpor ate the range of measured environmental concentrations, the RQ was calculated using the mea n, maximum, and 95th percentile of
measured environmental concentrations for surface w ater and sediment. Available soil concentrations were inadequate for calculating an R Q for soil organisms. RQs were calculated for aquatic and sediment dwelling organisms for acu te and chronic exposures to water and sediments. RQs were not calculated for terrestrial invertebrates and plants due to insufficient exposure data.

Maximum measured surface water concentrations were approximately 25 to >150 times lower than acute COCs, and 4 to 25 times lower than the c hronic COC for aquatic organisms (i.e., RQs ranged from 0.006 to 0.04 and 0.04 to 0.23, respect ively). The chronic COC for sedimentdwelling organisms was also not exceeded at the upp er range of measured environmental concentrations for the maximum published values or the more than 600 sediment measurements from the USGS. RQs for chronic risk t o sediment dwelling organisms ranged from 0.001 to 0.36. RQs were not calculated for te rrestrial invertebrates and plants **via** contact with contaminated water, sediment, or soil.

In conclusion, under the exposure scenarios assesse d in this this risk assessment, current environmental exposure concentrations are one to tw o orders of magnitude below hazard concentrations (RQs < 1) of concern for aquatic or sediment-dwelling organisms. The inability to assess potential risks to terrestrial invertebrates and plants is a major uncertainty associated
with this assessment.

## 1 BACKGROUND AND SCOPE

1.1 INTRODUCTION

As a part of EPA’s comprehensive approach to enhanc e the Agency’s existing chemicals management, in March 2012 EPA identified a work pla n of chemicals for further assessment under the Toxic Substances Control Act (TSCA). 2 After gathering input from stakeholders, EPA developed criteria used for identifying chemicals f or further assessment. 3 The criteria focused on chemicals that meet one or more of the following factors: (1) potentially of concern to children’s health (for example, because of reproduc tive or developmental effects); (2) neurotoxic effects; (3) persistent, bioaccumulative , and toxic (PBT); (3) probable or known carcinogens; (4) used in children’s products; or (5 ) detected in biomonitoring programs. Using this methodology, EPA identified a TSCA Work Plan o f chemicals as candidates for risk assessment in the next several years. In the priori tization process, 1,3,4,6,7,8-Hexahydro4,6,6,7,8,8-hexamethylcyclopenta- γ -2-benzopyran (HHCB) was identified for assessment based on high production volume, high estimated releases to the environment, moderate environmental persistence, moderate bioaccumulation , moderate toxicity in rodents, and high potential for human exposure.

The Agency is performing risk assessments on chemic als in the work plan. If an assessment identifies unacceptable risks to humans or the envi ronment, EPA will pursue risk management. The target audience for this risk assessment is pri marily EPA risk managers; however, it may
also be of interest to the broader risk assessment community as well as US stakeholders that are interested in HHCB. The information presented i n the risk assessment may be of assistance to other Federal, State, and Local agencies as well as to members of the general public who are
interested in the risks of HHCB.

The initial step in EPA’s risk assessment developme nt process, which is distinct from the initial prioritization exercise, includes scoping and probl em formulation. During these steps EPA reviews currently available data and information, i ncluding but not limited to, assessments conducted by others (e.g., authorities in other cou ntries), published or readily available reports, and published scientific literature. During scoping and problem formulation the more robust review of the factors influencing initial prioritiz ation may result in refinement – either addition/expansion or removal/contraction – of spec ific hazard or exposure concerns previously identified in the prioritization methodology.

HHCB (CASRN 1222-05-5) is a high production volume fragrance ingredient in consumer and commercial products. HHCB is one of the most common synthetic polycyclic musks and is used as an ingredient in consumer products including per fumes, cosmetics, shampoos, lotions, detergents, fabric softeners, air fresheners, and c leaning agents.

2 http://www.epa.gov/oppt/existingchemicals/pubs/work plans.html

3 http://www.epa.gov/oppt/existingchemicals/pubs/wpme thods.pdfPrevious assessments in Europe, Sweden, and Austral ia have evaluated the human and/or

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). The 2008 EU Risk Assessment Report (RAR) for HHCB (EC, 2008), concluded: “Ther e is at present no need for further information and/or testing and no need for risk red uction measures beyond those which are being applied already.” No potential risk concerns were identified for surface water and sediment in Sweden (SWECO, 2008) or for the marine environment (OSPAR, 2004). Potential
risk concerns were identified for HHCB in certain h ighly contaminated sediments in Europe (HERA, 2004) and in an initial screening study of p harmaceuticals, personal care products, and endocrine disrupting compounds in surface runoff fr om biosolids-amended land in Australia (Langdon et al., 2010).

A screening-level risk assessment for chemicals of emerging concern in California receiving waters recently identified a risk concern for HHCB in effluent dominated inland waterways and coastal embayments and recommended monitoring in in land waterways (Anderson et al., 2012). Further, California lists HHCB as a designat ed chemical for biomonitoring, however, California is not yet biomonitoring for these chemi cals (CA EPA OEHHA, 2014b). In addition, the Oregon Department of Environmental Quality lists HH CB as a priority persistent pollutant because it has a documented effect on humans, fish, wildlife, and plants (Oregon DEQ, 2010a; Oregon DEQ, 2011). Oregon also posts use, exposure pathways and release data for HHCB under this program (Oregon DEQ, 2010b). In addition , Minnesota classifies HHCB as a chemical of high concern (MDH, 2013).

HHCB has not been the subject of specific regulator y activity in the US, Canada, or the European Union (EU). HHCB is listed on the TSCA Inventory, is a High Production Volume (HPV) chemical substance, and is reported for Chemical Data Report ing (CDR), but is not otherwise subject to specific TSCA regulations. It is not listed on the Emergency Planning and Community Right-toKnow (EPCRA) Section 313 List of Toxic Chemicals. H HCB is approved as an inert ingredient for fragrance and non-food use in pesticide products un der the Federal Insecticide, Fungicide and Rodenticide Act (FIFRA). 4 In Canada, HHCB is listed on the Domestic Substan ces 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 assess ment. In Europe, HHCB is registered under the Registration, Evaluation, Authorisation and Res triction of Chemicals (REACH) regulation.

4 HHCB is commonly diluted in diethyl phthalate (DEP) , benzyl benzoate (BB), or isopropyl myristate (IPM ) before being added to fragrance formulations Rimkus (1999) . However, EPA recently proposed a ban for DEP as an inert ingredient in pesticide products (EPA, 2012c).

### 1.2 PROBLEM FORMULATION

EPA/OPPT relied on Agency guidance to develop this the risk assessment (EPA, 1998 and EPA 2014a). EPA followed the risk assessment paradigm b eginning with planning and scoping and problem formulation.

EPA/OPPT approached problem formulation by reviewin g other recently performed assessments and searching published literature from 2005, the date of the last literature review indicated in the EU RAR (EC, 2008), to May, 2012. T he literature review included HHCB chemistry, uses, sources including industrial relea ses, fate, exposure, and hazard to humans and ecological receptors. Other sources of information reviewed were unpublished reports provided by the Research Institute for Fragrance Ma terials (RIFM), the publicly available EPA Inventory Update Reporting (IUR) and Chemical Data Reporting (CDR) databases, and the US Geological Survey National Water Information Servic e (NWIS). Data was acceptable if it met standard quality criteria which varies according to the type of information (data quality criteria are described in each later section of this report)

.

During problem formulation, EPA/OPPT reviewed and s ummarized the following information:

• HHCB has an aggregate annual import volume of 3,126 ,728 lbs. HHCB was reported as imported to the US by three companies in 2012. Thr ee additional companies that reported to CDR are also assumed to import.

• TSCA uses for HHCB are as an ingredient in detergen ts, fabric softeners, dishwashing detergents, and commercial and consumer general pur pose cleaners. Non-TSCA uses include cosmetics and personal care products, which are regulated under the Federal Food, Drug, and Cosmetic Act.

• HHCB is not included on the list of substances for which information on environmental releases is collected as part of the Toxics Release Inventory (TRI) Program.

• Due to use in a wide variety of commercial and cons umer products, HHCB is continuously released to the environment in municip al wastewater and sludge.
Releases to the environment from industrial sites a re also expected.

• HHCB is not readily biodegradable and is expected p artition to organic-rich particles in sediment, soil and sludge, based on the physical-ch emical properties.

• HHCB is moderately persistent in the environment an d bioconcentration has been measured as low to moderate, depending on the speci es tested. HHCB has been found to have low bioaccumulation potential and to be bio transformed to more polar, less toxic, metabolites in some organisms.

• HHCB is highly toxic to aquatic organisms under bot h acute and chronic exposure conditions. Chronic effects of HHCB in aquatic org anisms may include effects on adult and embryo survival, behavior, respiration, and lar val development. Mechanistic studies (described in Appendix F) show that HHCB ma y interfere with steroid hormonereceptors, metabolism, and membrane transporters in fish and aquatic 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 U S Geological Survey (USGS) National Water Information System (NWIS) database and in the published scientific literature.

• Human health hazards and risks of HHCB have been pr eviously assessed by other organizations. Appendix A summarizes the conclusion s of the 2008 EU RAR.

• Human dermal and inhalation exposures occur intenti onally as a result of use of cosmetics and personal care products and may also o ccur during use of detergents and cleaners. HHCB is frequently measured in human bloo d and breast milk (see Appendix A for a summary of human biomonitoring data).

• Developmental effects occur at relatively high oral exposures (yielding a conservative LOAEL of 500 mg/kg/day) and no developmental toxici ty occurs at levels several times greater than levels detected in human breast milk ( in rodent studies, pups in the highdose group were estimated to be exposed to levels a pproximately 100 to 360 times the maximum level found in human milk samples (1,316 μg /kg fat)). A 2-year cancer study
is not available for HHCB.

• Based on review of available human health data, inc luding a risk assessment performed by the European Union, EPA/OPPT determined that fur ther assessment of human health risk was not currently needed.



### 1.3 CONCEPTUAL MODEL FOR ENVIRONMENTAL ASSESSMENT
The exposure scenarios within the scope of EPA/OPPT ’s HHCB assessment are depicted in the Conceptual Model (Figure 1-1).

As mentioned previously, HHCB is one of the most wi dely used polycyclic musk fragrance ingredients in a range of consumer products includi ng perfumes, cosmetics, shampoos, lotions, detergents, fabric softeners and household cleaners . Because exposure to consumers occurs intentionally for a majority of these uses, EPA/OPP T reviewed the available human health data and subsequently determined that further assessment of human health risks was not currently
needed.

After commercial or residential use of consumer pro ducts containing HHCB, the majority of
HHCB enters the wastewater stream. Releases to the environment from industrial sites were estimated to be at most < 10% of the HHCB productio n/import volume (see Appendix E: Release Estimates). Due to the ubiquity of fragranc ed products, consumer uses are a continuing
source of HHCB to wastewater. When HHCB enters wast ewater treatment plants (WWTPs), the chemical is negligibly volatile and expected to par tition to organic-rich particles during wastewater treatment based on the physical-chemical properties. However, HHCB removalduring treatment in WWTPs is suboptimal (see Chapte r 2, Fate in Wastewater) and HHCB is

present in both wastewater effluent and sludge (WER F, 2007).

During the wastewater treatment process, influent i s treated, producing sludge and effluent that is released to surface water at outfall sites, which may be located in freshwater, estuarine, or marine environments. Under this scenario, HHCB e nters the environment via the effluent and is diluted to varying degrees in streams, lakes , estuaries, and oceans where aquatic organisms may be exposed. In addition, thousands of sanitary sewer overflows (SSOs) occur each year in the US, resulting in releases of billi ons of gallons of untreated wastewater directly
to surface water. When HHCB has not been removed du ring the treatment process and is present in effluent, sorption to suspended solids i n the surface water and partitioning to sediment is expected. Under these conditions, HHCB may be available for uptake by fish, aquatic invertebrates, aquatic plants, and benthic organisms, leading to potential exposure. The route of exposure is likely dermal and branchial in aquatic organisms and can also be by oral ingestion in sediment and soil organisms. As HHCB i s negligibly volatile, it is likely to enter the environment during the wastewater treatment process or subsequently, from receiving waters.

Figure 1-1. Conceptual Model for HHCB Assessment

* Includes >90% of releases to the environment from all fragranced products such as soaps, detergents, fabric softeners, shampoos, cosmetics, and cleaners. Indus trial releases are <10% of releases and are not ind icated in the diagram. **EPA estimates that between 23,000 and 75,000 sani tary sewer overflows (SSOs) occur each year in the US, resulting in releases of untreated wastewater betwe en 3 billion and 10 billion gallons. Dotted line indicates the pathway was not assessed.

Because of its properties, HHCB also partitions to solid phases in the wastewater treatment process (sludge), and with further treatment can be come concentrated in biosolids. This organic carbon-rich material is disposed of by land fill or incineration, or may be utilized for land application to enhance physical soil properties as well as plant yield (Correa et al., 2005). WWTPs need to dispose of waste sludge continuously, and about half of all sludge is applied toagricultural land each year as biosolids. 5 Once biosolids containing HHCB are applied to land,

HHCB has a long half-life in soil and may be availa ble for uptake by plants and soil invertebrates, leading to potential exposure.



### 1.4 ANALYSIS PLAN FOR ENVIRONMENTAL ASSESSMENT

Based on the Conceptual Model described above, the scenarios evaluated in this assessment
are release of HHCB in wastewater to surface water and sediment. The exposure assessment is based on available monitoring data for the US. The ecological assessment endpoints evaluated for HHCB are acute and chronic toxicity to aquatic organisms and chronic toxicity to sediment organisms. Chronic toxicity to terrestrial inverteb rates and plants was not evaluated due to
insufficient data for HHCB concentrations in soil.

EPA/OPPT reviewed other recently performed assessme nts and searched published literature from 2005, the date of the last literature review i ndicated in the EU RAR (EC, 2008), to May, 2012. The literature review included HHCB chemistry , uses, sources including industrial releases, fate, exposure, and hazard to humans and the environment. Other sources of information reviewed were unpublished reports provi ded by the Research Institute for Fragrance Materials (RIFM), the publicly available EPA Inventory Update Reporting and Chemical Data Reporting databases, and the US Geolo gical Survey National Water Information Service (NWIS).

Key ecological hazard reports were obtained for cri tical review and data quality assessment. Details on the data quality criteria for ecological hazard assessment are included in Chapter 3.

Due to the availability of a sufficient quantity of monitoring data for the US, EPA/OPPT analyzed measured levels of HHCB and data from ecot oxicological studies of acceptable quality to assess the risks of HHCB to the aquatic environment by calculating conservative acute and chronic concentrations of concern (COCs). The potential risks to aquatic and benthic organisms were characterized by calculating the ris k quotient (RQ) for each route of environmental exposure ( **i.e.** , water and sediment) bas e d on measured media concentrations.

5 About 7.2 dry US tons of biosolids were beneficial ly used or disposed in 50 states in 2004. Of that total, approximately 55 percent was applied to soils for a gronomic, silvicultural, and/or land restoration pu rposes, or was likely stored for such uses. The remaining 45 percent was disposed of in municipal solid waste la ndfills, surface disposal units, and or incineration facilit ies. (NEBRA, 2007).

## 2 SOURCES AND ENVIRONMENTAL FATE

2.1 INTRODUCTION

As an initial step in assessing the environmental r isks of HHCB, physical and chemical properties of HHCB, sources of HHCB related to production and uses, and fate of HHCB in the environment
are described.

2.1.1 Physical and Chemical Properties

HHCB is an almost colorless viscous liquid with a m usk-like odor. The polycyclic chemical structure for HHCB is shown in Figure 2-1.

C H 3 CH 3

CH 3 CH 3 C H 3

O CH 3

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

Synonyms for HHCB include:

• 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamehyl-cyclope nta-gamma-2-benzopyran

• 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-hexamethyl-cyclop enta-[ 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

Technical HHCB consists of a mixture of isomers tha t are unspecified. Four diastereoisomers ( **i.e.** , an isomer that differs in the spatial arrangement of atoms in the molecule, but is not a mirror image) of HHCB are known to exist: two (-)/ 4S isomers (4S, 7R & 4S, 7S) have the characteristic odor, and the other two (+)/4R isome rs (4R, 7R & 4R, 7S) have little to no odor(DrugLead, 2009; http://www.druglead.com/cds/hhcb.html ). There are many other structural

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

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

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

Molecular formula C 18 H 26 O

Molecular weight 258.44 g∙mol -1

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

Melting point -10 to 0 °C (determined by cooling viscous liquid to –30 °C and gradual warm up)

Boiling point 320 °C at 760 mmHg (converted from 16 0 °C at 4 hPa)

Vapor pressure 0.0727 Pa (5.47 × 10 -4 mmHg) @ 25 °C (measured; OECD b Test Guideline 104)

Logarithmic octanol:water partition coefficient (log K OW ) 5.3 ("slow stirring" method) b ; 5.9 @ 25 °C (measured; OECD Test Guideline 117)

Water solubility 1.65 mg/L at pH 7 (at 25 °C) c ; 2.3 mg/L at 20 °C (measured; OECD Test Guideline 10 5)

Flash point 144 °C

Henry’s Law constant 1.13 x 10 -4 atm∙m 3 /mol (at 25 °C, calculated using measured vapor pressure and water solubility);

1.32 x 10 -4 atm∙m 3 /mol (at 25 °C, estimated using HENRYWIN program in EPI Suite v4.11);

3.65 x 10 -4 atm∙m 3 /mol (at 25 °C, measured) d

a Source: ( HSDB

, 2007 ) except as noted.

b OECD – Organization for Economic Cooperation and De velopment

c pH is not expected to effect HHCB solubility as it is non-ionizable

d Sources: EC, 2008; Artola-Garicano, 2002



### 2.2 PRODUCTION AND USES

This section discusses the US production volumes an d uses specifically for HHCB and more generally for fragrances. The manufacturing process for HHCB is described in Appendix B. TheHHCB commercial product is diluted (65 percent wet weight [ww]) in diethyl phthalate (DEP) 6

,

benzyl benzoate (BB), or isopropyl myristate (IPM) prior to compounding and formulation into products (HERA, 2004b). HHCB is a High Production V olume chemical that is widely used in cleaning and personal care products. Data on produc tion volume and uses are amenable for determining release estimates to the environment (S ee Appendix E). These data are also useful for understanding potential exposure routes and pat hways through which HHCB may enter the environment through industrial and consumer applica tions.

2.2.1 Production HHCB is one of the most widely used and consumed po lycyclic musks that represents 90 percent of the total US polycyclic musk market (EC, 2008; HERA, 2004a, respectively). Five US firms reported some non-confidential business infor mation (CBI) CDR data for HHCB in 2012 (EPA, 2014) 7 and a sixth reported only CBI information 7 . One of the five firms reported production at two sites. A list of the US producers of HHCB is provided in Table 2-2. Of the five companies listed, three reported importing HHCB, tw o did not indicate whether they are a manufacturer or importer. EPA/OPPT assumes that the se companies and the sixth company that reported only CBI information are also importe rs because HHCB is not manufactured in the US (IFRA, 2012a). HHCB is imported in liquid forms at a maximum concentration of ≥90 percent (EPA, 2014b).

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

Chemical Company a Reported CDR
Data in 2012 Manufacturer or

Importer

HHCB, Galaxolide, Musk GX, Abbalide Berje, Inc. Yes Importer

International Flavors & Fragrances, Inc. Yes Importer
Symrise, Inc. Yes Importer
S C Johnson & Son, Inc. Yes No data b

Firmenich, Inc. Yes No data b

a One company reported as a producer of HHCB in 2012, however all data were CBI so the company is not in cluded
in this table.

b ‘No data’ indicates that data are not available on whether the company is a manufacturer or importer. Source: EPA (2014b).

Between 1996 and 2000, the US consumption of synthe tic 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 (Somogy i and Kishi, 2001; as cited in Peck et al., 2006). Global musk production increased by 12.5 per cent between 1987 and 1996 and during

6 EPA plans to no longer approve DEP as an inert ingr edient in pesticide products (US EPA, 2012c).

7 These six producers may be an underestimate becaus e a production site is only required to submit a Fo rm U (the CDR reporting instrument) if it produces or imports more than 25,000 pounds of a chemical during the r eporting

year.this time period, production shifted from nitro mu sks, such as musk xylene, to polycyclic musks

(Rimkus, 1999; as cited in Luckenbach and Epel, 200 5). This shift, reflected in decreasing production rates of nitro musks and increasing prod uction rates of polycyclic musks (Hutter et al., 2005), is expected to continue for several rea sons. In June 2009, the European Chemicals Agency (ECHA) recommended xylene musk for authoriza tion under REACH (ECHA, 2009a). It was added to the authorization list in February 201 1, which means that musk xylene can be used only in cases where an authorization has been granted for a specific use (European Union, 2011). Additionally, in June 2009, the Internationa l Fragrance Association (IFRA) voluntarily phased out musk xylene through the IFRA Standards, part of the fragrance industry’s global selfregulatory program contained in the IFRA Code of Pr actice. The Code of Practice is mandatory for IFRA members, and membership accounts for appro ximately 90 percent of the global volume of fragrance materials (IFRA, 2011).

Data from IFRA have shown an increase in the volume of HHCB (point estimate opposed to range estimate) used in the US from the years (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 y r, respectively (IFRA, 2012b). 8 According to IFRA, the increase in HHCB use in the US is not a r eflection of increased use of this particular musk over others, but a reflection of increased mar ket demand for fragranced consumer products, which has expanded the market for HHCB an d 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 p olycyclic musks (Gatermann et al., 1999; as
cited in Peck et al., 2006). At present, it appears as though the US production volume (which includes import) trends for HHCB and musk xylene in the US are consistent for both chemicals. For example, the annual (non-CBI) production volume of HHCB has ranged between 1 and 10 million lbs since 1990 (Table 2-3). In addition, wi th the exception of 2002, the non-CBI IUR production volume for musk xylene has been consiste ntly reported to be <500,000 lbs (EPA, 2012d; 2012e). Therefore, and in contrast to the EU , the steady range of production volumes for both HHCB and musk xylene suggests that there h as not been a significant shift away from nitro musks to polycyclic musks in the US.

8 IFRA’s letter to EPA dated March 30, 2012 indicate d that the 2008 use volume of HHCB in the US was approximately 1,300 tons per year. This estimate w as revised by IFRA in an email dated June 29, 2012 to 1,600 tons per year in 2008.Table 2-3. US Production/Import Volume of HHCB

Chemical Production/Import Volume (in Thousands, K, or Milli ons, M, of lbs)

Year

1986 1990 1994 1998 2002 2006 2012

HHCB, Galaxolide, Musk GX, Abbalide 500K-1M 1M-10M 1M-10M 1M-10M 1M-10M 1M-10M 3.1M

Musk Xylene 500K

- 1M 10K

- 500K 10K

- 500K 10K

- 500K 10K

- 500K <500K NR

NR- No volume reported

Sources: EPA (2012d; 2012e; 2014b)

2.2.2 Uses Musks are considered important compounds to the fra grance 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, 20 09b).

HHCB is used as a fragrance ingredient in cleaning and personal care products because it is alkali-stable and does not discolor in light (Ash e t al., 2009; OSPAR, 2004; IFF, 2012; Fahlbusch et al., 2012). HHCB and other musks provide a uniqu e, 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 frag rances 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 dir t in the wash solution (Schreiber, 2004). Synthetic musks, including HHCB, also may be used t o mask chemical odors and can be found in products labeled “unscented,” but do not seem to be added to products labeled “fragrance free” (Potera, 2007).

The EU RAR (EC, 2008) estimated that in the EU, app roximately 86 percent of manufactured HHCB is used as part of a fragrance mixture, and 14 percent is used directly in the bulk formulation of products. HHCB is, however, generall y estimated to make up only two to four percent of fragrance compounds because fragrance mi xtures can contain from 50 to 300 ingredients (HERA, 2004a; Bickers et al., 2003). As shown in Table 2-4, the majority of fragrance oils (25 percent) are used in detergents, followed by fabric softeners (14 percent) and personal care products (13 percent), whereas on a w eight percent (wt%) basis, perfume extracts have the highest level of fragrance incorp oration at 20 wt% (Table 2-5). Given the importance of HHCB as a fragrance ingredient, EPA/O PPT expects that the use of HHCB has a similar distribution pattern as fragrance oils in g eneral with the largest percentage of volume used in detergents, fabric softeners, and personal care products.Table 2-4. Estimated Distribution of Fragrance Oil s by Use

Use Fragrance Oils (percent)

Detergent 25

Fabric softeners 14

Personal care 13

Bath and shower 10

Hair care 10
Soaps 9

Industrial and household cleaner 8

Other 6
Fine fragrances 5

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

Table 2-5. 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 appl ication of different cosmetics (COLIPA, 1987; as cited in Cadby et al., 2002).

b Estimates of typical fragrance levels in different products and maximum likely proportion of fragrance remaining on skin after normal product us e (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 pr oducts tested, including 100% ofdeodorants tested (Roosens et. al., 2007). A 2002 s urvey found that HHCB may be found at

levels between 0.02 and 0.90% in household cleaning products (IFRA, 2002 as reported in HERA, 2004b). Another recent study by Reiner and Kannan ( 2006) analyzed products for concentrations of HHCB and a degradant of HHCB, HHC B-lactone. In this study, HHCB concentrations were from <5 to 646,000 mg/kg sample and HHCB-lactone was <5 mg/kg in the same products. This study also showed that products within a single class might contain differing levels of HHCB. For example, it was noted that HHCB concentrations in two different laundry detergents were 79 and 84,900 mg/kg sample.

A wide range of consumer products contain HHCB. How ever, these studies indicate that levels varied significantly across and even within product types. This variation is not unexpected given the complexity of fragrance mixtures which often co ntain hundreds of constituents.

2.2.3 Conclusions of Production and Use Use volume of HHCB has increased gradually during 2 000-2011 due to increased demand for fragranced products. HHCB is used as part of fragra nce mixtures in products such as detergents and fabric softeners, and personal care products. T he concentration of HHCB in any of these products ranges widely across product types as well as within single product categories ( **e.g.** , laundry detergent).



### 2.3 ENVIRONMENTAL FATE

HHCB enters domestic wastewater treatment and is su bjected to removal by a variety of processes. HHCB and any degradation products genera ted in wastewater treatment are mainly discharged to receiving waters or disposed o f in biosolids. This section summarizes current knowledge of the fate and degradation of HH CB in waste water and the persistence,
bioconcentration and bioaccumulation of HHCB in the environment. Biological and abiotic reactions of HHCB as they relate to environmental f ate and distribution are summarized
below and in Table 2-6.

2.3.1 Environmental Persistence

2.3.1.1 Fate in Wastewater Treatment

The half-life of HHCB in activated sludge at concen trations of 5, 17.4, 25, 25 μg/L has been reported as 69, 10-15, 21, 33 hours, respectively ( Federle et al., 2002; Schaefer, 2005; Langworthy et al., 2000, as cited in EC, 2008). HHC B disappearance with subsequent appearance of more polar entities was observed (Lan gworthy et al., 2000; as cited in EC, 2008). The geometric mean from these studies for activated sludge half-disappearance time was is 22.5 hours. This corresponds to “moderate-to-slow” biodegradation in activated sludge; seeguidance in the Estimation Programs Interface (EPI) Suite v4.11 9 (EPA, 2011). The principal

degradation product of HHCB is HHCB-lactone, which can be present at levels as high as 50% of HHCB concentrations (Bester, 2005; Horii et al., 20 07). Formation of the lactone appears to occur via aerobic biodegradation. However, it has also been reported to be formed from treatment with hypochlorite (Kulich et al., 2010) a nd present in products containing HHCB, which the authors suggest may be as an impurity in technical HHCB (Reiner and Kannan, 2006).

HHCB is expected to partition strongly to solid pha ses based on its high measured log K OW of 5.9 (see Rimkus, 1999 for a summary of values for musks ) and the soil/sediment organic carbon partition coefficient (log Koc = 3.6-3.9; EC, 2008) which is supported by the estimated K OC (log basis) of 4.1-4.3 (KOCWINTM program v2.00; in EPI Su iteTM v4.11 9 , (EPA, 2011)). Values for both

K d (sorption coefficient) and K OC (organic carbon-normalized sorption coefficient) a re generally in the range of 3 to 4 on a logarithmic scale. This means that HHCB will be substantially removed by sorption to sludge in WWTPs; will have l ow mobility in soil; and will bind strongly to benthic and suspended sediment. The Human and Envir onmental Risk Assessment Project (HERA, 2004a) summarized both theoretical and measu red percent removal values (up to 2004) for European and US activated sludge-based treatmen t systems. Observed removal was generally >50 percent (usually >80 percent), and re moval percentage correlated well with removal of total suspended solids (Simonich et al., 2002). Kupper et al. (2006) reported overall
HHCB removal of 81% for a full-scale Swiss treatmen t plant, in line with the earlier results; and Horii et al. (2007) reported similar overall HHCB r emoval for wastewater treatment plants in Kentucky and Georgia.

Waste (settled) activated sludge in WWTPs is genera lly sent to anaerobic digesters to reduce sludge volume and organic load before disposal of t he concentrated material. Removal of HHCB
under anaerobic treatment conditions has been addre ssed in several studies. Xue et al. (2010) reported poor overall removal of HHCB in a full-sca le anaerobic/anoxic/aerobic system combined with membrane bioreactor for municipal was tewater reclamation, and little removal was attributed to biodegradation. In contrast Carba lla et al. (2007) reported good removal of HHCB (60-80%) in mesophilic and thermophilic anaero bic sludge digestion. Kupper et al. (2006) observed apparent anaerobic degradation of HHCB in the sludge line of a full-scale treatment system, but the mechanism was not confirmed. Theref ore, no direct evidence of anaerobic (methanogenic) biodegradation of HHCB exists.

2.3.1.2 Fate in Water HHCB will bind strongly to solid phases in the envi ronment; therefore to benthic and suspended sediment. Though HHCB is negligibly volat ile in WWTP, volatilization is an important process for HHCB in water. After 12.5 day s, about 40 percent of the radioactivity was lost by volatilization. In a more recent presen tation reporting on the same or extended experiments, with 0.5 μg/L of HHCB in river water, the half-life for disappearance of the

9 Available for download from http://www.epa.gov/opptintr/exposure/pubs/episuite. htm and http://www.epa.gov/oppt/exposure/pubs/halflife.htmparent substance was 43 hours (Federle et al., 2002 ). 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, 2005; as reported in EC, 2008). A significant fraction (u p to 16 percent) was volatilized after 28 days. These studies show that specific test conditi ons impact observed volatilization behavior, but that in any case volatilization is si gnificant and half-lives are on the order of days to weeks.

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 the modified river die-away tests of Langworthy et al. (2000), Federle et al. (2002), and Schaefer (2005) (all as cited in EC, 2008), which w ere designed to simulate river water conditions immediately below discharge of treated s ewage from activated sludge treatment plants, half-lives for HHCB disappearance of 10-69 hours were observed (see Table 2-6). Degradation is expected to be slower in streams tha t do not receive effluents from WWTPs or that are less impacted by such effluents as it can be expected that below WWTP discharge points, acclimated microbial populations will have developed.

In batch experiments (Langworthy et al., 2000; Federle et al., 2002; as cited in EC, 2008), effluent diluted in river water was spiked with rad iolabeled HHCB at concentrations of 1.0 or 0.5 μg/L. The observed degradation half-life of HHC B was 33 to 43 hours, depending on the starting concentration. Following 29 days, polar me tabolites made up 60% of the original radioactivity and the mass balance accounted for 92 % of the radioactivity.

In a series of die-away tests, a variety of metabol ites were formed at different rates and with increasing polarity over time (Schaefer et al., 200 5; as reported in EC, 2008). In the river dieaway test the total radioactivity of metabolites ro se from 25 to 56 percent between days 2 and 7 and to nearly 62 percent at 28 days. When the res ults were corrected for the amount
volatilized and the non-recovered fraction as well as for abiotic transformation, the primary biodegradation extent on day 28 was approximately 6 0 percent. About 40 percent of the radioactivity was lost by volatilization at 300 hou rs, but some of the volatilized radioactivity could have been in the form of degradation products , since degradation in water was also happening during the test.

2.3.1.3 Fate in Soil and Sediment HHCB has low mobility in soil. The fate of HHCB was studied using microcosms containing oak forest soil and agricultural soils, river sediment, and agricultural soil that routinely receives applications of biosolids from a domestic WWTP (slu dge amended soil) (Envirogen, 1998; as cited in EC, 2008). In the study, sealed flasks wit h HHCB-spiked soil at 10 μg/g soil (10 mg/kg) were incubated at laboratory ambient temperature fo r one year. Periodically the headspace was flushed for oxygen replenishment and volatile o rganics and carbon dioxide were captured.After incubation, and exhaustive extraction with et hyl 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, a nd agricultural soil, respectively. After one year, the average remaining HHCB for the four soil types was 14 percent of the HHCB initially present. The TLC results showed that HHCB was degra ded to various more polar metabolites. It was hypothesized that the majority of unrecovered r adiolabel became covalently bound to soil organic compounds ( **i.e.** , was immobilized by humification).

DiFrancesco et al. (2004) studied the dissipation of fragrance materials in biosolids-amended soils in a one year die-away experiment with four d ifferent soils, with and without spiking of the test materials. The four soils included sandy agric ultural soil; silty agricultural soil; clayey soil; and highly weathered oxide-rich soil. The anaerobic ally 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 from Georgetown, Delaware was 86 and 38 mg/kg dw in the years 2000 a nd 2002, respectively and from Wilmington was 43 and 22 mg/kg dw, respectively. Th e initial HHCB concentrations in spiked soil were 6 and 13 mg/kg soil, whereas the levels i n unspiked soil were 0.1 to 0.27 mg/kg. In all soils, the concentrations rapidly decreased. After one month, the concentrations of HHCB in the four soils were 30 to 90 percent of the initial con centration, and after 90 days, they ranged from 8 to 60 percent of the initial concentration. During the three months when the soil was frozen, concentrations of HHCB were stable. After o ne year, the residues ranged from below 10 to 14 percent of the initial concentration. The rat e of dissipation was higher in the soils with lower content of organic material. Loss processes c ould have included volatilization and leaching as well as biotransformation. The EC (EC, 2008) calculated half-disappearance times based on these data and reported values of 141 and 144 days for the spiked and unspiked biosolids-amended soils, respectively. In this stud y, leachates were also collected over 3 to 5 months, and the HHCB found amounted to 0.03 to 0.18 percent of the initial amount in the spiked soils. HHCB was not detected in leachate fr om unspiked soils (EC, 2008). The low percentage of HHCB initially in the sludge and pres ent in leachate was probably a result of both degradation and sorption.

Observed soil and sediment half-lives consistently exceeded 60 days (Table 2-6). Field
measurements on biosolids-amended soil indicated th at HHCB disappeared almost completely from soil within one year. The half-life based on u nfrozen conditions in biosolids-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 init ial concentrations. In the EU RAR (EC, 2008), a half-life of 105 days in the biosolids-amended so il was deemed most relevant for modeling the fate of HHCB in soil using the European Union S ystem for the Evaluation of Substances (EUSES) model, while 79 days was noted for the sedi ment (Envirogen, 1998; as cited in EC, 2008). EPA/OPPT agrees that these values are reason able for modeling and assessment

purposes.Table 2-6. HHCB Degradation Half-Lives and Half-Di sappearance Times for Environmental Media

Degradation Half-Life or Half-Disappearance Time a

Reference Water Sediment Soil Sludge Biosolids

Amended Soil

33 hours 21 hours (DT50) Langworthy et al. (2000 ) (as cited in EC, 2008)

43 hours 33

- 69 hours (DT50) Federle et al. (2002 ) (as cited in EC, 2008)

100 hours b (DT50) 10

- 15 hours (DT50) Schaefer (2005 ) (as cited in EC, 2008)

79 days 95 days

(forest soil) 105 days Envirogen (1998 ) (as cited in EC, 2008)
239 days (agricultural soil) Envirogen (1998 ) (as cited in EC, 2008)

141

- 144 days (DT50) DiFrancesco et al. (2004)

a DT50 is the time required for disappearance of 50% of the starting material. Half-disappearance time ( DT50) differs from degradation half-life in that it inclu des all mechanisms contributing to disappearance ( **e.g.** , volatilization in addition to degradation). Numbers that are DT50 values are indicated in the table; o therwise the data are degradation half-lives.

b Water with 10 mg/L suspended solids.

2.3.1.4 Fate in Air
Aschmann et al. (2001) studied the atmospheric oxidation (ph otooxidation) of HHCB and reported a rate constant for hydroxyl radical oxida tion 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-ra nge atmospheric transport is unlikely to be an important process for HHCB. Recent work of Villa et al. (2014) agrees, but also suggests that “moderate-range” (this could be called regional) at mospheric transport may occur. Any HHCB in the atmosphere is likely to be in the gas (not p articulate) phase. This is supported by
consistent results from three different models that predict extent of sorption to atmospheric particulates (aerosols) based on chemical propertie s. The fraction of HHCB sorbed to airborne particulates from AEROWIN v1.00 (EPI Suite v4.11) w ere 0.00149, 0.00288 and 0.00329 for the Junge-Pankow, octanol/air (K oa ) and Mackay models, respectively.

2.3.2 Bioaccumulation and Bioconcentration

Table 2-7 summarizes published bioaccumulation data for aquatic organisms, from Dietrich and Hitzfeld (2004) and Reiner and Kannan (2011).Fish

Dietrich and Hitzfield (2004) summarized informatio n on the bioconcentration and bioaccumulation potential of HHCB. BCFs and BAFs va ried with species but were generally lower than the commonly used criterion of 1,000 (lo g = 3) for characterizing bioaccumulation potential as moderate (EPA, 1999a; 1999b; Boethling et al., 2009). Measured BCFs of 620, 862 and 1,584 were reported for zebra fish, eel and blu egills, respectively (Table 2-7), whereas BAF values of 290, 510, 580, and 620 were reported for eel, tench, crucian carp and zebra mussels, respectively. These and other (chiefly monitoring) studies suggest that HHCB does not biomagnify.

A bluegill study was performed according to OECD Te st Guideline 305 and using Good Laboratory Practice (GLP) standards (EC, 2008; Rimk us, 1999). Bluegill sunfish ( **Lepomis**
**macrochirus** ) were exposed in a flow-through system to two conc entrations of radiolabeled HHCB. A solubilizer was used to prepare a solution (concentration 0.001 percent (w/v)). Identification of the parent compound in water and fish was performed using TLC/HPLC. Nominal exposure concentrations were 1 and 10 μg/L. The fish were exposed for 28 days; the elimination period was also 28 days. The eliminatio n rate constant (k 2 ) was estimated from the elimination curve (first-order kinetics), and deter mination of an uptake rate constant (k 1 ) was attempted using the following formula: k 1 = BCF × k 2 . However, this value could not be directly
calculated from the increase of concentrations in f ish due to rapid attainment of the final plateau level. Therefore, the BCF was derived from measured concentrations of parent compound in exposure water and the plateau concentr ation in fish. Based on the concentration of parent material, the BCF for whole fish was calc ulated to be 1,584 L/kg wet wt.

Reiner and Kannan (2011) and Nakata et al. (2007) s tudied bioaccumulation of HHCB in organisms from the upper Hudson River, USA; and mar ine tidal and shallow-water organisms from a Japanese site, respectively. In the first st udy, field (monitoring)-based BAFs were calculated as shown in Table 2-7). These and other (chiefly monitoring) studies suggest that HHCB is not subject to biomagnification, which occu rs when a substance achieves higher concentrations on a lipid-adjusted basis in organis ms than in their food (prey). Aquatic food chain modeling was conducted by the EPA using the B CFBAF © program (which uses the ArnotGobas model) in EPI Suite © v4.11. Consistent with the monitoring data, estima ted BAFs for middle and upper trophic fish were successively low er, not higher, than the estimated BAF for HHCB at the lowest trophic level.

Metabolism may account for the observation that mea sured BCFs and BAFs are lower than would be estimated based on the log K OW of HHCB (5.9). In this study, HHCB was metabolized to one or more polar metabolites in fish in a relative ly short time and then excreted to the water phase. Recent work of Fernandes et al. (2013) provi des evidence of metabolism of HHCB to a hydroxylated metabolite in European sea bass.Table 2

- 7 . BCFs and BAFs (L/kg ww) of HHCB in Aquatic Vertebrate s a
Organism Common Name BCF BAF
**Scardinius erythrophthalmus** Rudd 20

**Tinca tinca** Tench 510

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

**Danio rerio** Zebrafish 620

Smallmouth bass b 31

- 106
Largemouth bass b 30

- 146
White perch b 21

- 333

Catfish b 18

- 371

a source: Dietrich and Hitzfeld (2004) unless otherw ise noted

b source: Reiner and Kannan (2011)

Invertebrates
Bioaccumulation has been studied with two benthic o rganisms (Table 2-8). Non-biting midges ( **Chironomus riparius** ) exposed to 9.8 ± 1.4 μg/L, the concentrations in the organisms incre ased to a maximum level between days 1 and 3, and then t he level rapidly decreased to a new steady state with BCF values ranging from 85 to 138 L/kg wet wt. (Artola-Garciano et al., 2003). With the addition of piperonyl butoxide (PBO), a cy tochrome P450 inhibitor, the BCF was 525,
and it was concluded that the lower BCF values were likely caused by biotransformation of HHCB in this organism. In a flow-through experimen t with blackworms ( **Lumbriculus variegatus**  ), exposed to 4.6 ± 0.6 μg/L, uptake of HHCB reached a plateau after 3 days, and the measured BCF was 2,692 L/kg wet wt.

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

**terrestris** ) (Table 2-8). The steady-state concentration of a substance in earthworms is assumed to be proportional to the soil pore water concentra tion and a BCF can be calculated using an equation given by Jager (1998) (as cited in EC, 200 8):

BCF worm = (0.84 + 0.012 × **K** ow )/RHO earthworm

Where:

BCF worm = the earthworm bioconcentration factor (L/kg wet wt);

**K** OW = the n-octanol:water partition coefficient; and

RHO earthworm = the density of the earthworm set at 1 kg wet wt/ L.Table 2

- 8 . BCFs (L/kg ww) of HHCB in Benthic and Terrestria l Invertebrate s

Organisms 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

Plants Application of biosolids or reclaimed wastewater to agricultural land or crops may provide a pathway for uptake of HHCB by plants. Land applicat ion of reclaimed wastewater seems not to have been studied as a pathway for HHCB introductio n into the environment, but as reported in the EU 2008 assessment, in an unpublished study (Mü ller et al., 2002), transfer coefficients were determined for lettuce and carrots growing on biosolids-amended soil samples. The authors concluded that in spite of an unrealistical ly high concentration of HHCB in biosolids, no
relevant accumulation in leaves was observed and th at uptake ratios (the ratio C plant /C soil ) were
still well below the critical level of 1. Even for carrot root, transfer ratios were 0.1 and 0.48 in two different soils. Further, the low observed con centrations in the above ground parts of the carrot plant showed that there was no transport wit hin the plant. It was concluded that under
normal conditions transfer of HHCB from the soil to plants is not likely to be an important accumulation pathway.

2.3.3 Conclusions of Environmental Fate

HHCB is hydrophobic and partitions strongly to soli d phases. Therefore, HHCB is removed in WWTPs predominantly due to sorption to sludge and b iodegradation, but the degree of removal is considered suboptimal. HHCB has low pers istence in water and is moderately persistent in soil and sediment. Based on available data, HHCB is considered to be of low to
moderate concern for bioaccumulation. BCF values of 1,584 for bluegills and 2,692 for
**Lumbriculus** indicate moderate bioaccumulation potential. Howev er, BAF values are available for several aquatic organisms are in the range of 2 0 to 620, indicating low bioaccumulation. These studies, together with results of aquatic foo d-chain modeling (Arnot-Gobas model) and monitoring data for biota, suggest that HHCB is not subject to biomagnification.

## 3 ENVIRONMENTAL ASSESSMENT

3.1 INTRODUCTION

This assessment evaluated the environmental risk to the aquatic and terrestrial environments from the use of HHCB as a fragrance ingredient in c onsumer and commercial products. The HHCB environmental risk assessment is composed of t he following: (A) an exposure assessment; (B) an ecological hazard assessment; (C ) an environmental risk characterization; and (D) a discussion of uncertainties and data limi tations. Environmental risks were assessed by comparing levels of HHCB measured in the environmen t to the COCs determined from aquatic and benthic toxicity studies.



### 3.2 ENVIRONMENTAL EXPOSURE ASSESSMENT
For this assessment, environmental monitoring data consisting of measured levels of HHCB in wastewater, surface water, sediment, soil, and biot a were used to characterize environmental exposure to HHCB. The data for concentrations of HH CB measured in the environment in the US are from multiple sources, locations, and dates.

3.2.1 Estimated Environmental Releases Releases to the environment potentially occur throu ghout the life cycle of HHCB. In the US, the life cycle of HHCB consists of processing of import ed 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 product s, and disposal of HHCB-containing products to the environment. Compounding of fragran ce oils and formulation of end-use products occurs at industrial sites. A small amount of HHCB is released from these sites to the environment, including to water, mostly as a result of cleaning of compounding and blending equipment and cleaning of containers used to transp ort material. Specifically, the amount
released from industrial sites was estimated at a m aximum of 10 percent of the use volume (see Appendix E). EPA/OPPT assumed that the entire HHCB content of all of the various types of end-use products was released down-the-drain after use. Therefore, 90% of HHCB is assumed to be released to municipal wastewater after commer cial and consumer use of products containing HHCB.

3.2.2 Measured Levels in the Environment Studies published between 2004 and May 2012 that re ported measurements of HHCB in the US
environment were reviewed for this assessment. Dat a quality criteria included currency, geographic scope, accuracy/reliability, representat iveness, lack of bias, comparability and applicability. Details regarding the QA/QC of each individual study, laboratory reporting limits, and brief summaries of the studies from the publish ed literature are provided in Appendix G.Additionally, monitoring data in US waterways were collected from the USGS NWIS database up

to May 2012. All data collected from 2001-2012 cont ained in three USGS NWIS parameter codes (62075, 62823, 63209) for HHCB were included in this assessment. This database includes monitoring data as well as data collected for studi es of organic anthropogenic waste indicators, organic wastewater contaminants, and biologically-a ctive chemical mixtures in urban streams. USGS data from the NWIS database was accepted with the assumption that their internal methodologies were consistent and robust. These dat a were presumed to be collected under the guidance of the USGS National Field Manual for the Collection of Water-Quality Data, a publication which documents the methods, protocols, procedures and recommended practices for the collection of water-quality data (USGS, var iously dated). Data reporting procedures were presumed to follow USGS guidance (Oblinger Childres s et al., 1999).

Although ten times more data was available from the USGS as compared to data from published literature studies for effluent, surface water and sediment, the measured HHCB concentrations were within the same order of magnit ude. EPA/OPPT assumed that data from scientific publications were comparable to the USGS data and were used along with the USGS data as the basis of the exposure and risk characte rizations. Although sampling techniques varied depending upon the environmental media, gas chromatography/mass spectrometry (GC/MS) was the primary analytical methodology empl oyed in the detection and quantification of HHCB across all media. HHCB found in wastewater, surface water, sediment, biosolids, soil, and aquatic biota are summarized below and in Table 3-1. More detailed information regarding extraction, recovery, and reporting levels for the monitoring data is provided in Appendix G.Table 3-1. Measured Concentrations of HHCB in Biot a

Environmental Compartment Number of

samples (all US studies) Minimum Maximum Range of Reported Means Reference

Wastewater (μg/L)

Influent >135 0.043 12.7 0.42 – 9.03 Hope et al. (2012), Horii et al.(2007), Chase et al. (2012), Reiner et al. (2007a ), Osemwengie and Gerstenberger (2004) Effluent >135 0.010 13.0 0.02 – 5.31 Effluent discharged to ocean n/a ND e 2.5

--- d Anderson et al. (2012) Effluent (at stream or outfall) 47 0.090 a 3.4 b 0.98 – 1.18 USGS (code 62075 and 62823) Surface Water (μg/L)

Freshwater 94 ND f 1.6 0.00036 – 1.6 Osemwengie and Gerstenberger (2004), Anderson et al. (2012), Barber et al. (2011), Chase et al.

(2012)Peck and Hornbuckle(2004), Reiner and Kannan (2011), USGS (2011),
Klecka et al. (2010), Outfall, Stream, Lake, Reservoir, or Impoundment 4720 0.02 a 2.3 b 0.12 – 1.08 USGS NWIS (code 62075 and 62823) Groundwater (well) 1548 0.04 a 0.35 b 0.32 USGS NWIS (code 62075) Sediment (μg/kg dw)

Freshwater >59 ND e 388 1.1 - 144 c USGS ( 2008), Klecka (2010), Sapozhnikova (2010), Chase (2012), Reiner et al. (2011), Peck

et al. (2006) Bottom material (Lake, Resevoir, Impoundment, Stream) 606 17.68 a 212.87 b 67.99-87.46
USGS NWIS (code 63209)

Biosolids/Sludge (μg/kg dw) >58 13 427,000 13 - 177,000 c Kinney et al. (2006, 2008), Buyuksonmez et al. (2005), DiFrancesco et al. (2004) Soil (μg/kg dw)

Biosolids amended 6 1050 2770

--- d Kinney et al. (2008), Chase et al. (2012) Wastewater treated n/a <0.33 5.69

--- d Biota (Fish) μg/kg Fish (wet weight) 93 <1 5.4 3.0-4.8

Kannan et al. (2005), Reiner and Kannan (2011), Ramirez (2009), Kinney (2008) Fish (lipid weight) 23 <1 51.1

--- d Fish (tissue conc.) 30 n/a 2100 100 - 1800 Eel (lipid weight) 1 125 125

--- d Biota (other) ( μ g/kg ww)
Earthworm (tissue conc.) 9 ND g 3340

--- d Kinney et al.(2008) Zebra Mussel (lipid weight) 4 10.3 19.3

--- d Reiner and Kannan (2011) Shrimp , farmed (lipid weight) Shrimp, wild (lipid weight) 3 6 n/a

n/a 330

424

--- d Sapozhnikova et al. (2010 ) Marine mammal (wet weight) 25 <1 25 2.8 - 14 Kannan (2005) River Otter (wet weight) 3 2.4 3 2.8
Kannan (2005) Mink (wet weight) 4 2.2 5.3 3.7 Bird (wet weight) 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 e method detection limit not specified

f detection range =0.038-2.16 μg/L

g method detection limit =12.5 μg/kg dw3.2.2.1 Wastewater

In published studies, measured concentrations of HH CB in US WWTP influent and effluent varied widely. Influent values ranged from 0.043 to 12.7 μg/L and effluent values ranged from 0.010 to 13.0 μg/L (Table 3-1; for paired effluent/ influent values see also: Appendix G, Table G1). Effluent concentration was dependent on the typ e of waste treatment process employed and varied depending upon the season or month when the sample was collected; however, aerobically digested sludge reduced nitro and polyc yclic musk concentrations more rapidly than anaerobically digested sludge (Smyth et al., 2008). Reported values from a Canadian study confirmed that effluent values were process depende nt and concentrations were seasonally dependent and within an order of magnitude of those measured in the US (Smyth et al., 2008). Reported values summarized in the 2008 EU RAR were comparable to those found in the US, with one study from the UK that included a variety of treatment processes indicating a slightly wider range of influent concentrations (7.8-19.2 μg /L). The effluent values reported in the 2008 EU RAR were also comparable to those in the US, wit h the highest value (13.3 μg/L) reported for a study of 5 sewage treatment plants in Germany

.

The mean HHCB concentrations in wastewater effluent from stream and outfall discharge sites calculated from the USGS NWIS data (more than 40 da ta points) were 0.98-1.18 μg/L. The highest values were measured at outfall sites (95 th percentile: 3.4 μg/L). The mean effluent concentrations were comparable at stream and outfal l sites, but the range of concentrations was greater at outfall sites.

Very few studies have reported measuring metabolite s of HHCB other than HHCB-lactone, the apparent principal degradation product of HHCB. How ever, measured values of HHCB-lactone in the influent (maximum 1.15 μg/L) and effluent (m aximum 4 μg/L) of wastewater have been reported by Horii et al. (2007), and Reiner et al. (2007). In these studies, the range of
concentrations of HHCB-lactone in the effluent were more than twice the concentration in the influent and were reported to be comparable to thos e measured in Switzerland.

3.2.2.2 Surface Water In published studies, surface water concentrations of HHCB were found to range from nondetect (ND) to 1.6 μg/L and were dependent on their proximity to WWTP outfalls (Table 3-1; see also: Appendix G, Table G-2). The lowest repor ted concentrations were ND levels (reported detection range = 0.038-2.16 μg/L) at locations ups tream from WWTPs; higher concentrations
of HHCB were found downstream from the WWTPs. A max imum value of 1.6 μg/L was reported by Barber (2011) for a location at the North Shore Channel of the Chicago River at an area impacted by water reclamation plant effluent. These data support the assertion that the efficiency of HHCB removal in WWTPs plays a signifi cant role in the observed surface water

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

locations in 46 states), the mean calculated value for HHCB concentration in surface water was < 1.1 μg/L for all sites; the highest concentration s were measured at outfall sites. The 95 th percentile groundwater concentrations at well sites and surface water concentrations at streams and lakes, reservoirs, and impoundment site s were ≤0.35 μg/L (Table 3-1; see also: Appendix G, Table G-3 and Figures G-2 and G-3). Low er concentrations were consistently recorded in surface water samples collected from st reams and lakes, reservoirs, and impoundment sites when compared to values measured at outfall sites.

Values reported for surface water concentrations in the EU (EC, 2008) were similar to those found in the US, with the exception of areas of the water system in Berlin, Germany where very high proportions of effluents are present. The medi an concentration in sections with a high contribution of effluents was 1.48 μg/L.

3.2.2.3 Sediment Sediment concentrations varied from ND to 388 μg/kg dw in published studies (Table 3-1; 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 μg/kg dw) were found in a lake in Texas, a non-effluent impacted s ite. One of the highest concentrations (388 μg/kg dw) was reported along the upper Hudson River in New York, a river that receives treated wastewater discharge. Likewise, sediment along the Cuyahoga River in Ohio was found to have higher concentrations of HHCB in samples collected downstream of seven WWTPs (average= 144 μg/kg dw) when compared to those collected upst ream of the same WWTPs (average = 37 μg/kg dw). These studies suggest that effluents fro m WWTPs are a major source of synthetic musks that enter the environment and are present in the sediments of various water bodies. Reported sediment concentrations in the EU (EC, 200 8) and Asia (Lee et al., 2014) were similar to those in the US, with higher values found in Ber lin, Germany and in the sediment of contaminated brooks in Hessen, Germany.

Over 600 bottom material measurements (collected fr om USGS locations in 25 states) were available from the USGS NWIS (Table 3-1; see also: Appendix G, Table G-7 and Figure G-3). The

95 th percentile concentration at lake/reservoir/impound ment 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 concentrations (5%ile, mean, and 95%ile) were similar at both types of sites.

3.2.2.4 Biosolids and Soil Published concentrations of HHCB in biosolids range d from <100 to >100,000 μg/kg dw (Table 3-1; see also: Appendix G, Table G-4). These value s were dependent on a number of factors, including: season, date of collection, location, p opulation served, WWTP operations ( **e.g.** , municipal or industrial receiving waste stream, wat er volume, treatment type) and preparation methodologies prior to land application. Because th ese factors cannot be readily separatedfrom all values within the reported data set, it wa s not possible to determine which single

treatment variable had the largest impact on the re ported environmental concentrations. However, EPA observed that the US values for measur ements of biosolids were similar to those obtained in Ontario, Canada and likewise showed dif ferences depending upon location, treatment process, and season (Smyth et al., 2007). Reported concentrations of HHCB in sludge in the EU (EU RAR, 2008) were generally lower than those reported in the US.

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 am ended soils (Table 3-2; see also: Appendix G, Table G-5). The available data indicate that lan d application of treated waste water effluent or biosolids results in detectable quantities of HH CB in those soils.

3.2.2.5 Biota
The measured concentrations of HHCB in wildlife var ied by location, species, and method of reporting ( **e.g.** , lipid weight, tissue weight, or wet weight) (Tabl e 3-1; see also: Appendix G, Table G-6). Monitoring studies of aquatic biota wer e available only from the scientific literature, and the greatest number of sampling measurements wa s collected for fish (146 sampling measurements). On a lipid weight basis, measured va lues across various fish species and locations ranged from <1 to 51.1 μg/kg, whereas lev els in wild caught and farm raised shrimp were 330 and 424 μg/kg, respectively (Appendix G, T able G-6). Mean tissue weight concentrations of HHCB in different fish species ra nged from 100 to 1800 μg/kg (Appendix G, Table G-6). On a wet weight basis, HHCB has been re ported in a number of different species and trophic levels. The reported values were relati vely consistent across these species and ranged from <1 to 25 μg/kg (Appendix G, Table G-6).

3.2.2.6 USGS Data Analysis In addition to reported quantitative values, the US GS dataset includes values that are reported as less than the USGS laboratory reporting level 10 (LRL); data that are between the LRL and the long-term method detection level (LT-MDL); and data that are below the LT-MDL (Oblinger Childress et al., 1999). The value of the LRL is re ported with a “less than” remark code for samples in which the analyte was not detected. “Est imated” remark codes are noted for all values falling outside the calibration range becaus e of increased measurement uncertainty or values below the LT-MDL determined using informatio n-rich methods 11

.

For monitoring data sets where the geometric standa rd deviation was less than 3.0, values recorded as “less than LRL” or “estimated” were rep laced by the LRL divided by the square root of two, as per the EPA/OPPT’s guidance (EPA, 1994). Where the geometric standard deviation

10 The LRL for water sampling was 0.5 μg/L for samplin g dated 7/16/2001 – 9/30/2009 and was updated to 0. 05 μg/L for samples dating from 10/1/2009 to the prese nt.

11 The USGS defines **information rich methods** as “Classified as organic methods that use either mass spectrometric or photodiode array ultraviolet/visible spectroscop ic detection. These methods have qualifying inform ation that allows enhanced analyte identification.”was greater than 3.0, EPA/OPPT replaced values reco rded as “less than LRL” or “estimated”

with the LRL divided by two (EPA, 1994). This pract ice 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 concent rations and are biased towards the LRL.

High quality monitoring data (greater than 6800 dat a 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 sho wn in Table 3-1. 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 exp osure assessment.

3.2.3 Conclusions of Environmental Exposure HHCB has been detected and measured in wastewater, surface water, sediment, sewage sludge, soil, and aquatic biota in numerous studies . These data strongly suggest that HHCB is ubiquitous and potentially widespread in the enviro nment. However, these data reflect discrete locations and times, and the extent to whi ch they are representative of the overall
distribution of HHCB is not known.



### 3.3 ECOLOGICAL HAZARD ASSESSMENT

The environmental hazard assessment is based on pre vious hazard assessments including Balk and Ford (1999a, b); EU RAR (EC, 2008); HERA Projec t Report (HERA, 2004a); and Robust Summaries submitted under the US EPA HPV program (I FRA, 2003). In addition, a literature search was performed to identify peer-reviewed arti cles on ecotoxicity published between 2007 and May 2012. The search terms included freshw ater and saltwater fish, aquatic invertebrates, and aquatic plants; pelagic and bent hic organisms; acute and chronic sediment toxicity in freshwater and saltwater and terrestria l toxicity to soil organisms, birds, and mammals. The test species, test conditions, toxicit y endpoints, statistical significance, and strengths/limitations of the study were summarized and evaluated for data quality. Data quality inclusion criteria included: use of appropriate ana lytical and test controls, identification of test substance and test organism, stated exposure durati on time and administration route, a clear description of the effect endpoints, and transparen t reporting of effect concentrations. Guideline studies as well as studies using other pr otocols were included if they met data quality criteria. Specific criteria for exclusion are: stu dies that included HHCB as part of a mixture in
wastewater effluent or surface water and studies de scribed only in abstract form or in a language other than English.

Application of uncertainty factors based on establi shed EPA/OPPT methods (EPA, 2012f; 2013) were used to calculate lower bound effect levels (r eferred to as the concentration of concern; COC) that would likely encompass more sensitive spe cies not specifically represented by the available experimental data. Uncertainty factors ar e included in the COC calculation to account for differences in inter- and intraspecies variabil ity, as well as laboratory-to-field variability.These uncertainty factors are dependent upon the av ailability of datasets that can be used to

characterize relative sensitivities across multiple species within a given taxa or species group,
but are often standardized in risk assessments cond ucted under TSCA, since the data available for most industrial chemicals is limited (Ahlers et al., 2008).

A summary of the available ecotoxicity data for HHC B that were deemed adequate for consideration in this assessment are provided in ta bles and the studies selected for use in calculating risk quotients are described in more de tail in the appropriate section below.

3.3.1 Acute Toxicity to Aquatic Organisms Acute aquatic toxicity studies considered for this assessment are summarized in Table 3-2.

Additional acute toxicity values for were reported in Deitrich and Hitzfield (2004). For 96-hour toxicity tests using fathead minnow and zebrafish e mbryos, an EC 50 s of 0.39 mg/L and an LC 50 of >0.67, were reported, respectively. These data were not included in EPA’s consideration because they were reported in a secondary source an d sufficient study details were not provided in the report or found during the search f or information performed by EPA/OPPT. The zebrafish study appears to be described in the EU R AR (EC, 2008), but not used due to lack of experimental details.

The acute toxicity study in **Daphnia magna** (Wüthrich, 1996 as cited in Balk and Ford, 1999b and provided to EPA in IFRA, 2003) was selected fro m the available acute toxicity studies to calculate an RQ because **Daphnia** were the most sensitive species for acute toxic ef fects to HHCB. Although the lower end of the effects concen tration range for freshwater mussel is lower than the Daphnid effects concentration, the a uthors reported highly variable concentrations during the test as reflected by the effects concentrations being reported as a range, the upper end of which is approximately thre e times higher than the Daphnid effects
concentration. EPA notes that the differences in se nsitivity between freshwater and marine organisms appears, based on available data, to be l ess than an order of magnitude. Daphnid is a more representative species for this assessment bec ause the available monitoring data used for estimating exposure is largely freshwater. Further more, the study is reliable and demonstrates the acute effects (i.e. survival/immobilization) us ing appropriate, reproducible protocols.

The **Daphnia magna**  48-hour EC 50 of 0.282 mg/L was divided by an assessment factor (uncertainty factor [UF]) of 5 for invertebrates, a s per established EPA/OPPT methods (EPA, 2012f; 2013), to give an acute concentration of con cern (COC) of 0.0564 mg/L or 56.4 μg/L.Table 3

- 2 . Aquatic Toxicity Data for HHCB

- Acute Toxicity

Test Organism Fresh/

Salt

Water Test Guideline/ Study Type Duration

(hr-hour) Endpoint Concentration

(mg/L) Chemical

Analysis Effect References

Fish

- Freshwater

Bluegill sunfish **Lepomis macrochirus** Fresh OECD TG204/Flowthrough 96-hr a LC 50 0.452 Measured Mortality Wüthrich (1996) c

Aquatic Invertebrates

- Freshwater

Water flea ( **Daphnia**

**magna** ) Fresh OECD TG202/Semistatic 48-hr a EC 50 0.282 Measured Immobilization Wüthrich (1996) c

Freshwater mussel ( **Lampsilis cardium** ) Fresh Static b 48-hr glochidia LC 50 0.999 - >1.75

Measured Mortality Gooding et al. (2006) 96-hr juvenile EC 50 0.153 - 0.831 Growth

Aquatic Invertebrates

- Saltwater

Estuarine Copepods ( **Nitocra spinipes** ) Brackish SIS 1991/

Static 96-hr LC 50 1.9 Measured Mortality Breitholtz et al. (2003)

Marine copepod ( **Acartia tonsa** ) Salt ISO 1997/

Static 48-hr LC 50 0.47 Nominal Mortality Wollenberger et al. (2003)
Aquatic Plants

- Freshwater

Green algae ( **Pseudokirchneriella subcapitata**  ) Fresh OECD TG 201/Static 72-hr EC 50 >0.85 Measured Growth Van Dijk (1997) c 0.72 Biomass
Note: The shaded row indicates the principal study used for assessing acute risks to aquatic organism s.

a Value was calculated from the 21-day study

b Guideline not reported

c as reported in IFRA (2003), Balk and Ford (1999b) and EC (2008)

3.3.2 Chronic Toxicity to Aquatic Organisms

Chronic aquatic toxicity studies considered of acce ptable quality are summarized in Table 3-3.

Chronic toxicity values for rainbow trout ( **Oncorhynchus mykiss** ) and zebrafish ( **Danio rerio** ) with reported 21-day EC 50-repro and 21-day LC 50 of 0.282 mg/L and 0.452, respectively are reported

. However, these data were not included in EPA’s cons ideration because they were reported in a secondary source (Deitrich and Hitzfield, 2004) and sufficient study details were not provided in the report or found during the search for informati on performed by EPA/OPPT.

Chronic aquatic studies with HHCB are difficult to conduct in static or semi-static systems because the tendency of HHCB to associate with orga nic matter and surfaces may affect exposure concentrations. For example, the measured concentrations at the end of two chroniccopepod tests (Breitholtz et al., 2003; Wollenberge r et al., 2003) were only 2 to 19 percent of

the nominal concentration. Volatilization or sorpti on to organic material likely accounts for the disappearance of HHCB from the water phase. HHCB ha s been reported to have a half-life of hours to days in river water tests using radiolabel ed HHCB; HHCB may also be lost to the atmosphere via volatilization (Langworthy et al., 2000; Federle et al., 2002; as cited in EC, 2008). Additionally, HHCB will bind strongly to org anic material and sediment (average half-life of 128 days) (Envirogen, 1998; as cited in EC, 2008 ). Therefore, chronic toxicity values based on nominal (not measured) concentrations from otherwis e well-documented studies were not considered for this endpoint and the Concentration of Concern because the actual concentration of HHCB in the test system is unknown

.

Following adequacy review of the chronic tests summ arized in Table 3-3, the fish prolonged early life stage toxicity test (OECD TG 210) using fathead minnow ( **P. promelas** ) reported by
Croudace et al

**.** (1997) was selected as the study from which to cal culated a chronic RQ. The marine copepod ( **Acartia tonsa** ) study by Breitholtz et al. (2003) was selected in EPA’s draft assessment for HHCB; however, upon consideration of multiple comments regarding the use of this study for chronic RQ calculation, EPA/OPPT rec onsidered this selection. The marine copepod ( **Acartia tonsa** ) study by Wollenberger et al. (2003) was considere d less reliable than both the study of Breitholtz et al. (2003) and the fathead minnow study due to the fact that
concentrations of HHCB were not measured in the stu dy and due to other test design limitations described in the EU RAR (EC, 2008). Alt hough EPA notes that the differences in sensitivity between freshwater and marine organisms appears, based on available data, to be less than an order of magnitude, the fathead minnow is a more representative species for this assessment because the available monitoring data us ed for estimating exposures is largely freshwater. Furthermore, the study is reliable and demonstrates the chronic effects (i.e., survival and growth) using appropriate, reproducibl e protocols.

The study was conducted using a flow-through system , wherein fathead minnow eggs (<24 hours old) were exposed to nominal concentrations r anging from 0.0125 to 0.2 mg/L for 36 days. Corresponding mean measured concentrations (m easured via HPLC or GC 13 times during the test) were: 0.0091, 0.019, 0.037, 0.068 and 0.1 40 mg/L. Egg hatchability was not affected at any concentration. HHCB did have an effect on larva l survival and growth at 0.140mg/L. Compared to controls, larvae exposed to 0.140 mg/L experienced a 20 and 54% reduction in mean length and weight, respectively. Survival was reported to be 78% and of those that survived, larvae were generally smaller, underdevel oped and displaying erratic swimming behaviors. The authors identified 0.068 mg/L as th e NOEC for the study (based on survival and growth); the LOEC would be 0.140 mg/L.

EPA/OPPT calculated at a Maximum Acceptable Toxican t Concentration (MATC) effect concentration of 0 0.097 mg/L for survival. To der ive a chronic COC, the MATC was divided by an assessment factor (UF) of 10, according to estab lished EPA/OPPT methods (EPA, 2012f; 2013), to yield a chronic COC of 0.0097 or 9.7 mg/L for survival for aquatic organisms.Table 3

- 3 . Aquatic Toxicity Data for HHCB

- Chronic Toxicity

Test Organism Fresh/

Salt

Water Test Guideline/ Study Type Duration Endpoint Concentration (mg/L) Chemical

Analysis Effect References

Fish

- Freshwater Bluegill sunfish ( **Lepomis**
**macrochirus** ) Fresh OECD TG 204/Flow Through 21-day NOEC 0.093 Measured Clinical Signs Wüthrich (1996) d LOEC 0.182

Fathead

minnow ( **Pimphales promelas**  ) Fresh OECD TG 210/Flow Through 36-day NOEC 0.068

Measured Survival
Growth Development Croudace et al.

(1997b) c LOEC 0.140

MATC 0.097

Aquatic Invertebrates - Freshwater

Water flea ( **Daphnia**

**magna** ) Fresh OECD TG 202/

Semi-static 21-day NOEC 0.111

Measured Reproduction Wüthrich (1996) d LOEC 0.205

EC 50 0.293 Immobilization Aquatic Invertebrates - Saltwater

Marine copepods ( **Acartia tonsa** ) Salt OECD TG
Draft Invert Life Cycle/

Static 6-day EC 10 0.044

Measured Development Bjornestad (2007) b EC 50 0.131

NOEC 0.038

LOEC 0.075

5-day EC 10 0.037 Nominal Development Wollenberger et al. (2003) EC 50 0.059

Estuarine copepods ( **Nitocra spinipes**  ) Brackish ___ a /Static
Renewal 22-day NOEC 0.007

Measured Development Breitholtz et al. (2003) LOEC 0.02

Aquatic Plants - Freshwater
Green algae ( **Pseudokirch**
**neriella subcapitata**  ) Fresh OECD TG 201/Static 72-hr NOEC 0.201

Measured Growth

Biomass Van Dijk

(1997) c LOEC 0.466

Note: The shaded row indicates the principal study used for assessing chronic risks to aquatic organi sms. LOEC = Lowest Observed Effect Concentration
NOEC = No Observed Effect Concentration MATC = Maximum Acceptable Toxicant Concentration

a Test guideline and/or test type not reported

b As reported in EC (2008)

c As reported in Balk and Ford (1999b) and EC (2008)

d As reported in IFRA (2003); Balk and Ford (1999b); and EC (2008)3.3.3 Toxicity to Sediment-Dwelling Organisms

Toxicity studies in sediment-dwelling organisms inc luding amphipods, midges, oligochaete worms, polychaete worms, and mud snails are summari zed in Table 3-4.

The Amphipod, **Hyalella azteca,** was selected for calculating an RQ because it is a wellestablished test species for evaluating the effects of chemicals in the sediment environment and is widely distributed across North America (Pen nak, 1978). The study was conducted using established EPA and OECD test guidelines for measur ing lethal and sublethal effects of chemical exposure (i.e., growth, survival, and reproduction) . Although the sediment-dwelling species
identified as most sensitive in Table 3-4 is the Ne w Zealand mud snail, it is an invasive species in the US and currently is not ubiquitously distribute d throughout the US. Toxicity data are also available for the polycheate ( **Capitella sp.** ) for relevant endpoints of growth and reproduction
and the effects concentrations are similar to those for **Hyallela** . Although EPA notes that the differences in sensitivity between freshwater and m arine organisms appears, based on available data, to be less than an order of magnitu de, the freshwater **Hyalella** is a more representative species compared to **Capitella** for freshwater. Furthermore, the study is reliable and demonstrates the chronic effects (i.e., surviva l and growth) using appropriate, reproducible protocols.

Toxicity of HHCB to **H. azteca** was evaluated using five nominal concentrations of 0, 6, 14.5, 35, 83, 200 mg/kg and the measured concentrations were on average 49% of the nominal (i.e., 0, 3, 7.1, 16.3, 41 and 98 mg/kg, respectively) (Egeler, 2004 as cited in EC, 2008). Ten animals 7 to 14 days old (four replicates/concentration) were expos ed for 28 days. A solvent control was used in the test. There were no mortalities at 35 mg/kg (measured 16.3 mg/kg) during the study. At day 28, mortality was 88% at 83 mg/kg (measured 40. 67 mg/kg) and 100% at 200 mg/kg (measured 98 mg/kg). Growth (decreased length) was 9% below the control and total biomass decreased by 15% per replicate at 35 mg/kg (measure d 16.3 mg/kg). The reported LOEC was 35 mg/kg (measured, 16.3 mg/kg) and NOEC was 14.5 mg/k g (measured, 7.1 mg/kg).Table 3

- 4 . Sediment Toxicity Data for HHCB

Test Organism Fresh/

Salt Water Test

Guideline/ Study Type Duration EndPoint Concentration (mg/kg dw) Test

Analysis Effect Reference

Freshwater
Amphipod ( **Hyalella**

**azteca** ) Fresh OECD 218/
Renewal 28-day NOEC 7.1 b

Measured Growth Egeler (2004) c LOEC 16.3 b

MATC 10.8 Midge ( **Chironomus riparius**  ) Fresh OECD 218/
Renewal 28-day NOEC 200 b

Measured Emergence Egeler & Gilberg (2004a) c LOEC 400 b

EC 50 335 b Development Oligochaete ( **Lumbriculus variegatus**  ) Fresh OECD 218/
Renewal 28-day NOEC 16.2
Measured Reproduction Egeler & Gilberg (2004b) c LOEC 36.5

EC 50 44.5 Reproduction

Saltwater

New Zealand
mud snails ( **Potamopyrgus antipodarum**  ) Fresh/

Brackish Renewal a 120-day NOEC 0.81

Measured Reproduction Pedersen et al. (2009) LOEC 7.0

NOEC 7.0 Juvenile Growth LOEC 19.3

Polychaete ( **Capitella sp.** ) Salt Renewal a 120-day NOEC 1.5

Measured Reproduction Ramskov et al. (2009) LOEC 26

NOEC 26 Juvenile Survival LOEC 123 Note: The shaded row indicates the principal study used for assessing chronic risks to sediment-dwell ing organisms.

a Test guideline not reported

b Concentration adjusted based on measured concentra tions (recovery)

c as reported in EC, 2008

EPA/OPPT calculated at a Maximum Acceptable Toxican t Concentration (MATC) effect concentration of 10.8 mg HHCB/kg dw from this study . To derive a chronic COC for sedimentdwelling organisms, the MATC was divided by an asse ssment factor (UF) of 10, according to established EPA/OPPT methods (EPA, 2012f; 2013), to yield a chronic COC of 1.08 mg/kg dw.

3.3.4 Toxicity to Terrestrial Organisms The toxicity data for terrestrial organisms conside red for this assessment is summarized in
Table 3-5.

Invertebrates The earthworm ( **Eisenia fetida** ) was the most sensitive terrestrial (soil) inverte brate species tested (Chen et al., 2011a). In a two part test, Ch en et al. (2011a) observed the lethality of earthworms exposed to HHCB in a 14-day study and th e chronic effects in a 28-day study. In the acute study, adult earthworms were exposed for 14 d ays to concentrations of 0, 100, 140, 196, 274.4, 384.2, 537.8, 752.9, 1,054.1, and 1,475.8 mg HHCB (99 percent purity)/kg (air-dried soil) in acetone (Chen et al., 2011a). Observations were taken on days 7 a nd 14. No mortalitiesoccurred at the lowest concentrations of 100 and 14 0 mg/kg. However, there was 100%

mortality at ≥ 1,054.1 mg HHCB /kg. The 7-day LC 50 was reported as 489 mg HHCB/kg and the 14-day LC 50 as 392.4 mg HHCB/kg.

In the 28-day chronic test, earthworms were exposed to 3, 10, 30, 50, and 100 mg HHCB (99 percent purity)/kg (air-dried soil) in acetone. Aft er 28-days of exposure to HHCB there was a significant decrease in growth rate of earthworms w hich occurred at 100 mg HHCB /kg. Additionally, a significant decrease in reproductio n rate and cocoon production in earthworms exposed to 50 and 100 mg/kg were noted. A LOEC of 5 0 mg/kg was reported; the NOEC is 30 mg/kg. A MATC of 38.7 mg/kg was calculated for the earthworm. Because exposure data are
insufficient to assess HHCB risk to terrestrial inv ertebrates, EPA/OPPT did not calculate a COC
for the earthworm.

Plants For plants, two studies (An et al., 2009 and Chen e t al., 2010; summarized in Table 3-5) were located that evaluated the toxicity of HHCB to whea t growth and development. However, these studies were not conducted using EPA, OECD or other established or widely recognized protocols, thereby making it difficult for EPA/OPPT to evaluate their adequacy or the relevance of the endpoints measured. This lack of adequate to xicity data for terrestrial plants coupled with only limited soil concentration data and exter nal expert advice, does not support assessing risks to terrestrial plants and hence, COCs were no t derived from these studies.Table 3

- 5 . Soil Toxicity Data for HHCB

Test Organism Duration

hr=hour Test
Guideline /Study Type Endpoint Value Unit Test Analysis Effect Reference

Terrestrial Organisms

Springtail ( **Folsomia**
**candida** ) ISO/CD

11267 4-week NOEC 45 mg/kg dw Nominal Mortality and reproduction Klepka

(1997) a LOEC 105

Earthworm ( **Eisenia fetida**  ) ISO 11268-

1/OECD 207/Not
Refreshed 8-week NOEC 45 mg/kg dw Nominal Reproduction Gossmann (1997) a LOEC 105

OECD 207/Not reported 24-hr LC 50 32.60
μg/cm 2 not reported Mortality Chen et al. (2011b) 48-hr LC 50 11.87

72-hr LC 50 6.14

OECD 207/222*/

Not reported 7-day LC 50 489

mg/kg not reported Mortality

Chen et al. (2011a) 14-day LC 50 392.4

28-day NOEC 30
Reproduction

LOEC 50

MATC 38.7

Nematode ( **Caenorhab ditis elegans**  ) Not

reported 24-hr LC 50 >194.6

mg/L Nominal Mortality

Mori et al. (2006) 60-hr LOEC 9.8 Reduced growth
and maturation

72-hr LOEC 19.5 Decreased fecundity

Wheat ( **Triticum**
**aestivum** )

Not

reported/
Renewed Contingent

on control EC 50 143.4 mg/L Not reported Shoot elongation

An et al. (2009) 422.3 Root elongation

Contingent
on control LOEC 50 mg/L Not reported Seedling development – shoot height

21-Day LOEC 0.2 mg/L Not reported Reduced Chlorophyll synthesis

Not reported Contingent

on control NOEC 77

mg/kg dw Not reported** Reduced seed

germination Chen et al.

(2010) LOEC 194

LC 50 846.6

EC 50 2,123

EC 50 928.5

a As cited in Balk and Ford (199b) and EC(2008)

* Slightly modified guideline reported in the study **HHCB reported at 77.4% pure (Chen et al. 2010) LOEC=Lowest Observed Effect Concentration
NOEC=No Observed Effect Concentration MATC=Maximum Acceptable Toxicant Concentration3.3.5 Conclusions of Environmental Hazard Assessment

Ecotoxicity studies for HHCB have been conducted in fish, aquatic invertebrates, aquatic plants, sediment invertebrates, soil invertebrates, and ter restrial plants. HHCB was found to have high acute toxicity and moderate to high chronic toxicit y to aquatic organisms. The COCs derived for aquatic, soil, and sediment-dwelling organisms are summarized in Table 3-6.

Table 3

- 6

. Concentrations of Concern (COCs) for Environmental Toxicity

Environmental Toxicity Concentration of Concern (CO C)

Acute Toxicity, aquatic organisms 56.4 μg/L

Chronic Toxicity, aquatic organisms 9.7 μg/L

Chronic Toxicity, sediment-dwelling organisms 1.08 mg/kg dw (1080 μg/kg dw)



### 3.4 ENVIRONMENTAL RISK CHARACTERIZATION

3.4.1 Calculation of Risk Quotient (RQ) Values The goal of this environmental risk characterizatio n was to determine whether there are risks to the aquatic and benthic environments from the me asured levels of HHCB found in wastewater, surface water, or sediment. The data fo r environmental monitoring and toxicity
were used in this risk assessment to address the fo llowing key questions:

1. Does acute exposure to levels of HHCB measured in w astewater effluent in the US pose risks for adverse effects in aquatic invertebrates, fish, or plants?

2. Does chronic exposure to levels of HHCB measured in surface water in the US pose risks for adverse effects in aquatic invertebrates, fish, or plants?

3. Does chronic exposure to levels of HHCB measured in sediment in the US pose risks for adverse effects in sediment-dwelling invertebrates?

To answer these key questions, environmental risk w as characterized by calculating risk quotients or RQs (EPA, 1998; Barnthouse **et al.** , 1982); the RQ is defined as:

RQ = Environmental Concentration / Effect Level

An RQ equal to 1 indicates that the exposures are t he same as the concentration that causes effects. If the RQ is above 1, the exposure is grea ter than the effect concentration. If the RQ isbelow 1, the exposure is less than the effect conce ntration. The Concentrations of Concern

(COCs) for aquatic and benthic organisms shown in T able 3-6 and the environmental
concentrations shown in Table 3-7 were used to calc ulate RQs. The environmental concentration for each compartment ( **i.e.** , wastewater, surface water, sediment,) was
determined based on measured concentrations of HHCB (Table 3-1). RQs were not calculated for terrestrial invertebrates or plants due to insu fficient data for environmental concentrations.

Frequency, timing, and duration of exposure affects the potential for adverse effects in aquatic or benthic organisms. Aquatic exposure is assumed t o 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 measu red in surface water downstream from WWTPs even though some fate studies indicate that a significant percentage of HHCB may be lost to volatilization and metabolism. Exposure of sediment dwelling-organisms to HHCB is also
assumed to be continuous.

In order to characterize the degree of potential ri sk due to HHCB exposure, RQs were calculated
based on measured environmental concentrations for the maximum published levels and range of means from published studies of surface water an d sediment, and also for the mean and 95 th percentile for surface water at outfall sites and b ottom material from the USGS NWIS data (see Table 3-1 and Appendix G, Table G-8). Surface water concentrations of HHCB at outfalls are assumed to represent the maximum environmental aqua tic concentrations for HHCB that would be experienced acutely in streams that receiv e wastewater.

Table 3

- 7 . Environmental Concentrations Used to Calculate R Qs
Maximum
Published
Concentration Range of Mean

Published
Concentrations 95 th Percentile

USGS NWIS Mean

USGS NWIS

Surface Water (μg/L) 1.6 0.36 – 1.6 2.3 1.08

Sediment (μg/kg dw) 388 1.1 - 144 213 87

The calculated RQs are presented in Table 3-8 and r eflect the range of concentrations of HHCB
measured in the environment. In addition to the ma ximum published surface water values, the acute and chronic RQs for aquatic organisms in surf ace water were also calculated from the 95 th percentile USGS NWIS data using the concentrations measured in surface water at outfall sites. Though the concentration of HHCB in effluent discha rge may be diluted in surface water downstream from the WWTP outfall, the calculated RQ s for chronic risk to aquatic organisms
are considered conservative estimates based on actu al surface water concentrations.Table 3

- 8 . Calculated Risk Quotients (RQs) for HHCB

Calculated RQs Maximum

Published
Concentration Range of Mean

Published
Concentrations 95 th

Percentile

USGS NWIS Mean

USGS NWIS Range of RQs

Acute RQ
, aquatic organisms
Surface Water 0.03 0.006 to 0.03 0.04 0.02 0.006 to 0.04

Chronic RQ
, aquatic organisms,
Surface Water 0.16 0.04 to 0.16 0.24 0.11 0.04 to 0.23

Chronic RQ
, sediment-dwelling organisms 0.36 0.001 to 0.13 0.20 0.08 0.001 to 0.36

The range of RQs for acute toxicity to aquatic orga nisms was 0.006 to 0.04 (Table 3-8). The highest RQ of 0.04 was calculated using the 95 th percentile value from the USGS NWIS dataset
for HHCB in surface water as measured at an outfall site (see Table 3-1 and Appendix G, Table G-8). All RQs for acute toxicity to aquatic organi sms in Table 3-8 range from 25 to 167 times or 1 to 2 orders of magnitude below 1.

The range of RQs for chronic toxicity to aquatic or ganisms was 0.04 to 0.23 (Table 3-8). The highest RQ of 0.23 was calculated using the 95 th percentile value from the USGS NWIS dataset
for HHCB in surface water as measured at an outfall site (see Table 3-1 and Appendix G, Table G-8). All RQs for chronic toxicity to aquatic orga nisms in Table 3-8 range from 4 to 25 times or an order of magnitude below 1.

The maximum concentration for surface water as meas ured at an outfall site from the USGS dataset is 4.4 μg/L, a value well below both the ch ronic and acute concentration of concern for aquatic organisms.

The range of RQs for sediment-dwelling organisms wa s 0.001 to 0.36 (Table 3-8). The highest RQ of 0.36 was calculated using the maximum publish ed level of HHCB in sediment of 388 μg/kg dw (Table 3-1). This HHCB sediment concentra tion is from a single sample taken from the upper Hudson River near Troy, NY (Reiner, 2011) and a second sample (from the same study) taken near Albany, NY was 351 μg/kg. Both of these values are above the 95 th percentile value from the USGS monitoring data for bottom mate rial. It should be noted that of the more than 600 bottom material samples available in the U SGS dataset, a single value was identified as greater than the COC of 1080 μg/kg dw. This val ue is presumed to represent a “hotspot” as >99.9% of the remaining values were below the COC f or sediment dwelling organisms. Aside from this single value, all RQs for acute toxicity to aquatic organisms in Table 3-8 range from 3 to 1000 times or up to 3 orders of magnitude below 1.Insufficient exposure data were available to calcul ate risk for terrestrial invertebrates and

plants **via** contact with contaminated water, sediment, or soil

.

3.4.2 Key Sources of Uncertainty and Data Limitations The strength of this assessment is the calculation of risk based on a considerable quantity of surface water and sediment monitoring data and an a dequate ecological toxicity data set. However, a number of limitations in the approach cr eate uncertainties associated with this

assessment.

3.4.2.1 Representativeness of Exposure Concentrations Available monitoring data for HHCB in water, sedime nt, and soil were used in the assessment. A robust dataset of over 6800 sampling measurements were available from the USGS NWIS for wastewater effluent, surface water, and bottom mate rial along with over 500 sampling
measurements from the scientific literature. The da ta from both published studies and the
USGS NWIS database were used because the mean value s for surface water and sediment were within the same order of magnitude. The concentrati ons reported in the literature and in the USGS NWIS data are assumed to represent reasonable estimates, however the data may not
reflect the actual distribution of HHCB in the envi ronment. Calculation of the RQ based on the maximum value, the 95 th percentile, and the highest mean value was perform ed in an effort to capture the full range of concentrations in the env ironment.

3.4.2.2 Variability in Environmental Concentrations Although calculation of the RQ based on the maximum reported value from the literature, the

95 th percentile USGS NWIS value for surface water at ef fluent sites, and the highest mean value was performed in an effort to capture the full rang e of concentrations observed in this data set, higher values may exist in the environment. Variati on of use practices across the population and differences in sewage treatment plant inputs (i ndustrial, commercial or residential), removal efficiencies and receiving streams, may aff ect the final concentrations present in any single location or point in time. Unless actual env ironmental concentrations of HHCB in surface water or sediment are greater than one or two order s of magnitude higher than expected,
additional measurements or estimates will not affec t the risk calculation.

Use of all available data obscures trends such as e ffects of seasonality, differences in WWTP removal efficiencies, and differences in use trends ; therefore the ability in this assessment to predict or attribute concentrations to any particul ar variable was not possible. WWTP removal efficiencies for HHCB are reported to vary from 58 to >99%, therefore environmental concentrations may vary to a similar extent dependi ng upon other factors such as input and receiving environment. Well-designed studies that f ollow these factors over a period of time are needed to provide more definitive understanding of how these variables may be
interrelated or interact to effect final concentrat ions in any given media or location.3.4.2.3 Anaerobic Degradation

Waste (settled) activated sludge in WWTPs is genera lly sent to anaerobic digesters to reduce sludge volume and organic load before disposal of t he concentrated material. Removal of HHCB
under anaerobic treatment conditions has been addre ssed in several studies. Xue et al. (2010) reported poor overall removal of HHCB in a full-sca le anaerobic/anoxic/aerobic system combined with membrane bioreactor for municipal was tewater reclamation, and little removal was attributed to biodegradation. In contrast Carba lla et al

**.** (2007) reported good removal of HHCB (60-80%) in mesophilic and thermophilic anaero bic sludge digestion. Kupper et al. (2006) observed apparent anaerobic degradation of HHCB in the sludge line of a full-scale treatment system, but the mechanism was not confirmed. Theref ore, no direct evidence of anaerobic (methanogenic) biodegradation of HHCB exists, and t his is therefore an uncertainty in characterizing biodegradation.

3.4.2.4 Volatilization As noted in the document, volatilization seems not to be a major process in activated sludge treatment, although available data are limited. EPI Suite’s WWTP model (STPWIN) predicts negligible removal due to volatilization. Two studi es are cited in the section on fate in water, and they may appear at first glance to be in confli ct, however the amount of volatilization in lab studies versus that in the environment is highly de pendent on the test conditions. For lab studies, specifically, values will be dependent on design of test vessels, degree of mechanical aeration if any; and incubation temperature.

3.4.2.5 Isomers and Metabolites Information (hazard and exposure) of HHCB isomers a nd metabolites is very limited. The measurement of specific isomers of HHCB were not de scribed in the collected studies or USGS data set reviewed for this assessment and only a ve ry limited number of studies described the
occurrence of HHCB-lactone. Hazard as it relates to specific isomers of HHCB is also unknown.
More information on isomers and metabolites would b e needed to refine this assessment and provide a fuller understanding of if or how specifi c HHCB isomers or metabolites contribute to
the observed HHCB behavior in environmental media a nd their impact on organisms present in water, sediment and soil.

3.4.2.6 Deriving Concentrations of Concern from Single Spec ies Tests Ecological hazard of industrial chemicals is routin ely evaluated using a single species and within a controlled laboratory setting in the form of a to xicity test. There is uncertainty associated with extrapolating these single-species laboratory test results to concentrations intended to be protective of all species and environments. Applica tion of uncertainty factors (or assessment factors) based on established EPA/OPPT methods (EPA , 2012; 2013) were used to calculate lower bound effect levels (referred to as the conce ntration of concern; COC) that would likely encompass more sensitive species not specifically r epresented by the available experimental data. Uncertainty factors are included in the COC c alculation to account for differences in interand intraspecies variability, as well as laboratory -to-field variability. These uncertainty factors are dependent upon the availability of datasets tha t can be used to characterize relative sensitivities across multiple species within a give n taxa or species group, but are oftenstandardized in risk assessments conducted under TS CA, since the data available for most

industrial chemicals is limited (Ahlers et al., 200 8).

3.4.2.7 Assessment of Risk to Terrestrial Invertebrates or Plants Two studies with six total samples reported HHCB co ncentrations in soil following land application of either biosolids or wastewater efflu ent. The locations are limited to two sites and may not representative of overall US soil concentra tions given differences in sludge application practices, WWTP contaminant removal, or seasonal fa ctors. Measured concentrations of HHCB in non-amended soil were not available for comparis on. The limited data available on measured
levels of HHCB in biosolids or wastewater-amended s oil, prevented quantitative assessment of risk to terrestrial invertebrates and terrestrial p lants.

3.4.3 Conclusions of Risk Characterization
RQs were calculated based on measured values for th e maximum published levels and range of means from published studies of surface water and s ediment, and also for the mean and 95 th percentile for surface water at outfall sites and b ottom material from the USGS NWIS data in an effort to capture the spectrum of environmental con centrations. Uncertainty factors were used to account for inter-species and lab-to-field extra polation of hazards for organisms exposed
HHCB in the ambient environment. Risk estimates for terrestrial environments were not performed due to very limited monitoring and toxici ty data.



### 3.5 CONCLUSIONS OF ENVIRONMENTAL ASSESSMENT

This assessment evaluated the environmental risk to aquatic organisms from the use of HHCB as a fragrance ingredient in consumer and commercia l products. Scenarios of interest included
release of HHCB in wastewater to surface water or a pplication of biosolids to land. The assessment endpoints were acute and chronic toxicit y to aquatic organisms and chronic toxicity to sediment organisms. The risk assessment consider ed only direct exposure to aquatic organisms and sediment-dwelling organisms.

Acute and chronic RQs indicated that maximum measur ed surface water concentrations did not exceed COCs for aquatic organisms. The chronic COC was not exceeded for sediment-dwelling organisms at the upper range of measured environmen tal concentrations for the maximum published values or the more than 600 sediment meas urements from the USGS. Low risk
concerns were identified for both acute and chronic toxicity to aquatic organisms and for sediment-dwelling organisms. Therefore, unless envi ronmental concentrations increase by factor of 1 to 2 orders of magnitude, the assessmen t does not indicate risk concerns.

Insufficient data were available to calculate chron ic risk for terrestrial invertebrates and plants

**via** contact with contaminated water, sediment, or soil

.

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## APPENDICES

## 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 E U RAR on HHCB (EC, 2008). Other reviews are also available (OECD, 2009; EPA, 2003; HERA, 20 04a; 2004b). This review is not intended to be an exhaustive discussion of the toxicity of HHCB , but rather a summary of the available data. The reader is referred to the original documents fo r more detailed information.



### A-1 Human Hazard Characterization

A-1-1 Toxicokinetics (Absorption, Distribution, Metabolis m,

Excretion)

Available toxicokinetics data indicate that HHCB ha s poor dermal absorption and is extensively metabolized and excreted, with no evidence of signi ficant bioaccumulation. An **in vivo** human study reported approximately 0.1 percent dermal abs orption following exposure to HHCB (Hawkins et al., 1996a; Ford et al., 1999; as cited in EC, 2 008), while an **in vitro** dermal absorption study with human epidermal membranes sho wed 5.2 percent absorption (Green and Brian, 2001; as cited in EC, 2008). Human biomo nitoring data show that HHCB has been detected in blood, milk, and adipose tissue followi ng dermal exposures (see Table A-2). Data in humans after oral and inhalation exposures, includi ng 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, kidney s, and thyroid, with excretion primarily in the feces (Ford et al., 1999; Hawkins et al., 1995; as cited in EC, 2008). Intravenous st udies in rats and pigs indicated that HHCB is rapidly distributed and excreted primarily in the feces in the rat and in the urine in pigs (Hawkins et al., 1997; Hawkins, 1997; as cited in EC, 2008). No evidence
of accumulation in the feces or urine was seen in e ither the rat or pig. Apparent half-lives of elimination in the blood were longer in the rat tha n in the pig, with clearance from fat and skin considerably slower. HHCB and metabolites were foun d 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.



### A-1-2 Acute Toxicity
Several acute toxicity studies in rats and rabbits by the oral and dermal routes with HHCB are available (Minner and Foster, 1977; Moreno, 1975; a s cited in EC, 2008). Acute toxicity studies by the inhalation route are not available. The acu te 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.

### A-1-3 Subchronic/Repeated-Dose Toxicity

An oral range-finding study in rats with mean achie ved intakes up to 829 mg/kg-bw/day reported signs of liver toxicity at 347 mg/kg-bw/da y (lowest-observed-adverse effect concentration [LOAEL] (Api and Ford, 1999; Hopkins et al., 1996; as cited in EC, 2008); however, the follow-up 13-week repeated-dose oral (dietary) toxicity study showed no adverse, treatment-related effects on any parameters measure d. The no-observed-adverse-effect level (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 s tudy, no effects on organ weight or histopathology were reported in the reproductive organs at any dose level. Repeated-dose subchronic toxicity studies by the inhalation and d ermal routes were available, but were considered to be of limited value because they were not conducted according to guidelines or GLPs or they used fragrance mixtures in which HHCB was only present at low levels.



### A-1-4 Reproductive Toxicity and Fertility
No combined repeated-dose/reproductive/developmenta l or multigenerational reproductive toxicity studies on HHCB are available. Information on reproductive organs is available from the repeated-dose dietary toxicity study in rats; this study did not show any signs of toxicity to the reproductive organs examined at doses as high as ~1 50 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/artic le/safety-guidelines.html ) was conducted in rats in order to assess the potential for adverse effects to the neonate following exposure to HHCB through nursing. In this comprehen sive study, the basis for the dose selection for this study was a pharmacokinetic anal ysis by Hawkins et al. (1996b) (as cited in EC, 2008) that determined the oral doses required to pr oduce 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 mil k of lactating dams at 2.28 and 32.8 mg/L, respectively (see Human Biomonitoring below f or comparison to levels in human breast milk). Five groups of pregnant Sprague-Dawley rats (28/dose) were administered 0, 2, 6, or 20 mg/kg-bw/day HHCB (>95 percent purity) by gavage fr om gestation day (GD) 14 through weaning on lactation day 21 (Ford and Bottomley, 19 97; Jones et al., 1996; as cited in EC, 2008). Dams were allowed to litter and rear their offsprin g to weaning. F 1 offspring 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 measure d in the lactation study, actual intake of HHCB could not be determined.

After parturition, the offspring were counted, sexe d, weighed, and examined for external abnormalities. On postnatal day 4, the pups were we ighed and the litters were culled to four males and four females. During the pre-weaning peri od, pups were examined for several developmental milestones, including surface rightin g reflex, startle reflex, air righting reflex, and pupil reflex. Selected offspring from these lit ters (24/sex/dose) were allowed to reach sexual maturity and were then assessed for behavior al changes in motor coordination andbalance, activity, and avoidance. When the F 1 generation offspring reached 84 days of age, they

were mated (avoiding sibling pairings) and the repr oductive capacity was assessed by examining the females before and after mating to de termine time of pregnancy, estrous cyclicity, pre-coital time, pregnancy rates, and du ration of gestation. F 2 generation offspring were examined for abnormalities at parturition and then periodically until study termination on postnatal day 21. No adverse, treatment-related eff ects 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 NOAEL for maternal, reproductive, and developmental toxicity was 20 mg/kg-bw/day, based on no effects observed at the highest dose tested.



### A-1-5 Developmental Toxicity
In a pilot range-finding study of prenatal developm ental 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 thre e 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 stud y described above, pregnant SpragueDawley rats (25/dose) were administered 0, 50, 150, and 500 mg/kg-bw/day HHCB (>95 percent purity) by gavage on GDs 7 to 17 (Christian et al., 1997; Christian et al., 1999; as cited in EC, 2008). The dams were observed for clinical signs of toxicity. Food and water consumption and body weights were recorded at various intervals. On GD 20, the dams were sacrificed and a gross necropsy was performed. The numbers of corpor a lutea, implantation sites, resorption sites, live and dead fetuses, and live pup body wei ghts were recorded. Fetuses were examined for sex and underwent a gross examination. One half of the fetuses was examined for visceral
anomalies and the other half was examined for skele tal abnormalities. Signs of maternal toxicity in the dams consisted of significant incre ases in salivation, urine-stained fur, red or brown substance on the forepaws, and alopecia begin ning at 150 mg/kg-bw/day, with statistical significance at 500 mg/kg-bw/day. At ≥1 50 mg/kg-bw/day, statistically significant reductions in food consumption and maternal body ga in were observed, with the most severe reductions in body weight occurring during GDs 7 to 10. After completion of the dosing period (GDs 18 to 20), a significant rebound in food consu mption and body weight gains occurred, with values reported as being comparable to controls. Si gns of developmental toxicity were confined to the highest dose of 500 mg/kg-bw/day. Significan t reductions in offspring body weight were reported at the high dose; however, according to th e study authors, when severity, dose relationships, and historical ranges were taken int o consideration, these body weight reductions were not definitively considered to be t reatment-related. The number of litters or fetuses with morphological alterations (malformatio ns and variations) reported at the high dose was not statistically significantly increased when compared to the control group.
Malformations that did occur were few in number and were considered spontaneous in origin, except for the vertebral/rib malformations reported in only three fetuses that were not littermates. No other parameters were reported to be affe cted by treatment at any dose level. The

LOAEL for maternal toxicity was 150 mg/kg-bw/day, b ased on increases in clinical signs of toxicity and decreases in body weight, and the NOAE L was 50 mg/kg-bw/day. Although there is uncertainty concerning the effects reported in the offspring, a conservative LOAEL for developmental toxicity of 500 mg/kg-bw/day was conc luded and the NOAEL was 150 mg/kgbw/day.



### A-1-6 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 aberrati ons 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 Sl y, 1994; Kevekordes et al., 1998; as cited in EC, 2008).



### A-1-7 Carcinogenicity
A 2-year cancer study is not available for HHCB. As an alternative, EPA/OPPT evaluated the weight of the evidence from mutagenicity, subchroni c, and endocrine studies, to support the
conclusion that cancer data are not a critical data need.



### A-1-8 Additional Information

Irritation and sensitization studies in humans and animals showed that HHCB was not corrosive, irritating, or sensitizing in rabbits and humans. N o data on respiratory tract irritation were available. HHCB was considered to be a minimal eye irritant in rabbits and a possible photoirritant in rabbits and guinea pigs, but not in hum ans (Guillaume et al., 1973a; 1973b; Rubenkoenig and Ede, 1964; Lindstrum et al., 1978a; 1978b; 1978c; Harrison and Stolman, 1986; Gabriel and Mark, 1987; Folk and Dammers, 198 7; Shanahan and Alworth, 1987; Haynes, 1984; 1985; 1986; Sauer, 1980; Levenstein, 1973a; 1 973b; 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 designe d to test for potential endocrine disruption ( **i.e.** , estrogenic and/or anti-estrogenic activity), HHCB showed weak estrogenic and anti-estrogenic activity **in vitro** , dependent on the estrogen receptor type (Seinen et al., 1999; Bitsch et al., 2002; Schreurs et al., 2002; Schreurs et al., 2004; Schreurs et al., 2005a; Schreurs et al., 2005b; Gomez et al., 2 005; as cited in EC, 2008). Marginal repressing
effects were also found **in vitro** on the androgen and progesterone receptor. However , no estrogenic effects were seen in the **in vivo** uterotrophic assay.

Possible anti-estrogenic effects in zebrafish were assessed (Schreurs et al., 2004; as cited in EC, 2008). Transgenic zebrafish were exposed to 0.01, 0 .1, 1, and 10 μM HHCB with and without estradiol (E 2 ). The highest concentration of 10 μM HHCB was tox ic. Concentrations of 0.01, 0.1, and 1 μM HHCB were not estrogenic. Dose-dependent a ntagonistic effects were reported at 0.1 and 1 μM levels.Table_ApxA-1.Summary of Human Health Hazard InformationEndpointsHHCBCASRN (1222-05-5)References(as cited in EC, 2008)Acute oral toxicityLD50(mg/kg-bw)>3,000Minner and Foster (1977);Moreno (1975)Acute dermal toxicityLD50(mg/kg-bw)>3,250Minner and Foster (1977);Moreno (1975)Acute inhalation toxicityLC50(mg/L)No dataRepeated-dose toxicityOral (mg/kg-bw/day)Dermal (mg/kg-bw/day)Inhalation (mg/L/day)NOAEL = 150 (highest dose tested)No adequate dataNo adequate dataApi and Ford(1999);Hopkins**et al.**(1996)Reproductive toxicityOral (mg/kg-bw/day)Maternal NOAEL/LOAELReproductive NOAEL/LOAELDevelopmental NOAEL/LOAELNOAEL = 20 (highest dose tested)NOAEL = 20 (highest dose tested)NOAEL = 20 (highest dose tested)Ford and Bottomley (1997);Joneset al.(1996)Developmental toxicityOral (mg/kg-bw/day), ratsMaternal NOAEL/LOAELDevelopmental NOAEL/LOAELNOAEL = 50/LOAEL = 150NOAEL = 150/LOAEL = 500Christianet al.(1997);Christianet al.(1999)Genetic toxicity, gene mutation**In vitro**Genetic toxicity, gene mutation**In vivo**NegativeNegativeSan et al. (1994);Api and San (1999);Mersch-Sundermann et al. (1998a); Mersch-Sundermann etal. (1998b)Genetic toxicity, chromosomal aberrations**In vitro**Genetic toxicity, chromosomal aberrations**In vivo**NegativeNegativeApi and San (1999);Kevekordeset al. (1997);Curryand Putman (1995); Gudi and Ritter (1997); San andSly (1994); Kevekordes et al. (1998)Table_ApxA-1.Summary of Human Health Hazard InformationEndpointsHHCBCASRN (1222-05-5)References(as cited in EC, 2008)Additional informationCorrosivity/skin-eye Irritation/sensitizationRespiratory tract irritationCarcinogenicityEndocrine disruption**In vitro In vivo** NegativeNo dataNo dataWeakNegativeSee summary above

### A-2 Human Biomonitoring

Biomonitoring data describing human exposure levels of HHCB are summarized in Table A-2. Data are preliminary and limited based on small num bers of subjects, limited geographic representation, high variability, and incomplete de tails on methodology.

Several small studies have reported concentrations of HHCB in limited numbers of the general population in Asia and Europe. Only two small studi es 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 <10 0 samples and several of them did not provide data tables or details on the results of th e data. However, HHCB was detected in a high majority of the samples collected in the studies re ported here. Therefore, it can be assumed that exposure is widespread.

Both US studies took place in 2004 in the northeast ern part of the US. A study on adipose tissue in 49 residents of New York City undergoing liposuc tion reported a mean concentration of 178 ng/g lipid weight (lw) (range: 12 to 798 ng/g lw) (Kannan et al., 2005), which falls between the reported concentrations of HHCB in adipose tissue f rom two other available studies (Moon et al., 2012; Schiavone et al., 2010). A study of 43 K orean women reported a mean of 81 ng/g lw in adipose tissue (range: 28 to 211 ng/g lw) (Moon et al., 2012), while a small study in Italy reported a mean of 361 ng/g lw (range: 28 to 211 ng /g lw) (Schiavone et al., 2010).

Several small studies on concentrations of HHCB in breast milk have been conducted. In the US, a mean of 227 ng/g lw was reported for 39 samples c ollected 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 orde r to compare the results, the timing of the collection of the samples must be considered (lengt h 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 concent rations of HHCB in the US. A collection of matched maternal and cord blood samples in the Neth erlands indicated high detection and
correlation between maternal and cord blood serum. Concentrations reported in studies in the Netherlands, Austria, and China are so variable tha t they cannot and should not be compared (see Table A-2). Little data were provided in these studies, which were primarily done in 2005. Given that HHCB is sequestered in the fat, blood ma y not be an appropriate matrix for biomonitoring.

High variance in measurements may be a result of th e analytical methods for measuring HHCB in human matrices as they are still being developed and modified because the sample sizes are very small, the results should not be extrapolated to larger populations. Also, data from other countries may not represent those of the US since e xposure patterns may vary greatly between countries, especially for personal care products su ch as those that contain HHCB.Table_Apx A

- 2 . Human Biomonitoring Data for HHCB

Population Sampling Year(s) Levels Reference New York City residents undergoing liposuction; n = 49 (12 males, 37 females) 2003-2004

HHCB detected in adipose tissue in all samples Mean: 97 ± 88 ng/g ww

Range: 6.1-435 ng/g ww Mean: 178 ± 166 ng/g lw Range: 12-798 ng/g lw Kannan et al. (2005)

n = 43 women undergoing laparoscopy surgery for myoma 2007-2008

100 percent detection in adipose

tissue Mean: 81 ± 44 ng/g lw

Range: 28-211 ng/g lw Moon et al. (2012)

n = 12 surgical samples, Siena, Italy, 3 females, 9 males 2005-2006

HHCB detected in 92 percent of adipose tissue samples Mean: 361 ± 467 ng/g lw Schiavone et al. (2010)

n = 39 milk samples
from Massachusetts

women 2004

HHCB detected in 97 percent of breast milk samples; sample collection time not reported Mean ± SD: 227 ± 228

ng/g lw Range: <5-917 ng/g lw Reiner et al. (2007b)

10 primiparous mothers (25-29 years old ) 1999

2003/2004
HHCB detected in all breast milk samples;
collected 14-26
weeks after birth Median: 147 μg/kg fat

Range: 38-422 μg/kg fat Mean: 179 ± 111 μg/kg Large variability in individual samples Duedahl-Olesen et al. (2005)

n = 101 random samples of 266 breast milk samples collected from 44 primiparous women, Uppsala County 1996-2003, 14-21

days postpartum
HHCB detected in all breast milk samples Median: 63.9 ng/g lw

Range: 2.8-268 ng/g lw Lignell et al. (2008)

40 mothers (24-38 years old), Munich 1997-1998; HHCB

detected in 35/40 breast milk samples Median: 64 ng/g lw

Mean: 115 ng/g lw Range: 21-1,316 ng/g lw Liebl et al. (2000)Table_Apx A

- 2 . Human Biomonitoring Data for HHCB

Population Sampling Year(s) Levels Reference
Seoul, Korea n = 20 volunteers, >25 years old, 17 breast milk samples,
14 umbilical cord samples; 18 maternal serum samples 2007; breast milk

samples collected 3-10 days after delivery; HHCB
detected in 100 percent breast milk samples; 70 percent cord blood samples; 90 percent maternal serum samples Breast milk mean:

0.055-0.515 ng/g lw
Umbilical cord blood: 0.67-2.7 ng/g lw
Maternal serum: 0.17-1.4 ng/g lw Kang et al. (2010)

n = 110 mothers in
southwest China taking children for vaccines; 110 breast milk samples 2009 Median: 11.5 ng/g lw Range: <1.1-456.7 ng/g
lw Yin et al. (2012)

n = 10 breast milk samples from Chengdu, China; validating analytical
method 2009

HHCB detected in breast milk, all samples Range: 11.7-67.6 ng/g lw Wang et al. (2011)

100 volunteers in 3 cities in Yangtze River Delta, Shanghai, Shaoxing, and Wuxi 2006-2007

Breast milk samples
collected 1-2 weeks after delivery; 99 percent detection Median: 63 ng/g

Ranges: Shanghai: <5-278 Shaoxing: 5-782 Wuxi: 24-281 ng/g lw Zhang et al. (2011)

Serum samples from
volunteers in the
Netherlands
n = 42 maternal serum samples, 27 cord blood samples 2005 report

HHCB detected in 38/42 maternal serum samples and 26/27 cord blood samples Maternal: 0.15-3.2 ng/g

serum Cord blood: 0.11- 1.6 ng/g

serum Peters (2005)

Students at Medical University of Vienna n = 100 plasma samples (55 female, 45 male); 19-43 years old 2005 report

Detection in 91 percent of plasma samples; women had higher
levels than men but
lower values in 26-43 year olds Plasma concentrations:

Males, median: 260 ng/L Range: 98-540 ng/L Females, median: 580 ng/L Range: 290-885 ng/L Hutter et al. (2005)Table_Apx A

- 2 . Human Biomonitoring Data for HHCB

Population Sampling Year(s) Levels Reference Dept of Angiology at
Hanusch- Krankenhaus, Vienna;

n = 53 women >50 years old (51-71 years) Detection in

89 percent of plasma samples;
older women had higher
concentrations of

HHCB Maximum plasma

concentration: 6,900 ng/L No other data provided Hutter et al. (2010)

11 cities in China n = 204 (94 female, 110 male) 17-75 years old (median 25 years) 98 percent detection

in blood samples (almost all above LOQ) Whole blood

Median: 0.85 μg/L Maximum: 1.63 μg/L No data tables provided Hu et al. (2010)



### A-3 Summary of 2008 EU Human Health Risk Assessment

The human health hazards and risks of HHCB have bee n extensively assessed by several other reliable entities (EC, 2008; OECD, 2009; EPA, 2003; HERA, 2004a; 2004b). Each of these welldocumented, independent reviews concluded that ther e was no significant risk to human health from exposure to HHCB as used in household c leaning products. Moreover, the overall conclusions of the EU RAR (EC, 2008) were that ther e was no need for further information and/or testing and no need for risk reduction measu res beyond those already being applied. The EPA/OPPT has thoroughly reviewed these other as sessments and presents a summary of
the most recent EU RAR.



### A-3-1 Assumptions and Points of Departures Used in the EU

RAR
The EU RAR (EC, 2008) assessed risk in three popula tions for relevant endpoints: **workers** (exposures from handling during production and dilu tion, compounding, formulation, and professional cleaning); **consumers** (exposures from a wide variety of consumer product s such as perfumes, creams, toiletries, soaps, shampoos, and household and laundry cleaning products);
and **humans exposed via the environment** (exposures **via** food and water including fish, root crops, and mother’s milk).

Dermal absorption of 16 and 5.2 percent in rats and humans, respectively, was used. Since toxicokinetic data of HHCB following oral and inhal ation absorption are not available in humans or animals, intermediate defaults of 50 percent for oral absorption and 100 percent absorption from inhalation exposures were used.For general systemic effects following repeated exp osures, the oral NOAEL of 150 mg/kgbw/day from the 13-week subchronic toxicity study i n rats was used as a point of departure. Since NOAELs are not available following dermal or inhalation exposures, the administered dose from the oral 13-week toxicity study was conve rted into an internal dose (body burden) by taking into account different absorption factors. B ased 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 expos ures 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) toxi city study was used as a point of departure (Ford and Bottomley, 1997; Jones et al., 1996; as c ited 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 (simu ltaneous) dermal and inhalation exposures for compounding workers only (the only sc enario relevant for combined exposures).



### A-3-2 Risk to Workers
Oral exposures in workers were assumed to be mitiga ted by personal hygiene measures, and therefore, only risks following dermal and inhalati on 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 studie s by the inhalation route are not available, low-level, short-term exposures in workers, combine d with the low acute toxicity by the oral route, suggested no concern for workers for acute t oxicity by the inhalation route.

Given that HHCB showed no potential for skin/eye ir ritation, 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 potenti al was expected.

For both the dermal and inhalation exposure scenari os for the general systemic toxicity endpoint, a minimal margin of safety (MOS) of 100 w as used for comparison. The MOS was based on an interspecies factor of 10 (4 for metabo lic size differences and 2.5 for remaining differences), an intraspecies factor of 5, and a fa ctor of 2 for extrapolation from semi-chronic to chronic exposure. For each inhalation/dermal exposu re scenario, the calculated MOS was well above the minimal MOS of 100 (≥2,600). Combined exp osure routes for total body burdens for skin contact and inhalation for the compounding sce nario resulted in MOS values well above 100 (≥2,000). As a result, it was concluded that fo r 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 thos e 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 expos ure scenarios, there is, at present, no need for further information and/or testing and no need for risk reduction measures beyond those already being applied. Although there are no carcin ogenicity data for HHCB, it was concluded
that there was no concern for workers under these e xposure scenarios for carcinogenicity based on negative mutagenicity data and the lack of any apparent structural alerts that would

raise a concern.

For both the dermal and inhalation exposure scenari os 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 dif ferences and 2.5 for remaining differences) and an intraspecies factor of 5. For dermal, inhalation , and combined exposure scenarios, the calculated MOS was above 50 (≥261). As a result, it was concluded that for workers under these exposure scenarios for developmental/reproductive t oxicity, 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-3-3 Risk to Consumers
The main route of exposure for consumers was assume d to be dermal, with some inhalation exposures and no oral exposures.

Given that HHCB showed no potential for skin/eye ir ritation, corrosivity, or sensitization/ photosensitization, it was concluded that there was no concern for local effects in consumers following repeated exposures. Since there is a lack of skin and eye irritation potential, no significant respiratory tract irritation potential was expected.

For general systemic effects following repeated exp osures, a MOS of 200 (based on a factor of 10 for intraspecies differences, 4*2.5 for interspe cies differences, 2 for duration extrapolation, and 1 for dose-response) was used. The calculated M OS for dermal exposure was well above 200 (≥1,400). Therefore, it was concluded that ther e was no concern for general systemic toxicity in consumers following repeated dermal exp osures.

Given that HHCB was not found to be mutagenic in a wide array of studies, it was concluded that for mutagenicity for consumers, there is, at p resent, no need for further information and/or testing and no need for risk reduction measu res beyond those already being applied. Likewise, although there are no carcinogenicity dat a for HHCB, it was concluded that there was no concern for consumers under these exposure scena rios for carcinogenicity based on negative mutagenicity data and the lack of any appa rent structural alerts that would raise a

concern.

For the dermal exposure scenario for the developmen tal/reproductive toxicity endpoint, a minimal MOS of 100 was used for comparison. The MOS was based on an intraspecies factor of10, an interspecies species factor of 10 (4 for met abolic size differences and 2.5 for remaining

differences), and factor of 1 for dose-response. Th e calculated MOS was above 100 (MOS ≥189).
It was concluded that for consumers with a dermal e xposure scenario for developmental/ reproductive toxicity, there is, at present, no nee d for further information and/or testing and no need for risk reduction measures beyond those alrea dy being applied.



### A-3-4 Risk to Humans Exposed Indirectly via the Environme nt Exposure to HHCB by the inhalation route ( **via** air) was considered to be negligible. The main route of exposure to humans in the environment was oral ( **via** fish and root crops). Exposures were based on local and regional daily intake estim ates following production. A separate risk
characterization for breast-fed babies was conducte d for exposure **via** mother’s milk (see next section for more information).

For general systemic effects following repeated exp osures, an internal NOAEL of ≥75 mg/kgbw/day and a MOS of 200 (based on a factor of 10 fo r intraspecies differences, 4*2.5 for interspecies differences, 2 for duration extrapolat ion, and 1 for dose-response) were used. The calculated MOS for exposure **via** food was >3E+4 for the local production scenario a nd >8E+5 for the regional production scenario. Therefo re, it was concluded that for the local and regional exposure scenarios, there was no concern f or general systemic toxicity in humans exposed indirectly **via** the environment.

As with the other populations, since HHCB was not f ound to be mutagenic in a wide array of studies, it was concluded that for mutagenicity, th ere is, at present, no need for further information and/or testing and no need for risk red uction measures beyond those already being applied. Likewise, although there are no carc inogenicity data for HHCB, it was concluded that there was no concern for carcinogenicity based on negative mutagenicity data and the lack of any apparent structural alerts that would raise a concern.

For the oral exposure scenario for the developmenta l/reproductive toxicity endpoint, an internal NOAEL of ≥10 mg/kg-bw/day and a MOS of 100 were use d. 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 val ues for local and regional oral exposure scenarios were 3,846 and 10,000, respectively, indi cating no concern for developmental/ reproductive toxicity in humans exposed indirectly **via** the environment.



### A-3-5 Assessment of Risk for Breast-Fed Babies Exposed vi a

Mother’s Milk
The presence of HHCB in human milk was considered t o be the result of indirect environmental exposures from a variety of sources, including mate rnal exposure to consumer products; intake

**via** food, water, or air; and occupational exposures.The concentrations of HHCB in milk samples for huma ns and rats were compared. By using the

dose level in dams (20 mg/kg-bw/day, the highest do se tested from the oral peri/postnatal developmental toxicity study), the measurement of l evels of HHCB in the milk of dams (17.6 and 5.0 μg/mL at four or eight hours post dos ing, 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 ex posed to levels approximately 100 to
360 times the maximum level found in human milk sam ples.

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 avera ge daily milk consumption for the rat pup, were estimated. A comparison of these two estimates of uptake showed a difference of approximately three orders of magnitude between the levels of HHCB exposure in the rat (in which no adverse effects were found) and the human infant exposure.

Total internal (worst-case) combined exposures were estimated and compared to the internal NOAELs for general systemic toxicity following repe ated exposures and for developmental/ reproductive toxicity (≥75 and ≥10 mg/kg-bw/day, re spectively) in order to calculate MOS values. The worst-case combined exposure was estima ted from the sum of the worst-case estimates from three populations: (1) dermal and in halation exposures for scenario 2 for compounding workers; (2) dermal and inhalation expo sures for consumers; and (3) oral and inhalation, locally **via** the environment. For general systemic effects fol lowing repeated exposures, a MOS of 100 was used (based on 4*2.5 fo r metabolic size and other differences, a factor of 5 for intraspecies differences, and a fac tor of 2 for semichronic to chronic exposure extrapolation). The calculated MOS was well above 1 00 (798). For the developmental/ reproductive toxicity endpoint, a MOS of 50 (based on 4*2.5 for interspecies species differences and 5 for intraspecies differences) was used. The c alculated 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 informatio n and/or testing or risk reduction measures beyond those already being appli ed.



### A-4 Key Sources of Uncertainty and Data Limitations on Human Health

Overall, adequate screening-level animal toxicity d ata are available to characterize the human health hazard for HHCB. Toxicokinetics (by the derm al route), acute, repeated-dose, and developmental toxicity data are available to charac terize the human health hazard of HHCB. Although no multigenerational reproductive toxicity studies on HHCB are available, information on developmental and reproductive toxicity was obta ined from both the repeated-dose dietary toxicity study (reproductive organ data) and the pe ri/post-natal reproductive toxicity study with modified exposures. Several assays testing for end ocrine disruption, genotoxicity, and irritation/sensitization (including several studies in humans) are also available. The database isincomplete for toxicokinetic data in animals by the oral and inhalation routes; acute toxicity

data by the inhalation route; repeated-dose toxicit y data by the inhalation and dermal routes; and chronic toxicity/carcinogenicity. Data in human s on most toxicity endpoints are not
available.

One area of uncertainty concerns the effects report ed in offspring in the prenatal developmental toxicity study in rats (Christian et al., 1997; Christian et al., 1999; as cited in EC, 2008). When severity, dose relationships, and histo rical ranges were taken into consideration, the reductions in pup body weight were not definiti vely considered by the study authors to be treatment-related. Additionally, fetal malformation s observed were reported in only three fetuses from separate litters. Even though there is uncertainty surrounding these endpoints, the study authors, as well as other reliable entiti es, have concluded a conservative LOAEL for developmental toxicity of 500 mg/kg-bw/day from thi s 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 ne onate when exposed through nursing. Exposures in the F 1 offspring occurred **in utero** from GD 14 through lactation; there were no exposures 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 endp oints such as pup weight, pup survival, and postnatal death) and reproductive toxicity (reprodu ctive performance) as well as a battery of developmental neurotoxicity tests. Other reliable entities such as the 2008 EU RAR on HHCB (EC, 2008) and the 2009 SIDS Initial Assessment Pro file (OECD, 2009) also concluded that this study was adequate for assessing hazard for these e ndpoints.

Human biomonitoring data indicate that HHCB is pres ent in milk, fat, and blood, but the studies in the US are preliminary and of limited value for characterizing exposures. Specifically for the US, there are few biomonitoring studies in limited locations with small numbers of samples. HHCB has not been measured as part of the National Health and Nutrition Environmental Survey (NHANES), so the data are not representative . Exposures in the US are assumed to be similar to those in Europe, but it is unknown how m anufacture and use of products containing HHCB may differ between regions or cultures since u se of fragrances, personal care, and cleaning conventions may vary.



### A-5 Conclusions of Human Health Assessment

The available toxicokinetic data indicate that HHCB is poorly absorbed through the skin;
toxicokinetic data for the oral and inhalation rout es 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 preliminaryand 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 activit y has been demonstrated. HHCB is not mutagenic, corrosive, irritating, or sensitizing; h owever, minimal eye irritation and possible signs of photoirritation have been reported. Carcin ogenicity data are not available.

HHCB was initially selected for review based on a m oderate hazard concern for developmental toxicity and a high potential for exposure. However , following further review of the developmental and reproductive toxicity data, and t aking 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 fetuse s from separate litters exhibiting fetal malformations at 500 mg/kg-bw/day from this same st udy, and the lack of fetal morphological changes observed in the pilot studies, even at mate rnally lethal dosages (up to 1,000 mg/kgbw/day) lead to the conclusion that 500 mg/kg-bw/da y is a conservative LOAEL. No effects were reported in fetuses in the perinatal toxicity study with lactational exposures at doses similar to those reported in human milk and several orders of magnitude higher (20 mg/kgbw/day). Overall, a low concern for developmental t oxicity 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 r ecently the EU, and it has been concluded that the overall concern for human health hazards, inclu ding that for developmental toxicity, is low.

The review of human health hazard studies, biomonit oring studies in the US and elsewhere,
and the EU RAR showed an overall low risk concern f or human health, including the risk for developmental toxicity, from current use of HHCB.

## 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 Apx B-1. Chemical Synthesis of HHCB

HHCB is manufactured by a three-step reaction (Ullm ann, 2003; Zviely, 2002). First, a cycloaddition reaction of .alpha.-methyl styrene an d 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 hydroxyalkylated with prop ylene oxide in a Friedel-Crafts reaction using aluminum chloride as a catalyst. Third, the r ing closure of the resulting 1,1,2,3,3- pentamethyl-5-(-hydroxyisopropyl)indane (2) to 1,3, 4,6,7,8-hexahydro-4,6,6,7,8,8- hexamethylcyclopenta- γ

-benzopyran (HHCB; Galaxolide) is accomplished with paraformaldehyde and a lower aliphatic alcohol **via** the acetal or with paraformaldehyde and a carboxylic acid anhydride **via** the acylate.

1) Ullmann's Encyclopedia of Industrial Chemistry (200 3). 6th ed. Vol 1: Federal Republic of Germany: Wiley-VCH Verlag GmbH & Co. 2003 to Presen t, p. V14 145. 2) Zviely M; Kirk-Othmer Encyclopedia of Chemical Tech nology. (2005). NY, NY: John Wiley & Sons; Aroma Chemicals. Online Posting Date: Jan 2 5, 2002.

## Appendix C HHCB (*), HHCB DIASTEREOISOMERS (#1 TO #6), AND RELATED STRUCTURAL ANALOGS (#7 TO #15)

Table_Apx C

- 1

. HHCB, HHCB Diastereoisomers
, and Related Structural Analogs

CASRN Chemicals Structure Chemical Index Name On TSCA Inventory

* 1222

- 05

- 5

C H 3 CH 3

CH 3 CH 3 C H 3

O CH 3 1,3,4,6,7,8

- Hexahydro

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

- hexamethylcyclopenta- γ

-2-benzopyran Y

1 172339

- 62

- 7 Cyclopenta[g]

- 2

-benzopyran,1,3,4,6,7,8-hexahydro4,6,6,7,8,8-hexamethyl-, (4S,7S)- N

2 172339

- 63

- 8 Cyclopenta[g]

- 2

-benzopyran,1,3,4,6,7,8-hexahydro4,6,6,7,8,8-hexamethyl-, (4R,7S)- N

3 252332

- 95

- 9 Cyclopenta[g]

- 2

-benzopyran,1,3,4,6,7,8-hexahydro4,6,6,7,8,8-hexamethyl-, (4S,7R)- N

4 252332

- 96

- 0 Cyclopenta[g]

- 2

-benzopyran,1,3,4,6,7,8-hexahydro4,6,6,7,8,8-hexamethyl-, (4R,7R)- NTable_Apx C

- 1

. HHCB, HHCB Diastereoisomers
, and Related Structural Analogs

CASRN Chemicals Structure Chemical Index Name On TSCA Inventory

5 252933

- 48

- 5 Cyclopenta[g]

- 2

-benzopyran,1,3,4,6,7,8-hexahydro4,6,6,7,8,8-hexamethyl-, (4R,7R)-relN

6 252933

- 49

- 6 Cyclopenta[g]

- 2

-benzopyran,1,3,4,6,7,8-hexahydro4,6,6,7,8,8-hexamethyl-, (4R,7S) relN

7 1222

- 06

- 6 Cyclopenta[g]

- 2

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

8 857091

- 61

- 3 Cyclopenta[g]

- 2

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

9 102296

- 64

- 0 Cyclopenta[g]

- 1

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

10 135546

- 43

- 9 Cyclopenta[g]

- 1

-benzopyran,2,3,4,6,7,8hexahydro4,6,6,7,8,8-hexamethylN

11 135546

- 42

- 8 Cyclopenta[g]

- 2

- benzopyran, 1,3,4,6,7,8-hexahydro-1,6,6,7,8,8-hexamethylNTable_Apx C

- 1

. HHCB, HHCB Diastereoisomers
, and Related Structural Analogs

CASRN Chemicals Structure Chemical Index Name On TSCA Inventory

12 114109

- 63

- 6 Cyclopenta[h]

- 2

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

13 114109

- 62

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

14 78448

- 48

- 3 Cyclopenta[g]

- 2

- benzopyran, 6

- ethyl

1,3,4,6,7,8-hexahydro-4,6,8,8-tetramethylN

15 78448

- 49

- 4 Cyclopenta[g]

- 2

- benzopyran, 8

- ethyl

1,3,4,6,7,8-hexahydro-4,6,6,8-tetramethylN

*This compound was evaluated in this assessment.

## Appendix D CDR DATA FOR HHCB

The information in Tables D-1 through D-4 is from E PA’s non-CBI CDR database (EPA, 2014b) for the 2012 reporting cycle, for HHCB, CASRN 1222-05-5 . The chemical name used in the CDR database is cyclopenta[g]-2-benzopyran, 1,3,4,6,7,8 -hexahydro-4,6,6,7,8,8-hexamethyl-.

Table_Apx D

- 1

. CDR National HHCB Information a

Production volume (aggregated) 3,126,728 lbs

Maximum concentration (at manufacture or import sit e) > 90 percent

Physical form(s): Liquid ; Dry Powder, Liquid, Other

Number of manufacturing, processing, and use sites Not completely reported

. b

Number of reasonably likely to be exposed industria l manufacturing, processing and use workers Not completely reported

. c

Was industrial processing or use information report ed? Yes

Was commercial or consumer use information reported ? Yes

a More detailed information was not publically avail able on the CDR website because it was considered t o be CBI.

b The total number of sites reported by two submitter s is in the range of less than 65 to 123 but one su bmitter reported that the number of sites is “not known or reasonably ascertainable” and the other submitters did not report any information.

c The number of workers reported by one submitter is in the range of 25 – 49, but the other submitters e ither reported that the number of sites is “not known or reasonably ascertainable” or else did not report an y information.

Table_Apx D

- 2

. CDR HHCB Industrial Use Information

Type of Processing Industrial Sector (Based on NAIC S) Industrial Function

Processing

— incorporation into formulation, mixture, or reaction product Soap, Cleaning Compound, and Toilet Preparation Manufacturing Odor agents

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 Other (requires additional information) Odor agents

Processing—incorporation into article Plastics Material and Resin Manufacturing Odor agentsTable_Apx D

- 3

. HHCB CDR 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

Cleaning and Furnishing Care
Products Reported as “Not Known or reasonably ascertainable” and
“<1% - <30%” Reported as “Not known or reasonably ascertainable” and “No” by different submitters

Personal Care Products Reported as “Not Known or reasonably ascertainable” and
“<1% - <30%” “Not known or reasonably
ascertainable”

Air Care Products Reported as “Not Known or reasonably ascertainable” and
“<1% - <30%” Reported as “Not known or reasonably ascertainable” and “No” by different submitters

Laundry and Dishwashing
Products Reported as “Not Known or reasonably ascertainable” and
“<1% - <30%” Reported as “Not known or reasonably ascertainable” and “No” by different submitters

Plastic and Rubber Products
not elsewhere Covered <1% Not known or reasonably
ascertainable

Non-TSCA Use 1% - <30% Not known or reasonabl y ascertainable

Source: EPA (2014b)

Table_Apx D

- 4

. CDR Company Site Information (2012)

Company Site City State Zip
Code Manufacture Import Site
Limited

Berje, Inc. Berje, Inc. Carteret NJ 07003 No Yes N/A

International Flavors & Fragrances, Inc. Ashland, Inc. Hazlet PA 19067

- 3701 No CBI N/A

Firmenich, Inc. Firmenich, Inc. Plainsboro NJ 08543 CBI CBI

-- Symrise, Inc. Symrise, Inc. Branchburg NJ 08773 No Yes N/A

S C Johnson & Son,

Inc. S C Johnson & Son,

Inc. Sturtevant WI 53177 CBI CBI

--

Source: EPA (2014b)

## Appendix E MODELED RELEASE ESTIMATES ACCORDING TO
STAGE OF PRODUCTION AND USE

This Appendix contains modeling estimates of releas es of HHCB to the environment from industrial sites of all stages of the HHCB life cyc le and estimates of releases from consumer and commercial use of end-use products.

In general, the life cycle of a fragrance ingredien t such as HHCB includes:

• Manufacturing of the chemical substance,

• Compounding of fragrance oils containing the chemic al 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.

The release estimates are summarized in Table E-1 a nd 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. These release esti mates are based on the estimated 2011 use volume of 1,700 metric tons (3.74 million lbs) for HHCB (IFRA, 2012b). This volume represents the vast majority of the quantity of HHCB used to p repare fragrance oils or fragrance compounds within the United States, based on data c ollected from IFRA’s Volume of Use Survey (IFRA, 2012d).

Because total releases as a fraction of production volume is independent of the production volume or the number of sites, the estimation of re lease amounts was not revised to account for the CDR data subsequently reported in 2012 (3.1 million lbs). A discussion of the release estimates according to stages of the life cycle fol lows.



### E-1 Estimated Release from Manufacture and Import
As discussed in Chapter 2, HHCB is not currently ma nufactured in the US (IFRA, 2012a). For this release assessment, four companies are assumed to i mport HHCB to the US at five sites (see Appendix D, Table D-4). Releases are not expected t o result from import activities, but may occur at import sites if HHCB is also diluted and c ompounded onsite after import, as further
discussed below.Table_ApxE-1.Summary of Estimated Environmental ReleasesRelease Activity Release MediaRelease Factor(Amount Releasedper Amount Used,Percent)Numberof SitesNumber ofRelease Daysat Each Site(Days/Year)Daily Release(kg/Site-Day)Combined AnnualReleases of All Sites(kg/Year)EPA Models andAssumptionsImport and Compounding of Fragrance OilsUnloading 100 percent of HHCBat 75 °C from bulk containersinto dilution tankAir NA 5-49 2-18a9.02 × 10-30.812-0.884 EPA AP-42 Model (EPA,1991; Fehrenbacher andHummel, 1996)Transferring 65 percent of(diluted) HHCB at 75 °C fromdilution tank into totes fordistribution to customers orindoor vessels for compoundingAir NA 5-49 26-251a3.22 × 10-4-6.45 × 10-40.405-0.821 EPA AP-42 Model (EPA,1991; Fehrenbacher andHummel, 1996)Cleaning of bulk containers Water 0.07-0.2b5-49 2-18a13.2-37.8c1,190-3,400dEPA Bulk TransportResidual Model (EPA,1988; 1992)Cleaning compoundingequipmentWater 0.07-1b5-49 250 0.097-13.6 1,190-17,000 EPA Single Process VesselResidual Model (EPA,1988; 1992)Table_ApxE-1.Summary of Estimated Environmental ReleasesRelease Activity Release MediaRelease Factor(Amount Releasedper Amount Used,Percent)Numberof SitesNumber ofRelease Daysat Each Site(Days/Year)Daily Release(kg/Site-Day)Combined AnnualReleases of All Sites(kg/Year)EPA Models andAssumptionsBlending of Fragrance Oils into ProductsCleaning of transport containersWater,incineration,land0.3-3e950 250 0.021-0.2155,100-51,000 EPA/Office of PollutionPrevention and Toxics(OPPT) Drum ResidualModel (EPA, 1988; 1992)Cleaning of process vessels Water,incineration,land2 950 250 0.143 34,000 EPA/OPPT Multiple VesselResidual Model (EPA,1988; 1992)Conveying, mixing, andpackaging powder productsAir, water,incineration,land3.83 950 250 0.274 65,110 4.5 percent dust lossesfrom spray-drying unitwith 85 percent airpollution control deviceefficiency (4.5 percent ×0.85 = 3.825 percent)(OECD, 2010)Use of End ProductsDisposal of products Water 90.0-93.7 Notknown250 Not known 1,530,000-1,592,900Engineering judgmentassuming a 100 percentrelease scenarioTotal ReleasesWater 90.1-94.9 NA 2-250 NA 1,532,380-1,613,300Air NA NA 2-251 NA 1.22-1.71Water,incineration,land (uncertain)2.3-5 NA 250 NA 39,100-85,000Air, water,incineration,land (uncertain)3.83 NA 250 NA 65,110Table_ApxE-1.Summary of Estimated Environmental ReleasesRelease Activity Release MediaRelease Factor(Amount Releasedper Amount Used,Percent)Numberof SitesNumber ofRelease Daysat Each Site(Days/Year)Daily Release(kg/Site-Day)Combined AnnualReleases of All Sites(kg/Year)EPA Models andAssumptionsNOTES: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.aThe number of unloading days, which is equal to thenumber of release days, is calculated by dividingthe average HHCB throughput per site by the containervolume, assuming that one container is unloaded persite per day. Bulk containers are assumed to be 5,000 gallons and totes are assumed to be 550 gallons.The calculation for the cleaning of bulk containersis presented as a sample calculation of this variable:Low-end of range of number of unloading days (or release days) per site:HHCB throughput per site = (1,700,000 kg/year ÷ 49sites) ÷ (1 g/cm3× kg/1,000 g × cm3/0.0002642 gallons) = 9,166 gallons per site.Number of release days per site = 9,166 gallons/site-year ÷ 5,000 gallons/site-day = 1.83 days ~2 days.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/cm3× kg/1,000 g × cm3/0.0002642 gallons) = 89,828 gallons per site.Number of release days per site = 89,828 gallons/site-year ÷ 5,000 gallons/site-day = 18 days.bThe release factor, or the amount released per unitamount used, is equal for this release activity tothe residual amount in a transport container or processequipment. This residual amount is expressed as afraction 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).cThe calculation of this range of releases per siteis 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 residualgallons/container gallon × (1 g/cm3× kg/1,000 g × cm3/0.0002642 gallons) = 13.2 kg/site-dayUpperHigh-end of range of amount released per site-day =5,000 gallons/container-site-day × 0.002 residual gallons/container gallon × (1 g/cm3× kg/1,000 g × cm3/0.0002642 gallons) = 37.8 kg/site-daydThe calculation of this range of combined releasesof all sites per year is presented as a sample calculation of this variable:Lower-end of range of amount released per year = 13.2 kg released /site-day × 5 sites × 18 days of release = 1,190 kg/yearUpper-end of range of amount released per year = 37.8 kg released/site-day × 5 sites × 18 days of release = 3,400 kg/yeareResidual amount as a fraction of drum volume: (1)containers unloaded by pumping: 3 percent (high-end) and 2.5 percent (central tendency); (2) containersunloaded by pouring: 0.6 percent (high-end) and 0.3 percent (central tendency).

### E-2 Estimated Release from Compounding

The number of compounding sites in which HHCB is pr ocessed 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. E PA/OPPT expects that imported HHCB may be delivered to compounding sites. Compounding site s are classified under North American Industry Classification System (NAICS) code 32562, Toilet Preparation Manufacturing, and specifically under the Perfume Oil Mixtures and Ble nds Manufacturing subcategory. This NAICS
code is not mentioned in the 2006 non-CBI IUR data on HHCB; therefore, EPA/OPPT assumed that HHCB compounding may occur only at the five im port sites. This approach resulted in a high estimate of per-site release. Alternatively, t he number of sites was estimated to be equal to the number of IFRA member companies. The estimat ed HHCB volume used to compound fragrance oils was collected through surveys of IFR A member companies, which represent the vast majority of fragrance volume produced in the U S. IFRA North America currently has 49 member companies (IFRA, 2012d). If each member com pany 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 NAI CS 32562 by subcategory, but shows that the total number of companies within NAICS 32562 in creased by four percent from 2002 to 2007. EPA/OPPT 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 subcat egory in 2007. 12 Therefore, the number of compounding sites is estimated to include 5 to 49 s ites. The geographic distributions of all sites
under NAICS 32562 are shown in Table E-2.

In summary, compounding could occur at these sites under two scenarios: (1) after import, HHCB is diluted and compounded at the five import s ites; or (2) HHCB is delivered to and compounded at the 49 compounding sites identified i n the Census. During compounding, EPA/OPPT assumes that imported HHCB is unloaded int o large outdoor storage tanks for dilution to a 65 percent solution and is subsequent ly transferred to an indoor tank in the compounding facility for further processing. Typica lly, HHCB is present in the compounded fragrance oils at two to four percent (HERA, 2004a)

.

At room temperature (25 °C), HHCB is a non-volatile liquid with a vapor pressure of

5.47 × 10 -4 mmHg (see Table 2-1). Because HHCB is a highly vis cous liquid at concentrations ≥65 percent, concentrated HHCB may be heated to bec ome less viscous during transfer activities, during which HHCB may have sufficient v apor pressure to volatilize (EC, 2008). The unloading and transfer are expected to occur at tem peratures between 25 and 75 °C (EC, 2008). Therefore, there is potential for releases to air d ue to volatilization from unloading and transfer

12 Because the number of establishments under this su bcategory is not known, and most companies under NAICS 32562 are small companies with no more than o ne establishment, the 34 companies are assumed to represent 34 establishments (sites).Table_Apx E

- 2

. Geographic Distribution for Facilities under NAICS 32562 Toilet Preparation Manufacturing

State Number of

Establishments Establishments with 20

Employees or More Number of Employees

California 188 59 7,572

New Jersey 97 56 10,227

New York 62 24 6,243

Texas 60 20 3,134

Florida 59 16 1,531

Illinois 41 22 3,860
Pennsylvania 30 10 1,575
Georgia 24 6 652

Connecticut 22 11 2,582

Minnesota 18 10 1,469

Missouri 17 6 1,091

North Carolina 17 9 3,984

Ohio 17 11 2,370

Colorado 14 3 283

Tennessee 14 9 2,033
Washington 14 1 106

Wisconsin 12 3 190

Arizona 10 6

— a

Massachusetts 9 4 249

Arkansas 8 4 1,545
Maryland 8 3 500

Idaho 7 3 318

Indiana 7 3 378
Michigan 7 5 2,795

Utah 7 3

— b

Virginia 7 4 1,208
Kentucky 5 2

— a

Iowa 4 2

— b

Delaware 2 1

— b

Vermont 2 1

— c

a 250 to 499 employees.

b 500 to 999 employees.

c 100 to 249 employees. Source: USCB (2007a).of HHCB at elevated temperatures. Releases to other environmental media may also occur from

cleaning of transport containers and compounding ve ssels. The environmental release media may include water or incineration, depending on the method of cleaning (IFRA, 2012c).

No quantitative release information was found for t he compounding scenarios. Compounding releases were assessed using US EPA release models related to cleaning and transfer activities (Fehrenbacher and Hummel, 1996; EPA, 1988; 1991; 19 92). In the absence of information to indicate otherwise, EPA/OPPT assumed that container and equipment residues containing HHCB were released to water. EPA/OPPT estimated the number of release days per year for container unloading and transfer activities at comp ounding sites, assuming that one container is unloaded per site per day and using a default un loading rate of approximately 200 gallons per minute. EPA/OPPT estimated the number of release da ys per year for equipment cleaning by assuming that cleaning occurs after each batch, the re is one batch per site per day, and there are 250 days per year operation at each site.



### E-3 Estimated Release from Blending of Fragrance Oils
After compounding, fragrance oils containing HHCB a re blended with other materials to formulate commercial and consumer products. EPA/OPP T assumed that HHCB may be present in the formulated products at a maximum of 0.9 perc ent (HERA, 2004b). The 2006 non-CBI IUR reported a range of 100 to 999 industrial processin g and use sites specifically associated with
HHCB. The estimate of 100 to 999 sites includes bot h compounding and blending sites. Because there are approximately 49 compounding sites, EPA/O PPT estimated that there may be up to 950 blending sites (999 sites – 49 sites = 950 site s). Blending sites are classified under NAICS codes 32562, 325611, and 325612. A geographic distr ibution of these sites, many of which are located in California, New York, and Texas, is prov ided in Tables E-3 and E-4. Note that the data include sites within the NAICS industry sectors, an d may include sites that do not specifically
handle HHCB.Table_Apx E

- 3

. Geographic Distribution for Facilities under NAICS 325611 So ap 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

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_Apx E

- 4

. Geographic Distribution for Facilities under NAICS 325612 Polish and Other Sanitization Goods 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
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 commerci al and consumer products, releases may result from the cleaning of transport containers an d process vessels. Additionally, releases mayoccur as a result of conveying, mixing, and packagi ng of powder products containing HHCB. No

quantitative release information was found for the scenario of blending of fragrance oils; therefore, these releases were estimated using the OECD Emission Scenario Document (ESD) on
the **Blending of Fragrance Oils into Commercial and Cons umer Products**  (OECD, 2010). The ESD assumes that fragrance oils may be transported to b lending sites in drums. Further, the ESD indicates a potential for dust losses if the operat ion involves the blending of solid commercial and consumer products ( **e.g.** , powder detergent) (OECD, 2010). Because the volum e of HHCB formulated into solid and powdered products is not known, EPA/OPPT conservatively assumes that the entire volume is formulated into powder pr oducts, which results in an overestimate of the release amount. EPA/OPPT 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 for cleaning, and actual releases may n ot occur onsite. EPA/OPPT also assumes that one batch of fragrance product is blended per site per day and that equipment is cleaned after
each batch.



### E-4 Estimated Release from Use of Commercial and Consum er Products
No quantitative release information was found for t he use of commercial and consumer products containing HHCB. EPA/OPPT assumed that th e use of all end-use products containing
HHCB results in down-the-drain release of HHCB to w ater. The basis for this assumption is the information on the type of end-use products that fr agrance oils are used in, as summarized in Table 2-7. All of the types of end-use products lis ted with the possible exception of “other” and “fine fragrances” are products whose use is likely to result in down-the-drain release of these products, and these products together represent 89% of the use volume of HHCB. It is noted that the data in Table 2-7 is neither specific to H HCB nor to the use of HHCB in the US. However, a 100% release to water from the use of en d-use products has also been assumed in other HHCB risk assessments (EC, 2008 and HERA, 200 4b). This assumption is deemed to be adequate for the purpose of this release assessment as well. These releases, combined with upstream losses from compounding and blending, acco unt for 100 percent of the HHCB use volume. Wastewater containing detergent and cleanin g 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 c ombined releases from all sites on an
annual basis.

## Appendix F ADDITIONAL STUDIES

F-1 Endocrine Mechanisms and Molecular Pathways

Studies were available on the effects of HHCB on en docrine mechanisms and other molecular pathways and are briefly summarized here.

The ability of HHCB to bind and interfere with ster oid hormone receptors was investigated in liver cells of several aquatic organisms and also i n cell lines. Weak estrogenic activity of HHCB (purity unknown) was observed in competitive estrog en receptor binding assays with rainbow trout, carp, and the amphibian, **Xenopus laevis** , and only at relatively high concentrations (Dietrich and Chou, 2001; as cited in EC, 2008; Sch reurs et al., 2002; Schreurs et al., 2004).
HHCB showed no **in vivo** estrogenic activation of vitellogenin production i n carp ( **Cyprinus carpio**  ) (Seinen **et al.** , 1999). HHCB did induce vitellogenin gene expressi on 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 anta gonistic 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 (Schreur s **et al.** , 2004). No **in vivo** estrogenic activity was detected in a transgenic zebrafish assay, but a nti-estrogenic effects were observed **in vivo** (Schreurs et al., 2004). HHCB (53.5 percent in DEP ) was shown to inhibit estrogen-induced vitellogenin production in rainbow trout within the range of concentrations that have been
detected in tissues of fish from contaminated locat ions (Simmons et al., 2010). A study of the interactions of HHCB with fish metabolic systems in carp showed that HHCB significantly inhibited Phase I and Phase II enzymes involved in the synthesis and metabolism of steroids (Schnell et al

**.** , 2009).

The ecdysteroid agonist and antagonist activity of HHCB was assessed in an assay with the **Drosophila melanogaster** BII-cell line. HHCB did not show specific agonistic or antagonistic activity in this bioassay (Breitholtz et al., 2003).

HHCB inhibited multixenobiotic resistance (mxr) tra nsporters in gills of the marine mussel, **Mytilus californianus** , an effect that may increase the accumulation of o ther toxicants in the tissues (Luckenbach et al., 2004). The inhibitory e ffects of HHCB were only partially reversed after a recovery period in clean seawater (Luckenba ch and Epel, 2005)

**.**

## Appendix G ENVIRONMENTAL MONITORING DATA ANALYSIS

G-1 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 con centration of HHCB in influent is dependent on the source of waste received (municipa l, commercial, industrial) and the population served by the WWTP. The concentrations o f fragrance materials in effluent, however, are dependent upon the ability of these co mpounds to be eliminated during treatment as well as WWTP size, type, and process.

In the draft report of the Science Advisory Panel f or Chemicals of Emerging Concern in California’s Aquatic Ecosystems, the maximum concen tration of HHCB in wastewater effluent reported from within the state of California was 2. 78 μg/L (Anderson et al., 2012). The maximum aqueous concentration of HHCB in treated mu nicipal wastewater effluent discharged to the ocean was 2.5 μg/L. Details regarding the ex traction methods and individual QA/QC procedures for the specific studies were not provid ed in this summary document.

Quarterly samples of wastewater influent and efflue nt were taken over a one year period at two WWTPs in Texas, and synthetic musk fragrance co ncentration was measured (Chase et al., 2012). HHCB was consistently one of the more abunda nt synthetic musk fragrances present throughout quarterly sampling, though seven other m usk fragrances were also detected. The
authors concluded that the low concentration of the other musk fragrances, as compared to HHCB, could be due in part to a low influx of those compounds into the WWTP or sampling time. From samples prepared using solid phase extra ction methods, the influent concentration of HHCB was as high as 5.7 μg/L and the effluent co ncentration was as high as 6.1 μg/L as shown in Table G-1. Recoveries were reported to be consistently over 50%. The method detection limit was 0.004 μg/L and the method quant itation limit was 0.040 μg/L. Where synthetic musk fragrances were detected in blank sa mples, the amount present was below the
calculated method detection limit.

A state-wide survey in Oregon of trace metals and o rganic chemicals in municipal effluent was published in 2012 (Hope et al., 2012). Oregon's Senate Bill 737, enacted in 2007, required the state's 52 largest municipal WWTPs and water pollut ion control facilities to collect effluent samples in 2010 and analyze them for persistent org anic pollutants. These facilities are located state-wide and represent a variety of treatment typ es, service population sizes, geographic areas, and flow conditions. HHCB was detected in 2/ 102 (the reported limit of quantitation was 10 μg/L) samples with a median value of 12.5 μg/L a nd a maximum value of 13.5 μg/L. The samples were filtered and analyzed using mass spect rometry by directly injecting the sample without extraction. Each sampling batch included a laboratory method blank and a laboratory control sample, with the data reported as estimated if the sample result did not exceed ten
times the level in the blank.Wastewater influent and effluent concentrations wer e determined at two WWTPs, one located

in a rural area in Kentucky and the other from an u rban site in Georgia (Horii et al., 2007). These two WWTPs treat primarily domestic and commer cial 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 μg/L) were up to six times higher than the mean concentrations for the Georgia plant (0.42 μg/L). However, mean
concentrations of HHCB in the effluents did not dif fer significantly between the two sites (0.044 and 0.055 μg/L, respectively). Liquid-liquid extraction was performed on the wast ewater samples and the concentration of HHCB was determine d by GC/MS. Procedural blanks analyzed with the samples as a check for contamination durin g analysis did not reveal the presence of the target polycyclic musks. The limit of quantificatio n was reported as 10 ng/L and the average recovery for HHCB was 87 ± 4%.

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 iden tical treatment processes and serve cities of moderate population size (approximately 100,000 people) (Reiner et al., 2007a). Both WWTPs receive domestic and commercial discharge, an d although one also receives industrial discharge, the measured concentration of HHCB did n ot differ. HHCB was found in wastewater samples with influent and effluent concentrations i n the ranges of 4.76 to 12.7 and 0.010 to 0.098 μg/L, respectively. Liquid-liquid extraction was performed on the waste water samples and the concentration of HHCB was determined by GC/MS. Field blanks and procedural blanks were run with samples as a check for contamination and t he limit of quantitation, 10 ng/L, was set to
be 10 times the standard deviation found in the bla nks. The average recoveries for HHCB were

85 ± 4.3%.

WWTP effluent was sampled at two sites over a two-y ear period: Taylors Falls WWTP at Taylors Falls, Minnesota and St. Croix Falls WWTP at St. Cr oix Falls, Wisconsin (USGS, 2011). The sites were chosen because the WWTPs discharge into the St . Croix River upstream from endangered mussel populations. Mean concentrations of HHCB ove r the sampling period were 0.048 μg/L at Taylors Falls and 1.33 μg/L at St. Croix Falls. Water samples then were stored on ice and extracted within 48 hours of sampling using disposa ble, polypropylene solid-phase extraction (SPE) cartridges. Field and laboratory blank samples were analyzed to ensure that environmental samples were not contaminated during collection and processing and to assess potential contamination. The average concentration in blank samples was reported as 0.012 μg/L. Recovery with spiked samples was 84.4 with a standard deviation of 22.1%.

Osemwengie and Gerstenberger (2004) analyzed surfac e water from the confluence of three municipal sewage treatment effluent streams in the state of Nevada, for a period of 7 to 12 months. HHCB was consistently detected in higher monthly average concentrations (0.032 to 0.098 μg/L) relative to the other polycyclic musks. The sewage effluents were sampled from a dedicated effluent receiving stream (one that recei ves only sewage effluent and rarely runoff). Twice filtered water was drawn by and passed throug h a peristaltic or diaphragm pump and finally through a cartridge containing a 1:1 polyme thyl methacrylate:polystyrene crossed linked with 50% divinylbenzene sorbent. Fragrance free soa ps were used during extraction andanalysis. Prepared field blanks and laboratory bla nks were used to check for contamination.

The method reporting limit for HHCB was 0.02 ng/L b ased on a signal-to-noise ratio of 3 to 1. The average recovery was 97% for STP effluent and 9 9% for nanopure water.

Smyth et al

**.** (2008) measured the HHCB concentration in the infl uent and effluent over a oneyear period at six different WWTPs, located in Onta rio Canada, employing four different treatment processes. Influent concentrations were a s high as 40.3 μg/L and effluent concentrations were as high as 3.73 μg/L. Samples w ere subjected to liquid-liquid extraction with petroleum ether and cleaned up on deactivated silica gel. Concentrations of HHCB in field blanks were 2-3 orders of magnitude below concentra tions measured in wastewater influent and 1-2 orders of magnitude below concentrations me asured in effluent, therefore background contamination was considered negligible. Mean recovery of a deuterated surrogate (anthracene or phenanthrene) was 93% with a standard deviation of 23%. The MDL was reported as 0.011 μg/L. Lagoon treatment produc ed the lowest effluent concentration, although the authors noted that process temperature may have influenced the removal efficiency. Three WWTPs that utilize conventional a ctivated sludge processes were studied, with only small differences in effluent concentrati ons observed. Overall, effluent HHCB concentrations were an order of magnitude less than the influent HHCB concentrations, as
summarized in Table G-1.

A limited number of studies have evaluated the diff erences between different types of wastewater treatment and their ability to remove HH CB. A study by Simonich et al. (2002) of six wastewater treatment processes (activated sludg e, carousel, oxidation ditch, trickling filter, rotating biological contactor, and lagoon) in the U S and Europe concluded that overall removal (primary and secondary treatment) of fragrance mate rials ranged from 58.6-99.9%, with the lowest removal occurring at carousel plants. The hi ghest removal, 96.7-99.9%, was for lagoon plants. In Canada, Smyth’s 2008 study (Smyth et al ., 2008) of fragrance material removal by four WWTP processes (activated sludge, extended aer ation, oxidation ditch, lagoon) similarly found that lagoon treatment yielded the highest rem oval (>95%). Both authors suggested that the long retention times associated with lagoon tre atment is a likely factor in the high removal values, but suggest that other factors such as the ability of given treatment plant to nitrify may have a greater influence on fragrance material remo val than retention time alone.Table_Apx G

- 1

. Measured Concentrations of HHCB in Wastewater

Location Year n Influent

Concentration (μg/L) Effluent

Concentration (μg/L) Comments Reference

CA 2007

— a

— b 2.78 maximum Effluent dominated inland
waterway Anderson et al. (2012)

ND to 2.5
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 12.5 median;

13.0 maximum 52 municipal WWTPs

sampled Hope et al. (2012)

KY 2005

— a 0.043

- 7.032

2.499 mean 0.017

- 0.024;

0.021 mean 0.010-0.037;

0.044 mean Before chlorination Horii et al. (2007)

GA 0.284

- 0.522,

0.420 mean 0.051

- 0.418;

0.189 mean 0.028-0.098;
0.055 mean Primary

NY(a) 2005

— a 4.76

- 12.7

9.030 mean 4.120

- 8.150;

3.730 mean 2.810-3.730;

3.350 mean Primary Reiner et al. (2007a)

NY(b) 2005 1.8

- 11.5

5.940 mean 2.950

- 6.820;

5.310 mean 2.360-3.310;

2.700 mean Primary

NV 2000

-

2001 24

— b 0.0326

- 0.0979;

0.0569 mean Osemwengie
and Gerstenberger (2004)

MN

WI 2007 7

— b 0.048 mean

1.330 mean Taylors Falls WWTP

St. Croix Falls WWTP USGS (2011)

Ontario,
Canada 2003

-

2004 3 7.75

- 14.6

8.38-14.0

1.81-7.21

5.03-40.30

6.64-10.8

3.78-19.0 0.0546

- 0.673

2.477-3.730

0.497-1.130

2.290-3.240

2.230-3.340

0.914-1.680 Lagoon

Extended aeration
Oxidation ditch Activated sludge Activated sludge Activated sludge Smyth et al. (2008)

a Not reported.

b Not measured. ND: not detected, detection limits not specified.

### G-2 Measured Concentrations in Surface Water

Because synthetic musk fragrances, such as HHCB, ar e used in many consumer products, these compounds enter WWTPs through down-the-drain practi ces. The ability of fragrance materials to be eliminated during treatment depends on a vari ety of factors inherent to the WWTP, such as WWTP size and type (municipal or industrial), pr ocesses utilized to treat waste, and populations served (rural versus urban). Consequen tly, 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 disch arge areas, with peak environmental concentrations typically occurring near effluent di scharge points (Heberer et al., 1999; Peck et al., 2006). Although the concentration of HHCB in w astewater and surface water may vary, HHCB is continuously present and constitutes a cons tant exposure source.

A science advisory panel convened by the Southern C alifornia Coastal Water Research Project was charged to identify potential sources of chemic als of emerging concern, evaluate their fate and effects, and ultimately provide guidance for de veloping monitoring programs that assess those chemicals with the highest potential to cause effects in the state’s receiving waters (Anderson et al., 2012). All data were blank censor ed to ensure that the reported compounds were in the sample at the time of collection and no t artifacts of sample processing and analysis. The established reporting level for HHCB was 0.5 μg /L. Three scenarios were evaluated: Scenario 1, a WWTP effluent-dominated inland (fresh water) waterway; Scenario 2, a coastal embayment that receives both WWTP effluent and stor mwater discharge; and Scenario 3, offshore ocean discharge of WWTP effluent. The maxi mum values reported for each of these
scenarios are recorded in Tables G-1 and G-2.

Water samples were taken from the North Side water reclamation plant effluent and the North Shore Channel of the North Branch of the Chicago Ri ver, both located in the Chicago, Illinois metropolitan area (Barber et al., 2011). This plan t collects sewage from residential (96 percent) and commercial/industrial (3.4 percent) sources. Fo llowing treatment, the water reclamation plant effluent is discharged to the North Shore Cha nnel. There were no statistical differences in concentrations of HHCB in these two locations, indi cating minimal in-stream dilution or transformation. The mean concentration was reported as 1.6 μg/L. Glass fiber filtered samples were extracted using polystyrene divinylbenzene sol id phase extraction and eluted with methylene chloride. Field blanks, laboratory blanks and distilled water matrix spikes were analyzed for qualtiy assurance, however, specific d etails regarding laboratory reporting limits for the water samples were not provided.

HHCB was detected in the surface waters of the Cany on Lakes System and North Fork Brazos River near Lubbock, Texas in 2010 (Chase et al., 2012). Reported concentrations in the lake ranged from ND to trace amounts, whereas concentrat ions in the river were appreciably higher, ranging from 0.077 to 0.794 μg/L. The low lake conc entrations were expected to be similar to groundwater since the source was groundwater from a land applied site, whereas occurrence in
the stream water was attributed to direct release o f WWTP effluent. The semi permeable membrane device used for sampling was dialyzed into hexane and reduced in volume prior toanalysis by GC/MS. Recoveries were reported to be c onsistently over 50%. The method

detection limit was 0.001 μg/L and the method quant itation limit was 0.005 μg/L. Where synthetic musk fragrances were detected in blank sa mples, the amount present was below the
calculated method detection limit.

Peck and Hornbuckle (2004) characterized the concentrati on of several synthetic musk fragrances in western Lake Michigan with water samp les that were collected from 1999 to 2001. A lake-wide annual mass balance analysis show ed that WWTP discharge was the major source of the synthetic musks, whereas major loss m echanisms were outflow and volatilization. The average measured concentration of HHCB over thi s time period was 4.7 ng/L. The resin and glass fiber filters used in water sampling were ext racted with acetone/hexane. Separations were performed with hexane, dichloromethane/hexane and methanol followed by analysis using GC/MS techniques. The average recovery of flu oranthene- **d 10**  was 56±34%. Field and laboratory blanks were utilized with the average sa mple to blank ratio for the dissolved–phase listed as 7.9. The limit of detection was not repor ted.

Water concentrations were measured in 2006 for HHCB from three locations along the upper Hudson River, with mean concentrations for HHCB ran ging from 0.00395 to 0.0251 μg/L (Reiner and Kannan, 2011). The highest concentration was fo und near Albany, New York, which corresponded to the location with the highest popul ation. Liquid-liquid extraction with hexane and dichloromethane was performed prior to analysis by GC/MS. Average recoveries of HHCB ranged from 85-98%, with the standard deviation bel ow 15% for all analytes. Field blanks and procedural blanks were analyzed with the samples as a check for contamination. The limit of quantitation for HHCB was 0.001 μg/L.

The USGS (USGS, 2011) and the National Park Service cooperated on a study to determine the
occurrence of wastewater indicators in the St. Croi x National Scenic Riverway in Minnesota and Wisconsin. Samples of treated wastewater effluent f rom two WWTPs, located in Taylors Falls, Minnesota and St. Croix Falls, Wisconsin were colle cted from 2007 to 2008, with concentrations up to 1.33 μg/L; however HHCB in surface water samp led from the St. Croix River at two locations in MN was not detected. Semi-permeable me mbrane devices (SPMD) were used in conjunction with polar organic chemical integrative samplers (POCIS). Analytes were extracted from SPMDs by means of a two stage dialysis with a solvent. Analytes were separated from the POCIS extracts by means of a liquid chromatography gradient elution. The concentration of HHCB in the field blank sample collected at the St. Croix River above Rock Island near Franconia, MN was 0.012 μg/L. The method reporting limit for H HCB was listed as “0.5e” where **e** denotes variable performance during initial method validati on.

Klecka et al., 2010 in a report prepared to a multi -board work group, summarized studies on chemicals of emerging concern that could pose threa ts to water quality in the Great lakes
watershed. The mean value of HHCB in surface water was 0.008 μg/L from 50 samples with a detection frequency of 70%. Specific information r egarding detection limits and analytical methodologies was not provided as the report provid ed summary statistics only.Concentrations of HHCB were detected in low concent rations in Lake Mead, with a mean value

of 0.00036 ± 0.0003 μg/L in 2001, although this wat erbody is downstream of the confluence of three WWTPs in Las Vegas, Nevada (Osemwengie and Ge rstenberger, 2004). The low concentrations of HHCB, compared to the concentrati ons 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. Twice filtered water was drawn by and passed through a pe ristaltic or diaphragm pump and finally through a cartridge containing a 1:1 polymethyl met hacrylate:polystyrene crossed linked with 50% divinylbenzene sorbent. The low concentrations of the analytes in the Lake Mead extracts were analyzed using GC/MS in the selected-ion mode. Fragrance free soaps were used during extraction and analysis. Prepared field blanks and laboratory blanks were used to check for contamination. The method reporting limit for HHCB was 0.02 ng/L based on a signal-to-noise ratio of 3 to 1. The average recovery was 97% for S TP effluent and 99% for nanopure water.

Passive sampling devices were deployed within two n earshore study zones along the north shore of Lake Ontario, Canada near the municipaliti es of Pickering and Ajax, Ontario, and Port Hope, Ontario (Helm et al., 2012). Time-weighted average dissolved wate r concentrations were estimated based on chemical amounts sequestered in the device. Organic wastewater chemicals, like HHCB, were greatest in the vicinity of the Pickering-Ajax WWTP, and then dramatically declined. Calculated sampling device-b ased concentrations of HHCB ranged from approximately 0.0002 to 0.010 μg/L near Pickering-A jax and from 0.1 to 0.3 ng/L near Port Hope. Recovery of HHCB was >79% as determined by ex tracting passive sampling devices spiked with the target compound. Measured concentra tions in the laboratory and field blanks were 5 to >10 times lower than the samples.

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 results were obtained from two replicate extractions of each sample by means of li quid–liquid extraction. The measured HHCB concentration within the bay of Hamilton (highly in fluenced by WWTP effluents) was 0.007 μg/L. With increasing distance from Hamilton Harbor , the measured concentration decreased and stabilized at 0.002 μg/L. The authors indicated that the decrease of the measured concentrations might be traced back to dilution eff ects or that musk fragrances are stable under the conditions found in Lake Ontario; however , they concluded that further investigation
is needed.Table_Apx G

- 2

. Measured Concentrations of HHCB in Surface Water

Location Date n Concentration (μg/L) Comments Reference

CA 2008 — a 0.970 maximum Resevoir Anderson et al. (2012) Chicago, IL 2006-2007 3 1.6 mean North Shore Channe l of the Chicago River (Impacted by water reclamation plant effluent) Barber et al. (2011)

Lubbock, TX Fall 2010

Spring 2010 — a 0.045

0.077-0.794 Canyon Lakes System

North Fork Brazos River Chase et al. (2012) Michigan 1999-2001 13 0.0047 ±0.0025 Lake Michigan, background locations Peck and
Hornbuckle (2004) Catskill, NY Troy, NY Albany, NY 7/2006

5/2006 5/2006 2

2

2 0.00395 mean

0.00736 mean

0.0251 mean Upper Hudson River Reiner and

Kannan (2011) Taylors Falls, MN St. Croix Falls, WI Franconia, MN Sunrise, MN 2007-2008 8 0.048

1.33

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 0.00006-0.001,

0.00036 mean Lake Mead Osemwengie
and Gerstenberg er (2004) Great Lakes 50 0.00008-0.180; 0.0008 median; 0.00844 mean;

0.027 95 th %ile Great lakes watershed, summary of

studies Klecka et al. (2010)

Lake Ontario,
Canada June 2008 — a 0.00012- 0.00034
estimated

0.00021-0.010
estimated Port Hope

Pickering-Ajax Helm et al. (2012)

Lake Ontario,
Canada 2005 4 0.007 Hamilton Harbour Andresen et al. (2007)

a Not reported. ND: not detected (detection range =0.038-2.16 μg/L )



### G-3 Measured Concentrations in Sediment

Studies of measured environmental concentrations of fragrance compounds in the sediment in North America are limited. Synthetic musk fragrance s 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.Chase et al. (2012) measured the concentration of HHCB in the sediment of a playa lake, a noneffluent impacted study site located in proximity t o an area treated by a WWTP. The primary water sources are rain and urban runoff, as these p layas are integral for storm water retention. In this study, sampling was performed in the summer and fall timeframes of 2010. The
measured concentrations of HHCB were 2.13 and 1.43 μg/kg dw in the summer and fall of 2010, respectively, as shown in Table G-3. Samples were e xtracted with 1:1 acetone/hexane. Recoveries were reported to be consistently over 50 %. The method detection limit was 0.3 μg/kg and the method quantitation limit was 0.33 μg /kg. Where synthetic musk fragrances were detected in blank samples, the amount present was below the calculated method
detection limit.

The USGS investigated the occurrence and distributi on of organic waste contaminants in the
Tinkers Creek watershed and two other tributaries t o the Cuyahoga River in northeastern Ohio (USGS, 2008; also summarized in Klecka **et al.,** 2010). The effluent from WWTPs constitutes a continuous and sometimes high proportion of the flo w in Tinkers Creek and its tributaries (sometimes ≥80 percent). HHCB was detected in the s treambed sediment in 29 percent of upstream sites and 86 percent of downstream sites. The range of values is 20 to 390 μg/kg dw with an estimated median value of 60 μg/kg dw. The reported upstream concentrations were <60 μg/kg dw, whereas the downstream concentrations ranged from 20 (estimated value) to 390 μg/kg dw with three of the seven downstream sam ples having measured values >200 μg/kg dw. Sediment samples were extracted using a pressur ized water/isopropyl alcohol extraction. For HHCB, the long term method detection level was reported as 16.5 μg/kg and the laboratory reporting level was listed as 50 μg/kg. Method and replicate blanks were used to check for

contamination.

Water and sediment samples were collected in May an d July 2006 at three locations along the upper Hudson River: Troy, Albany, and Catskill, Ne w York (Reiner and Kannan, 2011). WWTPs in each of these New York cities discharge effluent s into the Hudson River. HHCB was detected in all sediments sampled. Concentrations of HHCB w ere 72.8 μg/kg dw from Catskill, 351 μg/kg dw from Albany, and 388 μg/kg dw from Troy. Samples were extracted with a 1:3 hexane and dichloromethane mixture. Average recoveries of HHC B ranged from 85-98%, with the standard deviation below 15% for all analytes. Field blanks and procedural blanks were analyzed with the samples as a check for contamination. The limit of quantitation for HHCB in sediment was 5 μg/kg. The authors noted that these sediment conc entrations 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 th e Chesapeake Bay, US (Magothy, Corsica, and Rhode Rivers) were collected in 2007 (Sapozhnik ova et al., 2010). The Rhode River is the only one that receives effluent from a WWTP. Concen trations of HHCB detected in sediments were as high as 9.2 μg/kg dw for the Magothy River, 2.3 μg/kg dw for the Rhode River, and 1.6 μg/kg dw for the Corsica River. The reported av erage HHCB concentration was 1.1±2.2 μg/kg dw. The Magothy River is the most urbanized r iver among the three tributaries studied
and was reflected in the concentration of HHCB foun d in the sampling. Thirty-nine total samples were collected, but details regarding the n umber of samples for each location were notprovided. Samples were extracted with a 1:1 mixture of methylene chloride and acetone prior

to GC/MS analysis. Reagent blanks and replicate sam ples were analyzed with each batch of samples and blank subtracted as needed. HHCB recove ries were 71% and recoveries in blank samples spiked with musk standards were 96±12%. The limit of detection was not specified.

Sediment samples from Lake Erie and Lake Ontario we re measured in August of 2003 (Peck **et al.**  , 2006). The concentration in the Lake Erie core sa mple were blank corrected using the average mass of each compound found in four method blanks. Sample extraction was performed with dichloromethane and exchanged into h exane. The LOD for the core was
defined as 3 times the standard deviation. The LOD in the Lake Ontario core was 5.1 μg/kg. The HHCB concentration measured in Lake Ontario (16 μg/ kg dw) was an order of magnitude higher than that of Lake Erie (3.2 μg/kg dw). This differe nce 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 conc entration of HHCB over time was extrapolated from the collected sediment core s amples and the HHCB concentration was correlated to US fragrance consumption from 1980 to 2006. The authors concluded that HHCB concentration found in the sediment samples reflect ed the trend of fragrance consumption over this time period.

Table_Apx G

- 3

. Measured Sediment Concentrations at Locations in th e US

Location Year Concentration μg/kg dw Comments Reference

Lubbock, TX Summer 2010

Fall 2010 2.13

1.43 Canyon Lake System Chase et al. (2012) Tinkers Creek, OH Cuyahoga River, OH 2006 20 a

- <60, 37 average

20 a -390, 144 mean Upstream of 7 WWTPs

Downstream of 7 WWTPs USGS (20 08 );
Klecka et al. (2010) Troy, NY Albany, NY Catskill, NY 2006 388

351

72.8 Upper Hudson River, NY Reiner and Kannan (2011)

Magothy River, MD Rhode River, MD Corsica River, MD 2007 ND

- 9.2

ND - 2.3

ND - 1.6 1.1 avg Tributary, Chesapeake Bay Sapozhnikova et al. (2010)

Lake Erie
Lake Ontario 2003 3.2

16 Great Lakes Peck et al. (2006 )

a Reported as an estimated value. ND = not detected, limit of detection not specified



### G-4 Measured Concentrations in Biosolids and Sludge

The removal of fragrance compounds from wastewater has been reported in the US, Canada, and EU in several studies, with overall removal eff iciencies ranging from as low as 29.9 percentto as high as 99.7 percent (typical values were 70 to 80 percent) depending upon the type of

wastewater treatment (Simonich et al., 2002; Deblon de et al., 2011; Smyth et al., 2008). A fraction of HHCB is expected to preferentially part ition to organic-rich biosolids during wastewater treatment based on the high measured log K OW values of 5.3 to 5.9 (see Table 2-1)
and the estimated K OC values of 4.1 to 4.3 (EPI Suite v4.11, see Chapter 2, Section C, Environmental Fate). These solids, produced through the process of wastewater treatment, are subsequently disposed of by either landfill or inci neration. They may also be utilized for land application to fertilize plants or to improve the q uality of soil, and may then become available for plant and terrestrial biota uptake.

In a study by Kinney et al. (2006), the presence an d concentration of organic waste contaminants was determined in solid materials prod uced during wastewater treatment. Nine different biosolid products from municipal WWTPs in seven different states were analyzed. Six of these products were intended for commercial, hom eowner, and municipal use and are available to the public for purchase. Three of the products are used in agriculture. HHCB was
detected in all nine biosolids. HHCB was extracted from the biosolid using accelerated solvent extraction and measured by GC/MS. Each sample was c omposited from bulk material and samples were analyzed in triplicate. All samples ha d similar treatment (secondary activated sludge and chlorine disinfection); however, the pre paration and applied form of the solid varied. The concentration of HHCB varied, with one site sample measuring as high as 177,000 μg/kg dw. All other measured HHCB concentra tions were below 4,000 μg/kg dw. As this high value was more than 40 times higher than the next highest measured concentration, and more than four orders of magnitude higher than the lowest reported concentration, the authors considered it as an “outlier.” Excluding th is value, the reported median value for HHCB was 1,461 μg/kg dw. Extractions were performed usin g accelerated solvent extraction wand a
70:30 acetonitrile: water solvent mixture. At least one laboratory spike and laboratory blank were evaluated for each set of extractions and quan tifications; however HHCB was not detected in any of the blanks hence no contaminatio n was indicated. The method detection limit was noted as 16.5 μg/kg and the recoveries ra nged from 54-107%. Differences in biosolid preparation techniques and WWTP characteristics ( **e.g.** , location, population served, volumes of wastewater, and retention times) likely account for the wide variation in HHCB concentrations measured, with no one factor being clearly delineat ed in this study.

A second study by Kinney et al. (2008) was performed to examine bioaccumulat ion in earthworms from soil obtained from three Midwest ag ricultural fields, which had been
amended with biosolids or swine manure. The biosoli d source had a concentration of 427,000 μg/kg dw, although details regarding this s ource were not provided. HHCB and tonalide were detected in one of the blank samples at a conc entration 1 to 3 orders of magnitude lower than reported values. The method detection limit fo r HHCB was reported as 12.5 μg/kg dry weight.

Biosolid samples were collected from the North City Wastewater Reclamation Plant of the City of San Diego to study biosolid degradation on a wee kly schedule for 19 weeks (Buyuksonmez and Sekeroglu, 2005). Samples were taken from the r eturn activated sludge line after purgingthe stagnant sludge from the sampling line. The pre sence of HHCB was determined using

GC/MS following extraction (sequentially with methy lene chloride, ethyl acetate and hexane) with a Soxhlet apparatus for 24 hours. The average concentration of HHCB in biosolids was reported to be 7,840 μg/kg dw, although the number of samples taken and information regarding blanks and the limit of detection were no t provided.

In a study by DiFrancesco et al

**.** (2004), digested sludge was collected from two sit es over a period of two years. Anaerobically digested and de watered sludge was obtained from two activated sludge plants in Delaware: Georgetown WW TP and Wilmington WWTP. Background concentrations of fragrance materials in anaerobica lly digested sludge from the municipal WWTPs were determined in July 2000 and again in Mar ch (Georgetown) and October (Wilmington) 2002. Sludge was dewatered using a gra vity separator at Georgetown WWTP and a belt filter press at Wilmington WWTP. Both WWTPs added polymers to assist in sludge dewatering, and the Georgetown facility also used a luminum chloride for phosphorus removal. Samples were extracted using accelerated solvent ex traction (dichloromethane) and were analyzed using GC/MS. The HHCB concentrations range d from 21,800 to 86,000 μg/kg dw. The overall concentrations in this limited sampling wer e higher in samples collected in 2000 and at the Georgetown site. Table G-4 includes a summary o f measured concentrations of HHCB in biosolids. Only data with all matching confirmation ions and a signal-to-noise ratio greater than 5 were accepted.

For comparison, studies of biosolids from Canada ar e also included in Table G-4. HHCB concentrations in digested biosolid were studied by Smyth et al. (2007a, 2007b) in Canada. One oxidation ditch with tertiary sand filtration, one extended aeration activated sludge plant, and three conventional activated sludge plants that dis charge into the Grand River, Ontario, Canada were used as sampling sites with four sampling even ts from November 2003 to July 2004. Extraction was performed using supercritical fluid (5% acetone in hexane) or microwave assisted techniques (acetone and hexane), measureme nt was performed using GC/MS. Method blanks were employed to confirm the absence of cont amination. 3 out of 15 blanks showed concentrations of HHCB that were 2-3 orders of magn itude below concentrations measured in sludge samples. The method detection limit was repo rted as 41 μg/kg. HHCB concentrations in digested biosolid ranged from 9,430 to 55,500 μg/kg dw. Smyth et al. (2007a) reported that these values were in range with previous reported v alues in Canada (Lee et al., 2003) and the United Kingdom (Stevens et al., 2003).

Yang and Metcalfe (2006) also studied biosolids fro m WWTPs in Ontario, Canada. Samples were collected from the Peterborough WWTP over a one-yea r period, and the HHCB concentrations were measured in the raw, return activated sludge, and digested biosolids. Concentrations of HHCB did not vary significantly for any one samplin g type over the one-year study. However, there was a trend for the synthetic musks to accumu late in the return activated sludge and to finally concentrate in the digested biosolids. The annual average concentrations of HHCB were 3,309.9 μg/kg dw in the return activated sludge and 6,788.4 μg/kg dw in the digested biosolid. Extractions were performed by accelerated solvent e xtraction using a 1:1 mixture of **n** -hexane and ethyl acetate. Field blank and method blank samples were analyzed and recovery studieswere conducted in triplicate with blank subtraction . All recoveries of synthetic musks were

>80%. The limit of quantitation was determined from spiking experiments into surrogate matrices and corresponded to a 5:1 signal-to-noise ratio. Method LOQs were reported to vary between 0.4 and 4.0 ng/L for analytes in sewage sam ples; the specific value for HHCB was not reported.

Table_Apx G-4. Measured Concentrations of HHCB in Biosolids and Sludge

Location Collection

Year N Concentration

μg/kg dw Preparation/Comments Reference

WI 5/2003

11/2003 3

3 1,400 (average) 1,470 (average) Rotary kiln dried Kinney et al. (2006)

CO 5/2003

11/2003 3

3 3,700 (average) 2,670 (average) Composted + plant material

TX 8/2003

11/2003 3

3 1,100 (average) 977 (average) Composted + yard material

WA 11/2003 3 933 (average) Composted with sawdust

WA 5/2004 3 13 (average) Composted with yard waste

KS 9/2003

11/2003 3

3 2,370 (average) 1,970 (average) Air dried and turned

AZ 6/2003

1/2004 3

3 1,300 (average) 697 (average) biosolid

AZ 6/2003

1/2004 3

3 1,100 (average) 767 (average) Dewatered biosolid + polymer

IA 4/2005 3 177,000 (average) Thermophilic and mesophillic digestion + polymer

Unknown 2008 3 427,000 b No details given Kinney et al. (2008)

San Die go, 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

Ontario,
Canada (5 locations) 2003

-

2004 7

10

12

11

11 9,430 (mean) 40,300 (mean) 42,000 (mean) 55,500 (mean) 46,300 (mean) Aerobically digested biosolids

Aerobically digested biosolids Anaerobically digested biosolids Anaerobically digested biosolids Anaerobically digested biosolids S myth et al. (2007a)

Canada (site 1/site 2) 2003

-

2005 — a 36,500/20,100

25,600/17,500 73,700/45,300 23,800/20,500 Primary sludge, wet

Primary sludge, air dried Waste activated sludge, wet Waste activated sludge, air dried Smyth et al. (2007b)Table_Apx G-4. Measured Concentrations of HHCB in Biosolids and Sludge

Location Collection

Year N Concentration

μg/kg dw Preparation/Comments Reference

Canada Peterborough,

ON 2/2003

4/2003

7/2003

10/2003

Annual

average 1

3

3

3

3

3
3

3

3

3

3
3 4,514.4 3,638.8 (± 787.7) 7,896.7 (± 2,130.8) 2,857.4 (± 18.6) 5,103.9 (± 635.0) 7,006.8 (± 931.3) 3,356.1 (± 1,058.6) 2,156.6 (± 504.7) 6,477.6 (± 1,065.3) 2,482.4 (± 340.6) 2,340.4 (± 360.2) 5,772.76 (± 115.2) 3,302.5 3,309.9 6,788.4 469.6 (± 22.4) Raw sludge

Return activated sludge Digested sludge Raw sludge Return activated sludge Digested sludge Raw sludge Return activated sludge Digested sludge Raw sludge Return activated sludge Digested sludge Raw sludge Return activated sludge Digested sludge
Trucked biosolid Yang and Metcalfe (2006)

a Not reported.

b Value was not reported as mean, average, or median.

Studies summarized in Table G-4 reported analytical reliability using accelerated solvent extraction or supercritical fluid extraction follow ed by GC/MS techniques. Studies of HHCB in biosolids are somewhat limited, however, because th e studies that detected and measured HHCB reported concentrations in the range of <100 t o >100,000 μg/kg dw. The measured concentration depended on a variety of factors, inc luding WWTP operations such as receiving waste stream (municipal or industrial), date of col lection, location, population served, water volume, treatment type, and preparation methodologi es prior to land application. Because these factors cannot be independently separated fro m values within the reported data set, it is not possible to assign sensitivity to a single fact or. Comparison with values reported for studies in Ontario, Canada are within the reported range fo r those from the US and also similarly show differences depending upon location, treatment proc ess, and season.



### G-5 Measured Concentrations in Soil

Limited data were available that described concentr ations of HHCB in soil. Though many studies are designed to understanding the biodegradation or fate of HHCB in the soil compartment, there are data showing that land application of was tewater effluent or amendment of soils with
biosolids resulted in measurable concentrations of HHCB. The high frequency of HHCB detection in biosolids and wastewater suggests that they may be an important organic waste contaminant source to terrestrial environments, par ticularly when applied to soil. Table G-5
summarizes the measured concentrations in HHCB in s oil.In 2008, Kinney et al. (2008) reported HHCB concentrations in agric ultural soil amended with

biosolids. Three sites were studied. Site 1 (mini mally affected site) was a non-irrigated soybean
field that had not been amended with either human o r livestock waste for at least the previous seven years. Soil and earthworm samples were collec ted from this field in June and September of 2005. Site 2 (biosolid amended site) was a no-ti ll, non-irrigated soybean field receiving
biosolid as a fertilizer for the first time. Site 3 (manure-amended site) was a non-irrigated cornfield receiving liquid swine manure as an organ ic fertilizer. While HHCB was not detected at site 3, HHCB was detected at site 1 in June 2005 at measurable levels, up to 633 μg/kg dw. HHCB concentrations were highest at site 2 and rang ed from 1,050 to 2,770 μg/kg dw. Extractions were performed with a 70:30 acetonitril e: water solvent mixture. HHCB and tonalide were detected in one of the blank samples at a concentration 1 to 3 orders of magnitude lower than reported values. The method de tection limit for HHCB was reported as 12.5 μg/kg dry weight.

In a study by Chase et al. (2012), soil core sample s were collected quarterly over a two-year period at sites outside the city limits of Lubbock, Texas to determine the amount of sorption after land application of treated wastewater and gr oundwater recharge. The wastewater at this site is treated through primary and secondary treat ment processes for removal of suspended solids and organic carbon through different operati ng flow streams at the WWTP. Some of the treated effluent is released to a secondary effluen t reservoir, which is then applied **via** center pivot irrigation at a land application site. Two si tes were studied both inside and outside the pivot irrigation system; however, small differences in measured HHCB concentrations were observed between the two locations. Three sample de pths were studied: 0 to 6, 12 to 18, and 24 to 30 inches below the surface. Over the time pe riod of the study, at locations inside the irrigation system, concentrations ranged from <0.33 to 1.98 μg/kg dw at 0 to 6 inches, <0.33 to 1.25 μg/kg dw at 12 to 18 inches, and ND to 1.25 μg /kg dw at 24 to 30 inches. For locations outside the irrigation system, concentrations range d from <0.33 to 5.69 μg/kg dw at 0 to 6 inches, <0.33 to 0.77 μg/kg dw at 12 to 18 inches , and <0.33 to 0.79 μg/kg dw at 24 to 30 inches. The highest reported value (5.69 μg/kg dw) was measured outside the irrigation system in the winter of 2009; however, excepting this valu e, concentrations were below 1.98 μg/kg dw over the time period of the study. Soil concentrati ons were found to be lowest during the summer months, possibly due to volatilization. Extr actions were performed with 1:2 acetone/hexane. Recoveries were reported to be cons istently over 50%. The method detection limit was 0.3 μg/kg and the method quantitation lim it was 0.33 μg/kg. Where synthetic musk fragrances were detected in blank samples, the amou nt present was below the calculated method detection limit. Groundwater was also sample d and found to range from ND to <5 ng/L.Table_Apx G

- 5

. Measured Concentrations of HHCB in Soil

Location Year Site Soil Concentration (μg/kg dw) Re ference

Midwest, US Agricultural fields 2005 Minimally affected ND

- 633 Kinney **et al.**  (2008) Biosolid amended 1,050

- 2,770

Manure amended ND Lubbock, TX Land application

site 2009

-

2010 **0**

**- 6 inches**  **12**

**- 18 inches**  **24**

**- 30 inches**  Chase **et al.** (2012) Inside pivot irrigation <0.33

- 1.98 <0.33

- 1.25 ND

- 0.90 Outside pivot irrigation <0.33

- 5.69 <0.33

- 0.77 <0.33

- 0.79
Ontario, Canada 2003 Biosolid amended 2.0 ± 0.1 (after 1 day) 2.8 ± 0.4 (after 2 weeks) Yang and
Metcalfe (2006)

ND: not detected; method detection limit = 12.5 ug /kg dw (Kinney); 0.3 μg/kg dw (Chase)

A Canadian study of biosolid-amended agricultural f ields similarly showed that HHCB persisted in soils for the first two weeks post application, but concentrations declined thereafter (Yang and Metcalfe, 2006). Extractions were performed by accelerated solvent extraction using a 1:1
mixture of **n** -hexane and ethyl acetate. Field blank and method blank samples were analyzed and recovery studies were conducted in triplicate w ith blank subtraction. All recoveries of synthetic musks were >80%. The limit of quantitatio n was determined from spiking experiments into surrogate matrices and corresponde d to a 5:1 signal-to-noise ratio. Method LOQs were reported to vary between 0.2 to 1.9 μg/kg wet weight for analytes in soil; the specific value for HHCB was not reported.

More studies are needed, however, to understand the influence and integration of time, degradation mechanisms (such as volatilization, tra nsformation, and leaching), and organic matter content on long-term concentrations. Soil st udies, although clearly limited, do indicate that land application of treated wastewater effluen t or biosolids results in detectable quantities of HHCB in soil and hence represents a potential ro ute through which HHCB may enter terrestrial ecosystems.



### G-6 Measured Concentrations in Biota

The liver, fillet, and fat of many aquatic organism s, including fish, shrimp, zebra mussels, and aquatic mammals and birds, have been sampled from U S waters in a variety of locations (Table G-6). These data confirm the widespread occu rrence 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 expectedor detected in fish collected from this site. A tot al of 18 to 24 adult fish of the same resident

species were collected from each sampling location during late summer and fall of 2006. Fish sampled at each site were divided into six composit es, each containing three or four adult fish of similar size. The composites revealed the presen ce of HHCB at maximum concentrations ranging from 300 to 2,100 μg/kg tissue. Lipid deter minations were made gravimetrically and extractions were performed with 1:1 dichloromethane -hexane. Each analytical batch contained one blank, at least one continuing calibration veri fication sample and two laboratory control samples. No target pharmaceuticals were detected in blank samples. The method detection limit for HHCB was reported as 12 μg/kg tissue conc entration.

Concentrations of polycyclic musks, including HHCB, in fish were collected in 2006 from Troy, Albany, and Catskill along the upper Hudson River i n the eastern region of the state of New York (Reiner and Kannan, 2011). There are 148 WWTPs that discharge treated wastewater into the Hudson River, which flows southward from its he adwaters in the Adirondack Mountains and ultimately empties into the Atlantic Ocean at N ew York City. The measured levels of HHCB varied across fish species, with a range of <1 to 3 9 μg/kg lipid weight; the overall highest concentrations were measured in white perch liver. Biological samples were extracted with 3:1 dichloromethane-hexane. Average recoveries of HHCB ranged from 85-98%, with the standard deviation below 15% for all analytes. Field blanks and procedural blanks were analyzed with the samples as a check for contamination. The limit of quantitation for HHCB in fish samples was 1 μg/kg.

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 analyz ed for the presence of synthetic musks. HHCB was detected in all samples, with concentratio ns ranging from 48 to 683 μg/kg lipid weight (average 199 μg/kg lipid weight) in farmed s hrimp and 66 to 762 μg/kg lipid (334 ng/g lipid weight) in wild shrimp (HHCB in US wild shrim p max 330 μg/kg lipid weight, n=6). Farmraised shrimp from Indonesia (n = 1) showed the hig hest concentration of HHCB, followed by farm-raised shrimp from the US (max 424 μg/kg lipid weight, n = 3). No trends could be discerned as the sampling was not statistically sig nificant; however, the presence of HHCB in shrimp tissues, both farmed and wild caught, indica tes the widespread exposure of aquatic biota to HHCB. Lipid content was determined gravime trically and measured as dichloromethane extractible non-volatiles. Reagent blanks and replicate samples were analyzed with each batch of samples and blank subtr acted as needed. HHCB recoveries were 71% and recoveries in blank samples spiked with mus k standards were 96±12%. The limit of detection was not specified.

Kannan et al. (2005) reported the presence of HHCB in the tissues of higher trophic level aquatic organisms and humans (see Appendix A for su mmary of human biomonitoring studies). Among the wildlife species analyzed, tissue concent rations ranged from <1 to 25 μg/kg ww. Blubber tissue obtained from spinner and bottlenose dolphins from coastal waters off the state of Florida contained the highest levels of HHCB. Sa mple tissues were ground with anhydrous
sodium sulfate and extracted with mixed solvents of dichloromethane and hexane (3:1).Average recoveries of HHCB ranged from 85-98%. Proc edural blanks were analyzed with every

set of six samples to check for laboratory contamin ation and to correct sample values, if necessary. Procedural blanks contained trace levels of HHCB. The limit of quantitation (LOQ)
was set to be twice the concentration that was foun d in blanks. In wildlife samples, the LOQ for HHCB was 1 μg/kg wet weight.

Seven to eight carp were collected 200 meters from the drinking water intake area on Lake Mead, Nevada on a monthly basis over a period of on e year (Osemwengie and Gerstenberger, 2004). The measured HHCB concentration in the fish tissue ranged from 1.4 to 4.5 μg/kg ww with an annual mean concentration of 3.0 μg/kg ww. Lipids were extracted from carp tissue using chloroform. Prepared field blanks and laborat ory blanks were used to check for contamination. The method reporting limit for HHCB was not specified.

Kinney et al. (2008) measured HHCB concentrations in earth worm tissue obtained from amended soils where biosolids or liquid swine manur e was applied. These values were compared to values reported for worms collected fro m a soybean field, which had not been
amended with human or livestock waste for at least the previous seven years. HHCB concentrations in worms 30 days post application (3 ,340 μg/kg ww) were an order of magnitude higher than the levels measured after 156 days (131 μg/kg ww). Tissue
concentrations in worms from the manure amended sit e were 49 μg/kg after 139 days. Extractions were performed with a 70:30 acetonitril e: water solvent mixture. HHCB and tonalide were detected in one of the blank samples at a concentration 1 to 3 orders of magnitude lower than reported values. The mean ear thworm spike recovery was reported to be 40% and the MDL was 12.5 μg/kg dw.Table_Apx G

- 6

. Measured Concentrations of HHCB in Biota *

Species Year n Location HHCB
Concentration (μg/kg) Reference

μg/kg, lipid weight

White perch (liver) 2006
Hudso

n River 3 Troy, NY

Albany, NY Catskill, NY 6.27-19.9

7.58-22.5

13.7-27.9 Reiner and

Kannan (2011)

Channel catfish (liver) 3

1 Troy, NY Catskill, NY 11.1-39

<1 Smallmouth bass (liver) 3

3

3 Troy, NY Albany, NY Catskill, NY <1-11.1

1-31.9

<1 Largemouth bass (liver) 1

1 Albany, NY Catskill, NY 10.9

8.22 White catfish (liver) 1

1 Albany, NY Catskill, NY 6.56

5.79 Brown bullhead (liver) 3 Catskill, NY <1-51.1
Zebra mussel 4 Catskill, NY 10.3-19.3 American eel (whole body) 1 Catskill, NY 125
Shrimp, wild caught Shrimp, farm raised 2006 6

3 US

US 330 max

424 max Sapozhnikova et

al. (2010)

μg/kg, tissue concentration

Earthworm 2005

3

3

3 Midwestern US -minimally affected site
-biosolid amended site
-manure amended site 61 a ; ND b
3,340 c ; 131 d

ND e ; 49 f Kinney et al.

(2008)

Largemouth bass 2006 6 Chicago, IL
North Shore Channel 1,300 mean; 1,800 max Ramirez et al.

(2009)
Smallmouth buffalo fish 6 Dallas, TX Trinity River 800 mean; 1,800 max
Bowfin 6 Orlando, FL
Little Econlockhatchee

River 100 mean;

300 max

Common carp 6 Phoenix, AZ
Salt River 1,800 mean; 2,100 max
White sucker 6 West Chester, PA Taylor Run 1,800 mean; 2,000 max
μg/kg, wet weight

Sea otter (liver) 1993-

1999 8 Monterey Bay, CA <1-3.2 g Kannan et al. (2005)
Harbor seal (liver) 1996-

1997 3 Central CA coast 4.4-5.5; 4.8 mean

California sea lion (liver) 1993-

1996 3 Central CA coast 1.5-4.4; 2.8 mean

River otter (liver) 1997 3 Grand River, MI 2.4-3; 2 .8 meanTable_Apx G

- 6

. Measured Concentrations of HHCB in Biota *

Species Year n Location HHCB
Concentration (μg/kg) Reference

Bottlenose dolphin (blubber) 1994-

2000 4 FL coast 4.2-20.5; 12 mean

Striped dolphin (blubber) 1995-

1997 3 FL coast 8.1-25; 14 mean

Pygmy sperm whale (blubber) 2000 1 FL coast 6.6

Atlantic sharpnose shark (liver) 2004 3 Indian River Lagoon, FL

coast 4.6-5.2; 4.8 mean

Mink (liver) 1997 4 Aurora/Plainfield, IL 2.2-5.3; 3.7 mean Common merganser (liver) 1999 2 Buffalo Harbor, NY 3.7-4.2; 4 mean Greater and lesser scaup (liver) 1995-

1999 2 Niagara River, NY 1.9-2.7; 2.3 mean

Mallard (liver) 1995 1 North Tonawanda Creek,

NY 2.7

Atlantic salmon (skin on fillet) 2003 6 Farmed and wild, local
market NY <1-3.2 h

Smallmouth bass (liver) 2003 3 Effley Falls Reservo ir and Rock Pond, NY 4.3-5.4; 4.8 mean

Carp 2001 84 Lake Mead, NV 1.4-4.5; 3.0 mean Osemwengie
and Gerstenberger (2004)

* Note that human biomonitoring studies are summarize d in Appendix A, Section A

- 2 ND Not detected, method detection level = 12.5 μg/k g dw

a Measured June 2005.

b Measured September 2005.

c 31 days after application of biosolids.

d 156 days after application of biosolids.

e 30 days after application of liquid swine manure.

f 139 days after application of liquid swine manure.

g 38 percent of samples contained detectable concentr ations.

h 83 percent of samples contained detectable concentr ations.

### G-7 USGS National Water Quality Information System Data

Monitoring data collected by the USGS NWQL and avai lable from the NWIS database up to May
2012 were obtained for environmental concentrations of HHCB within the US (USGS, 2012). The following categories of HHCB data were availabl e from the USGS NWIS: water, filtered recoverable; water unfiltered, recoverable; and sol ids recoverable, dw. These data were grouped by medium type and site type, as shown in T able G-8. The Medium and Site Code definitions are provided in Table G-9.

The data compiled herein include values that are re ported as less than the USGS LRL; values that are between the LRL and the LT-MDL; and values that are below the LT-MDL (Oblinger
Childress **et al.** , 1999). Similar categories of data were available for both types of water samples. However, for unfiltered water from effluen t/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 p arameter codes ( **i.e.,** 62075, 62823, and 63209) for HHCB were included. The LRL for water sa mpling 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-M DL 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” w ere replaced by the LRL divided by the square root of two, as per the EPA OPPT guidance do cument (EPA, 1994). Likewise, where the geometric standard deviation was >3.0, values recor ded as “less than LRL” or “estimated” were replaced by the LRL divided by two (EPA, 1994). The se values are summarized in Table G-7 below. It should be noted, therefore, that resultin g 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 represe nt quantitative measured concentrations.

Table_Apx G

- 7

. Summary of Substituted Values (μg/L) for Water Samp les

7/2001 to 9/2009 (LRL=0.5 μg/L) 10/2009-2012

(LRL=0.05 μg/L)

Geometric Standard Deviation < 3 0.35 0.035

Geometric Standard Deviation < 3 0.25 0.025

USGS data from the NWIS database was accepted with the assumption that their internal methodologies were consistent and robust. These dat a were presumed to be collected under the guidance of the USGS National Field Manual for the Collection of Water-Quality Data, a publication which documents the methods, protocols, procedures and recommended practicesfor the collection of water-quality data (USGS, var iously dated). Data reporting procedures were

presumed to follow USGS guidance (Oblinger Childres s et al., 1999).

For each of the data groupings (Table G-8), box plo ts 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 pre sented in Table G-8.

Table_Apx G

- 8

. Summary of Box Plots for USGS HHCB Data

USGS Parameter Code Medium/Site HHCB Concentration

5 th

Percentile Q1 Median Mean Q3 95 th

Percentile n % of n

<LRL

62075 Water, filtered, recoverable; HHCB (μg/L)
Effluent/stream 0.35 0.35 1.40 1.18 1.82 2.16 10 40
Effluent/outfall 0.09 0.35 0.76 0.98 1.2 3.40 27 48
Surface water/outfall 0.35 0.35 0.65 1.08 1.70 2.30 41 48
Surface water/stream 0.10 0.35 0.35 0.35 0.35 0.35 2803 98
Surface water/lake, reservoir, impoundment 0.04 0.35 0.35 0.33 0.35 0.35 263 100

Groundwater/well 0.04 0.35 0.35 0.32 0.35 0.35 1548 99

62823 Water, unfiltered, recoverable; HHCB (μg/L)
Effluent/outfall 0.14 0.14 1.02 1.01 1.85 2.26 10 40
Surface water/stream 0.02 0.05 0.14 0.12 0.14 0.15 1568 100
Surface water/lake, reservoir, impoundment 0.03 0.14 0.14 0.12 0.14 0.14 45 100

63209 Solids, recoverable, dry weight; HHCB (μg/kg)
Bottom material/lake, reservoir, impoundment 27.93 41.02 65.42 87.46 106.08 212.87 122 99

Bottom material/stream 17.68 24.75 40.00 67.99 67.19 200.00 484 92

n: number of sample measurements; LRL: laboratory r eporting levelTable_Apx G

- 9

. USGS Medium and Site Codes

Medium Codes

Surface water Water on the surface of the earth stored or transported in rivers, streams,

estuaries, lakes, ponds, swamps, glaciers, or other aquatic areas. It also may refer to water in urban drains and storm-sewer systems.

Effluent Treated or untreated wastewater after use at a faci lity or wastewate r treatment

plant, or from combined sources, such as combined-s ewer overflows or tile drainage systems.

Groundwater Water below the surface of the earth contained in t he saturated zone. It does

not include soil moisture or interstitial water.

Bottom material A mixture of mineral and organic matter that compos e 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 s urface

- water body ( **e.g.** , the

point where wastewater is returned to a stream). T ypically, the discharge end of an effluent pipe.
Lake, reservoir, impoundment An inland body of standing fresh or saline water th at is generally too deep to

permit submerged aquatic vegetation to take root ac ross 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 containi ng 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 subsur face material. The diameter of a well is typically much smaller than the depth. Wells are also used to artificially recharge groundwater or to pressurize oil and gas production zones. Additional information about specific kinds of well s should be recorded under the secondary site types or the Use of Site field. Und erground waste-disposal wells should be classified as waste-injection wells.

Stream A body o f running water moving under gravity flow in a defi ned channel. The

channel may be entirely natural, or altered by engi neering practices through straightening, dredging, and (or) lining. An entir ely artificial channel should be qualified with the "canal" or "ditch" secondary sit e type.

Figure G-1 shows box plots for filtered water from three different mediums and associated sites. HHCB concentrations are highest from effluen t at outfall or stream sites and from surface water at outfall sites with higher values at outfal l sites, although the range of values varied from ND to >3 μg/L. The mean effluent concentration s were comparable, but the range of concentrations was greater at outfall sites. This m ay 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. M ean values for filtered water were <1.5 μg/L for all medium and site types.

Groundwater concentrations at well sites and surfac e water concentrations at streams and lakes, reservoirs, and impoundment sites ranged fro m ND to <0.5 μg/L. For effluent/outfallfiltered/non-filtered samples, the HHCB concentrati ons were higher in unfiltered samples

(Figure G-2). This may be attributed to the presenc e of HHCB in suspended solids. Again, significantly lower concentrations were consistentl y recorded in filtered surface water samples collected from streams and lakes, reservoirs, and i mpoundment sites—areas that are unaffected or nominally affected by wastewater trea tment outfall.

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

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

Unfiltered

Figure Apx G-3. Monitoring Data Summary from USGS NWIS for HHCB Concentrations in SolidsThe majority of the data available from the USGS NW IS data set for solids was from bottom

material. Bottom material data were predominantly d ivided between two types of sites: lake/reservoir/impoundment and streams (Figure G-3) . The 95 th percentile concentration range at lake/reservoir/impoundment sites was 213 μg/kg, and the mean concentration was 87 μg/kg. The 95 th percentile concentration at stream sites was 200 μ g/kg and the mean concentration was 68 μg/kg. The range of data was s imilar for both site types, with the Q3 value for lake/reservoir/impoundment sites slightly higher than that observed for stream sites, though it is unclear what this may be attributed to

.