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-hexamethylcyclopenta-γ-2-

benzopyran 

CASRN: 1222-05-5 

H3C

CH3

CH3

CH3

H3C

O

CH3

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

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

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

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 Species 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 HUMAN HAZARD CHARACTERIZATION ............................................................................................................... 78

A-1-1 Toxicokinetics (Absorption, Distribution, Metabolism, 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 HUMAN BIOMONITORING ............................................................................................................................... 84

A-3 SUMMARY OF 2008 EU HUMAN HEALTH RISK ASSESSMENT ................................................................................. 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 Environment ................................................................... 90

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

A-4 KEY SOURCES OF UNCERTAINTY AND DATA LIMITATIONS ON HUMAN HEALTH ........................................................... 91

A-5 CONCLUSIONS OF HUMAN HEALTH ASSESSMENT ................................................................................................. 92

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

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

ANALOGS (#7 TO #15) ...................................................................................... 95
Appendix D CDR DATA FOR HHCB ....................................................................................... 98

Appendix E MODELED RELEASE ESTIMATES ACCORDING TO STAGE OF PRODUCTION AND 

USE ............................................................................................................... 100

E-1 ESTIMATED RELEASE FROM MANUFACTURE AND IMPORT .................................................................................... 100

E-2 ESTIMATED RELEASE FROM COMPOUNDING ..................................................................................................... 104

E-3 ESTIMATED RELEASE FROM BLENDING OF FRAGRANCE OILS ................................................................................. 106

E-4 ESTIMATED RELEASE FROM USE OF COMMERCIAL AND CONSUMER PRODUCTS ........................................................ 109

Appendix F ADDITIONAL STUDIES .................................................................................... 110

F-1 ENDOCRINE MECHANISMS AND MOLECULAR PATHWAYS .................................................................................... 110

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

G-1 MEASURED CONCENTRATIONS IN WASTEWATER ............................................................................................... 111

G-2 MEASURED CONCENTRATIONS IN SURFACE WATER ............................................................................................ 115

G-3 MEASURED CONCENTRATIONS IN SEDIMENT ..................................................................................................... 118

G-4 MEASURED CONCENTRATIONS IN BIOSOLIDS AND SLUDGE ................................................................................... 120

G-5 MEASURED CONCENTRATIONS IN SOIL ............................................................................................................. 124

G-6 MEASURED CONCENTRATIONS IN BIOTA .......................................................................................................... 126

G-7 USGS NATIONAL WATER QUALITY INFORMATION SYSTEM DATA ......................................................................... 131
LIST OF TABLES 

Table 2-1. Physical-Chemical Properties of HHCBa

 ...................................................................... 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 Oils 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-Disappearance Times for Environmental 

Media .......................................................................................................................... 30

Table 2-7. BCFs and BAFs (L/kg ww) of HHCB in Aquatic Vertebratesa ....................................... 32

Table 2-8. BCFs (L/kg ww) of HHCB in Benthic and Terrestrial Invertebrates ............................. 33

Table 3-1. Measured Concentrations of HHCB in Biota ............................................................... 36

Table 3-2. Aquatic Toxicity Data for HHCB - Acute Toxicity ......................................................... 42

Table 3-3. Aquatic Toxicity Data for HHCB - Chronic 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 Environmental Toxicity .................................. 49

Table 3-7. Environmental Concentrations Used to Calculate RQs .............................................. 50

Table 3-8. Calculated Risk Quotients (RQs) for HHCB .................................................................. 51

LIST OF APPENDIX TABLES 

Table_Apx A-1. Summary of Human Health Hazard Information ............................................... 82

Table_Apx A-2. Human Biomonitoring Data for HHCB ................................................................ 85

Table_Apx C-1. HHCB, HHCB Diastereoisomers, and Related Structural Analogs ....................... 95

Table_Apx D-1. CDR National HHCB Informationa ....................................................................... 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 Facilities under NAICS 32562 Toilet Preparation 

Manufacturing ................................................................................................. 105

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

Detergent Manufacturing ................................................................................ 107

Table_Apx E-4. Geographic Distribution for Facilities under NAICS 325612 Polish and Other 

Sanitization Goods Manufacturing .................................................................. 108

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

Table_Apx G-4. Measured Concentrations of HHCB in Biosolids and Sludge ........................... 123

Table_Apx G-5. Measured Concentrations of HHCB in Soil ....................................................... 126
Table_Apx G-6. Measured Concentrations of HHCB in Biota*

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

Table_Apx G-7. Summary of Substituted Values (μg/L) for Water Samples ............................. 131

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 Environmental Protection Agency (US EPA), 

Office of Chemical Safety and Pollution Prevention (OCSPP), Office of Pollution Prevention and 

Toxics (OPPT). The Workplan Chemical Risk Assessment for 1,3,4,6,7,8-Hexahydro-4,6,6,7,8,8-

hexamethylcyclopenta-γ-2-benzopyran (HHCB) was prepared based on currently available data 

and any additional information received during the public comment period and peer review 

process. Mention of trade names does not constitute 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 & Sustainable 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 support from Abt Associates, Eastern Research 

Group, Versar and SRC, Inc. 

Acknowledgements 

The following individuals contributed to portions of the draft document: 

Jonghee Koh (visiting scientist, Korean Environmental 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 risk 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/riskassess.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 

E2 Estradiol 

EC European Commission 

ECHA European Chemicals Agency 

EPA Environmental Protection Agency 

EPCRA Emergency Planning and Community Right-to-Know 

EPI Estimation Programs Interface 

ESD Emission Scenario Document 

EU European Union 

EUSES European Union System for the Evaluation of Substances 

FIFRA Federal Insecticide, Fungicide and Rodenticide Act 

GC/MS Gas chromatography/mass spectrometry 

GD Gestation day 

GLP Good Laboratory Practice 

HERA Human and Environmental Risk Assessment Project 

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

HPLC High performance liquid chromatography 

HPV High production volume 

IFRA International Fragrance Association 

IPM Isopropyl myristate 

IUR Inventory Update Reporting Rule 

k1 Uptake rate constant 

k2 Elimination rate constant 

Kd Sorption coefficient 

kg Kilogram(s) 

KOC Organic carbon-normalized sorption coefficient 

KOW Octanol:water partition coefficient 

lbs Pounds 

LDR Larval development ratio 

LOAEC Lowest-observed-adverse-effect concentration

LOAEL Lowest-observed-adverse-effect level 

LOEC Lowest-observed-effect concentration 
Log KOC Logarithmic organic carbon partition coefficient 

Log KOW Logarithmic octanol:water partition coefficient 

LRL Laboratory reporting level 

LT-MDL Long-term method detection level 

lw Lipid weight 

MATC Maximum Acceptable Toxicant Concentration 

MOS Margin of safety 

NAICS North American Industry Classification System 

ND Not detected 

NHANES National Health and Nutrition Examination Survey 

NOAEC No-observed-adverse-effect concentration 

NOAEL No-observed-adverse-effect level 

NOEC No-observed-effect concentration 

NWIS National Water Information System 

NWQL National Water Quality Laboratory 

OCSPP Office of Chemical Safety and Pollution Prevention 

OECD Organisation for Economic Co-operation and Development 

OPPT Office of Pollution Prevention and Toxics 

OSPAR Oslo Paris 

PNEC Predicted no-effect concentration 

REACH Registration, Evaluation, Authorisation and Restriction of Chemicals 

RHOearthworm Bulk density of earthworm 

RAR Risk assessment report 

RQ Risk quotient 

SCAS Semi-continuous activated sludge 

SIAM SIDS Initial Assessment Meeting 

SIAR SIDS Initial Assessment Report 

SIDS Screening Information Data Set 

SSO Sanitary sewer overflow 

TG Test Guideline 

TLC Thin-layer chromatography 

TSCA Toxic Substances Control Act 

UF Uncertainty factor 

UFtotal Total uncertainty factor 

US United States 

USGS United States Geological Survey 

wt% Weight percent 

ww Wet weight 

WWTP Wastewater treatment plant 

yr(s) Year 
EXECUTIVE SUMMARY 

As a part of EPA’s comprehensive approach to enhance the Agency’s existing chemicals 

management, in March 2012 EPA identified a work plan 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 identifies 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, shampoos, lotions, detergents, fabric 

softeners, and household cleaners. HHCB is not produced in the US, but is imported. The 

volume of HHCB imported into the US has been reported 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 musks are not being replaced with other 

synthetic musks, as may be occurring in Europe. 

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

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

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

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

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

bioaccumulation into aquatic organisms is low to moderate, depending on the species in which 

it is measured. 

Problem formulation resulted in the assessment focusing on environmental risk to the aquatic 

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

products. Exposure pathways of interest included discharge of wastewater to surface water or 

application of biosolids to land. The assessment endpoints 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 European 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 information for HHCB. 

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

(USGS) National Water Information System (NWIS) database were used in the risk assessment. 

Levels of HHCB measured in wastewater, surface water, sediment, and soil were identified from 

these sources. More than 6800 data points for effluent, surface water and sediment were 

available for HHCB in the NWIS database. Soil monitoring data was limited to two published 

studies. 

 

1 http://www.epa.gov/oppt/existingchemicals/pubs/workplans.html
Ecotoxicity studies for HHCB have been conducted in fish, aquatic invertebrates, aquatic plants, 

sediment invertebrates, soil invertebrates, and terrestrial plants. Concentrations of concern 

(COCs) were derived from these studies according to established EPA/OPPT methods (EPA 

2012f; 2013). This assessment considered only direct exposure to aquatic and sedimentdwelling organisms via contact with contaminated water or sediment. Based on half-lives in 

water, sediment and continuous wastewater inputs, exposure to organisms was considered 

continuous. 

Potential risks were calculated using the risk quotient (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 concern (the COC). A RQ greater than one 

indicates there may be risk of concern. To incorporate the range of measured environmental 

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

measured environmental concentrations for surface water and sediment. Available soil 

concentrations were inadequate for calculating an RQ for soil organisms. RQs were calculated 

for aquatic and sediment dwelling organisms for acute 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 chronic COC for aquatic organisms (i.e., RQs 

ranged from 0.006 to 0.04 and 0.04 to 0.23, respectively). The chronic COC for sedimentdwelling organisms was also not exceeded at the upper range of measured environmental 

concentrations for the maximum published values or the more than 600 sediment 

measurements from the USGS. RQs for chronic risk to sediment dwelling organisms ranged 

from 0.001 to 0.36. RQs were not calculated for terrestrial invertebrates and plants via contact 

with contaminated water, sediment, or soil. 

In conclusion, under the exposure scenarios assessed in this this risk assessment, current 

environmental exposure concentrations are one to two 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 enhance the Agency’s existing chemicals 

management, in March 2012 EPA identified a work plan 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 for 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 reproductive 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 of chemicals as candidates for risk 

assessment in the next several years. In the prioritization 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 chemicals in the work plan. If an assessment 

identifies unacceptable risks to humans or the environment, EPA will pursue risk management. 

The target audience for this risk assessment is primarily 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 in 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 development process, which is distinct from the initial 

prioritization exercise, includes scoping and problem formulation. During these steps EPA 

reviews currently available data and information, including but not limited to, assessments 

conducted by others (e.g., authorities in other countries), published or readily available reports, 

and published scientific literature. During scoping and problem formulation the more robust 

review of the factors influencing initial prioritization may result in refinement – either 

addition/expansion or removal/contraction – of specific 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 perfumes, cosmetics, shampoos, lotions, 

detergents, fabric softeners, air fresheners, and cleaning agents. 

 

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

3 http://www.epa.gov/oppt/existingchemicals/pubs/wpmethods.pdf 
Previous assessments in Europe, Sweden, and Australia 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: “There is at present no need for further 

information and/or testing and no need for risk reduction measures beyond those which are 

being applied already.” No potential risk concerns were identified for surface water and 

sediment in Sweden (SWECO, 2008) or for the marine environment (OSPAR, 2004). Potential 

risk concerns were identified for HHCB in certain highly contaminated sediments in Europe 

(HERA, 2004) and in an initial screening study of pharmaceuticals, personal care products, and 

endocrine disrupting compounds in surface runoff from 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 inland waterways (Anderson et al., 

2012). Further, California lists HHCB as a designated chemical for biomonitoring, however, 

California is not yet biomonitoring for these chemicals (CA EPA OEHHA, 2014b). In addition, the 

Oregon Department of Environmental Quality lists HHCB 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 regulatory activity in the US, Canada, or the European 

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

substance, and is reported for Chemical Data Reporting (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. HHCB is approved as an inert ingredient for 

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

Rodenticide Act (FIFRA).4

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

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

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

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

 

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 beginning with planning and scoping and 

problem formulation. 

EPA/OPPT approached problem formulation by reviewing 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. The literature review included HHCB 

chemistry, uses, sources including industrial releases, fate, exposure, and hazard to humans and 

ecological receptors. Other sources of information reviewed were unpublished reports 

provided by the Research Institute for Fragrance Materials (RIFM), the publicly available EPA 

Inventory Update Reporting (IUR) and Chemical Data Reporting (CDR) databases, and the US 

Geological Survey National Water Information Service (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 summarized 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. Three additional companies that 

reported to CDR are also assumed to import. 

• TSCA uses for HHCB are as an ingredient in detergents, fabric softeners, dishwashing 

detergents, and commercial and consumer general purpose 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 consumer products, HHCB is 

continuously released to the environment in municipal wastewater and sludge. 

Releases to the environment from industrial sites are also expected. 

• HHCB is not readily biodegradable and is expected partition to organic-rich particles in 

sediment, soil and sludge, based on the physical-chemical properties. 

• HHCB is moderately persistent in the environment and bioconcentration has been 

measured as low to moderate, depending on the species tested. HHCB has been found 

to have low bioaccumulation potential and to be biotransformed to more polar, less 

toxic, metabolites in some organisms. 

• HHCB is highly toxic to aquatic organisms under both acute and chronic exposure 

conditions. Chronic effects of HHCB in aquatic organisms may include effects on adult 

and embryo survival, behavior, respiration, and larval development. Mechanistic 

studies (described in Appendix F) show that HHCB may interfere with steroid hormone 
receptors, 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 US Geological Survey (USGS) National 

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

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

organizations. Appendix A summarizes the conclusions of the 2008 EU RAR. 

• Human dermal and inhalation exposures occur intentionally as a result of use of 

cosmetics and personal care products and may also occur during use of detergents and 

cleaners. HHCB is frequently measured in human blood 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 toxicity 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 approximately 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, including a risk assessment performed 

by the European Union, EPA/OPPT determined that further 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 widely used polycyclic musk fragrance 

ingredients in a range of consumer products including perfumes, cosmetics, shampoos, lotions, 

detergents, fabric softeners and household cleaners. Because exposure to consumers occurs 

intentionally for a majority of these uses, EPA/OPPT 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 products 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 production/import volume (see Appendix E: 

Release Estimates). Due to the ubiquity of fragranced products, consumer uses are a continuing 

source of HHCB to wastewater. When HHCB enters wastewater treatment plants (WWTPs), the 

chemical is negligibly volatile and expected to partition to organic-rich particles during 

wastewater treatment based on the physical-chemical properties. However, HHCB removal 
during treatment in WWTPs is suboptimal (see Chapter 2, Fate in Wastewater) and HHCB is 

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

During the wastewater treatment process, influent is 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 enters 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 billions of gallons of untreated wastewater directly 

to surface water. When HHCB has not been removed during the treatment process and is 

present in effluent, sorption to suspended solids in 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 is 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. Industrial releases are <10% of releases and are not indicated in 

the diagram. 

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

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

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 become concentrated in biosolids. This 

organic carbon-rich material is disposed of by landfill 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 to 
agricultural 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 available 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 invertebrates and plants was not evaluated due to 

insufficient data for HHCB concentrations in soil. 

EPA/OPPT reviewed other recently performed assessments and searched published literature 

from 2005, the date of the last literature review indicated 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 provided by the Research Institute for 

Fragrance Materials (RIFM), the publicly available EPA Inventory Update Reporting and 

Chemical Data Reporting databases, and the US Geological Survey National Water Information 

Service (NWIS). 

Key ecological hazard reports were obtained for critical 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 ecotoxicological 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 risk quotient (RQ) for each route of 

environmental exposure (i.e., water and sediment) based on measured media concentrations. 

 

 

5

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

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

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

surface disposal units, and or incineration facilities. (NEBRA, 2007). 
2 SOURCES AND ENVIRONMENTAL FATE 

2.1 INTRODUCTION 

As an initial step in assessing the environmental risks 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 musk-like odor. The polycyclic chemical 

structure for HHCB is shown in Figure 2-1. 

H3C

CH3

CH3

CH3

H3C

O

CH3

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

Synonyms for HHCB include: 

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

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

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

• Galaxolide 

• Galaxolide 50 

• Galaxolide 50BB 

• Galaxolide 50IPM 

• Galaxolide White 

• Abbalide 

• Pearlide 

Technical HHCB consists of a mixture of isomers that 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 isomers (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 shown in Table 2-1. 

Table 2-1. Physical-Chemical Properties of HHCBa

Molecular formula C18H 26O

Molecular weight 258.44 g∙mol-1

Physical form Colorless liquid; highly viscous liquid at 20 °C and 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 160 °C at 4 hPa) 

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

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 105) 

Flash point 144 °C 

Henry’s Law constant 

1.13 x 10-4 atm∙m3

/mol (at 25 °C, calculated using 

measured vapor pressure and water solubility); 

1.32 x 10-4 atm∙m3

/mol (at 25 °C, estimated using 

HENRYWIN program in EPI Suite v4.11); 

3.65 x 10-4 atm∙m3

/mol (at 25 °C, measured)d 

a Source: (HSDB, 2007) except as noted.

b OECD – Organization for Economic Cooperation and Development 

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 and uses specifically for HHCB and more 

generally for fragrances. The manufacturing process for HHCB is described in Appendix B. The 
HHCB 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 Volume chemical that is widely used in 

cleaning and personal care products. Data on production volume and uses are amenable for 

determining release estimates to the environment (See Appendix E). These data are also useful 

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

environment through industrial and consumer applications. 

2.2.1 Production 

HHCB is one of the most widely used and consumed polycyclic 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 information (CBI) CDR data for HHCB in 2012 

(EPA, 2014)7 and a sixth reported only CBI information7. 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, two did not indicate whether they are a 

manufacturer or importer. EPA/OPPT assumes that these companies and the sixth company 

that reported only CBI information are also importers 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 Companya 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 datab

Firmenich, Inc. Yes No datab

aOne company reported as a producer of HHCB in 2012, however all data were CBI so the company is not included 

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 synthetic musk fragrances increased by 

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

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

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

 

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

7

 These six producers may be an underestimate because a production site is only required to submit a Form U (the 

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

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

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

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

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

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

was added to the authorization list in February 2011, 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 International 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 Practice. The Code of Practice is mandatory 

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

volume of fragrance materials (IFRA, 2011). 

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 yr, respectively (IFRA, 2012b).8

 According to 

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

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

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

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

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

cited in Peck et al., 2006). At present, it appears as though the US production volume (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, with the exception of 2002, the non-CBI IUR 

production volume for musk xylene has been consistently 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 has 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 indicated that the 2008 use volume of HHCB in the US was 

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

tons per year in 2008. 
Table 2-3. US Production/Import Volume of HHCB

Chemical 

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

Year 

1986 1990 1994 1998 2002 2006 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 fragrance industry because of their unique 

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

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

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

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

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

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

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

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

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

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

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

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

free” (Potera, 2007). 

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

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

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

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

300 ingredients (HERA, 2004a; Bickers et al., 2003). As shown in Table 2-4, 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 weight percent (wt%) basis, perfume 

extracts have the highest level of fragrance incorporation at 20 wt% (Table 2-5). Given the 

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

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

used in detergents, fabric softeners, and personal care products. 
Table 2-4. Estimated Distribution of Fragrance Oils by Use

Use Fragrance Oils (percent) 

Detergent 25 

Fabric softeners 14 

Personal care 13 

Bath and shower 10 

Hair care 10 

Soaps 9 

Industrial and household cleaner 8 

Other 6 

Fine fragrances 5 

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

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 application of different cosmetics (COLIPA, 1987; 

as cited in Cadby et al., 2002). 

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

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

Source: Cadby et al., 2002 

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

One study found HHCB in 55% of the personal care products tested, including 100% of 
deodorants tested (Roosens et. al., 2007). A 2002 survey found that HHCB may be found at 

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

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

concentrations of HHCB and a degradant of HHCB, HHCB-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. However, these studies indicate that levels 

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

the complexity of fragrance mixtures which often contain hundreds of constituents. 

2.2.3 Conclusions of Production and Use 

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

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

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

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

laundry detergent). 

2.3 ENVIRONMENTAL FATE 

HHCB enters domestic wastewater treatment and is subjected to removal by a variety of 

processes. HHCB and any degradation products generated in wastewater treatment are 

mainly discharged to receiving waters or disposed of in biosolids. This section summarizes 

current knowledge of the fate and degradation of HHCB 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 fate 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 concentrations 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). HHCB disappearance with subsequent 

appearance of more polar entities was observed (Langworthy 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; see 
guidance in the Estimation Programs Interface (EPI) Suite v4.119

 (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., 2007). 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) and 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 phases based on its high measured log KOW 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 KOC (log 

basis) of 4.1-4.3 (KOCWINTM program v2.00; in EPI SuiteTM v4.119, (EPA, 2011)). Values for both 

Kd (sorption coefficient) and KOC (organic carbon-normalized sorption coefficient) are generally 

in the range of 3 to 4 on a logarithmic scale. This means that HHCB will be substantially 

removed by sorption to sludge in WWTPs; will have low mobility in soil; and will bind strongly to 

benthic and suspended sediment. The Human and Environmental Risk Assessment Project 

(HERA, 2004a) summarized both theoretical and measured percent removal values (up to 2004) 

for European and US activated sludge-based treatment systems. Observed removal was 

generally >50 percent (usually >80 percent), and removal percentage correlated well with 

removal of total suspended solids (Simonich et al., 2002). Kupper et al. (2006) reported overall 

HHCB removal of 81% for a full-scale Swiss treatment plant, in line with the earlier results; and 

Horii et al. (2007) reported similar overall HHCB removal for wastewater treatment plants in 

Kentucky and Georgia. 

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

sludge volume and organic load before disposal of the concentrated material. Removal of HHCB 

under anaerobic treatment conditions has been addressed in several studies. Xue et al. (2010) 

reported poor overall removal of HHCB in a full-scale anaerobic/anoxic/aerobic system 

combined with membrane bioreactor for municipal wastewater reclamation, and little removal 

was attributed to biodegradation. In contrast Carballa et al. (2007) reported good removal of 

HHCB (60-80%) in mesophilic and thermophilic anaerobic 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. Therefore, 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 environment; therefore to benthic and 

suspended sediment. Though HHCB is negligibly volatile in WWTP, volatilization is an 

important process for HHCB in water. After 12.5 days, about 40 percent of the radioactivity 

was lost by volatilization. In a more recent presentation reporting on the same or extended 

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

 

9

 Available for download from http://www.epa.gov/opptintr/exposure/pubs/episuite.htm and 

http://www.epa.gov/oppt/exposure/pubs/halflife.htm
parent 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 disappearance of HHCB in river water (Schaefer, 2005; 

as reported in EC, 2008). A significant fraction (up to 16 percent) was volatilized after 28 

days. These studies show that specific test conditions impact observed volatilization 

behavior, but that in any case volatilization is significant 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 were designed to simulate river water 

conditions immediately below discharge of treated sewage 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 that 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 radiolabeled HHCB at concentrations of 1.0 or 

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

starting concentration. Following 29 days, polar metabolites 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 metabolites were formed at different rates and with 

increasing polarity over time (Schaefer et al., 2005; as reported in EC, 2008). In the river dieaway test the total radioactivity of metabolites rose from 25 to 56 percent between days 2 and 

7 and to nearly 62 percent at 28 days. When the results were corrected for the amount 

volatilized and the non-recovered fraction as well as for abiotic transformation, the primary 

biodegradation extent on day 28 was approximately 60 percent. About 40 percent of the 

radioactivity was lost by volatilization at 300 hours, but some of the volatilized radioactivity 

could have been in the form of degradation products, since degradation in water was also 

happening during the test. 

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 (sludge amended soil) (Envirogen, 1998; as 

cited in EC, 2008). In the study, sealed flasks with HHCB-spiked soil at 10 μg/g soil (10 mg/kg) 

were incubated at laboratory ambient temperature for one year. Periodically the headspace 

was flushed for oxygen replenishment and volatile organics and carbon dioxide were captured. 
After incubation, and exhaustive extraction with ethyl acetate and/or acetone/hexane, an 

aliquot of the solvent was used for TLC. After one year, significant amounts of polar metabolites 

were found, and HHCB was present at 4, 7, 9 and 35 percent of the initial concentration in the 

river sediment, forest soil, sludge-amended soil, and agricultural soil, respectively. 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 degraded to various more polar metabolites. It 

was hypothesized that the majority of unrecovered radiolabel became covalently bound to soil 

organic compounds (i.e., was immobilized by humification). 

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

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

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

and highly weathered oxide-rich soil. The anaerobically digested and dewatered sludge was 

obtained from activated sludge WWTPs in Georgetown, Delaware (100 percent domestic 

sewage; 10 percent solids) and Wilmington, Delaware (70 percent domestic sewage; 17 percent 

solids). The unamended concentration of HHCB in the digested sludge from Georgetown, 

Delaware was 86 and 38 mg/kg dw in the years 2000 and 2002, respectively and from 

Wilmington was 43 and 22 mg/kg dw, respectively. The initial HHCB concentrations in spiked 

soil were 6 and 13 mg/kg soil, whereas the levels in unspiked soil were 0.1 to 0.27 mg/kg. In all 

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

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

from 8 to 60 percent of the initial concentration. During the three months when the soil was 

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

to 14 percent of the initial concentration. The rate of dissipation was higher in the soils with 

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

leaching as well as biotransformation. The EC (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 study, 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 from unspiked soils (EC, 2008). The low 

percentage of HHCB initially in the sludge and present 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 that HHCB disappeared almost completely 

from soil within one year. The half-life based on unfrozen 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 initial concentrations. In the EU RAR (EC, 2008), 

a half-life of 105 days in the biosolids-amended soil was deemed most relevant for modeling 

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

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

2008). EPA/OPPT agrees that these values are reasonable for modeling and assessment 

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

Degradation Half-Life or Half-Disappearance Timea

Reference 

Water Sediment Soil Sludge BiosolidsAmended 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 hoursb

(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)

aDT50 is the time required for disappearance of 50% of the starting material. Half-disappearance time (DT50) 

differs from degradation half-life in that it includes all mechanisms contributing to disappearance (e.g., 

volatilization in addition to degradation). Numbers that are DT50 values are indicated in the table; otherwise 

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 (photooxidation) of HHCB and 

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

second-1. This leads to a calculated atmospheric half-life of 3.4 hours for HHCB (based on a 

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

an important process for HHCB. Recent work of Villa et al. (2014) agrees, but also suggests that 

“moderate-range” (this could be called regional) atmospheric transport may occur. Any HHCB 

in the atmosphere is likely to be in the gas (not particulate) phase. This is supported by 

consistent results from three different models that predict extent of sorption to atmospheric 

particulates (aerosols) based on chemical properties. The fraction of HHCB sorbed to airborne 

particulates from AEROWIN v1.00 (EPI Suite v4.11) were 0.00149, 0.00288 and 0.00329 for the 

Junge-Pankow, octanol/air (Koa) 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 information on the bioconcentration and 

bioaccumulation potential of HHCB. BCFs and BAFs varied with species but were generally 

lower than the commonly used criterion of 1,000 (log = 3) for 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 bluegills, 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 Test Guideline 305 and using Good 

Laboratory Practice (GLP) standards (EC, 2008; Rimkus, 1999). Bluegill sunfish (Lepomis 

macrochirus) were exposed in a flow-through system to two concentrations of radiolabeled 

HHCB. A solubilizer was used to prepare a solution (concentration 0.001 percent (w/v)). 

Identification of the parent compound in water and fish was performed using TLC/HPLC. 

Nominal exposure concentrations were 1 and 10 μg/L. The fish were exposed for 28 days; the 

elimination period was also 28 days. The elimination rate constant (k2) was estimated from the 

elimination curve (first-order kinetics), and determination of an uptake rate constant (k1) was 

attempted using the following formula: k1 = BCF × k2. However, this value could not be directly 

calculated from the increase of concentrations in fish 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 concentration in fish. Based on the concentration 

of parent material, the BCF for whole fish was calculated to be 1,584 L/kg wet wt. 

Reiner and Kannan (2011) and Nakata et al. (2007) studied bioaccumulation of HHCB in 

organisms from the upper Hudson River, USA; and marine tidal and shallow-water organisms 

from a Japanese site, respectively. In the first study, 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 occurs when a substance achieves higher 

concentrations on a lipid-adjusted basis in organisms than in their food (prey). Aquatic food 

chain modeling was conducted by the EPA using the BCFBAF© program (which uses the ArnotGobas model) in EPI Suite© v4.11. Consistent with the monitoring data, estimated BAFs for 

middle and upper trophic fish were successively lower, not higher, than the estimated BAF for 

HHCB at the lowest trophic level. 

Metabolism may account for the observation that measured BCFs and BAFs are lower than 

would be estimated based on the log KOW of HHCB (5.9). In this study, HHCB was metabolized to 

one or more polar metabolites in fish in a relatively short time and then excreted to the water 

phase. Recent work of Fernandes et al. (2013) provides 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 Vertebrates

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 bassb 31-106

Largemouth bassb 30-146

White perchb 21-333

Catfishb 18-371

a

source: Dietrich and Hitzfeld (2004) unless otherwise noted 

b

source: Reiner and Kannan (2011) 

Invertebrates 

Bioaccumulation has been studied with two benthic organisms (Table 2-8). Non-biting midges 

(Chironomus riparius) exposed to 9.8 ± 1.4 μg/L, the concentrations in the organisms increased 

to a maximum level between days 1 and 3, and then the level rapidly decreased to a new 

steady state with BCF values ranging from 85 to 138 L/kg wet wt. (Artola-Garciano et al., 2003). 

With the addition of piperonyl butoxide (PBO), a cytochrome P450 inhibitor, the BCF was 525, 

and it was concluded that the lower BCF values were likely caused by biotransformation of 

HHCB in this organism. In a flow-through experiment 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 calculated 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 concentration and a BCF can be calculated using an 

equation given by Jager (1998) (as cited in EC, 2008): 

BCFworm = (0.84 + 0.012 × Kow)/RHOearthworm

Where: 

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

KOW = the n-octanol:water partition coefficient; and 

RHOearthworm = the density of the earthworm set at 1 kg wet wt/L. 
Table 2-8. BCFs (L/kg ww) of HHCB in Benthic and Terrestrial Invertebrates

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 2,395a Jager (1998) 

aCalculated

Plants 

Application of biosolids or reclaimed wastewater to agricultural land or crops may provide a 

pathway for uptake of HHCB by plants. Land application of reclaimed wastewater seems not to 

have been studied as a pathway for HHCB introduction into the environment, but as reported in 

the EU 2008 assessment, in an unpublished study (Müller 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 unrealistically high concentration of HHCB in biosolids, no 

relevant accumulation in leaves was observed and that uptake ratios (the ratio Cplant/Csoil) 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 concentrations in the above ground parts of the 

carrot plant showed that there was no transport within 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 solid phases. Therefore, HHCB is removed in 

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

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

persistent in soil and sediment. 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. However, BAF values are available 

for several aquatic organisms are in the range of 20 to 620, indicating low bioaccumulation. 

These studies, together with results of aquatic food-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 consumer and commercial products. The 

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

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

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

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

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 biota were used to characterize environmental 

exposure to HHCB. The data for concentrations of HHCB 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 throughout the life cycle of HHCB. In the US, the 

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

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

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

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

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

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

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

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

(see Appendix E). EPA/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 commercial and consumer use of products 

containing HHCB. 

3.2.2 Measured Levels in the Environment 

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

environment were reviewed for this assessment. Data quality criteria included currency, 

geographic scope, accuracy/reliability, representativeness, 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 published 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 contained in three USGS NWIS parameter 

codes (62075, 62823, 63209) for HHCB were included in this assessment. This database includes 

monitoring data as well as data collected for studies of organic anthropogenic waste indicators, 

organic wastewater contaminants, and biologically-active chemical mixtures in urban streams. 

USGS data from the NWIS database was accepted with the assumption that their internal 

methodologies were consistent and robust. These data 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, variously dated). Data reporting procedures were 

presumed to follow USGS guidance (Oblinger Childress 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 magnitude. 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 characterizations. Although sampling techniques 

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

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

of HHCB across all media. 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 Biota 

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

Effluent >135 0.010 13.0 0.02 – 5.31 (2004)

Effluent discharged to ocean n/a NDe 2.5 ---

d

Anderson et al. (2012)

 Effluent (at stream or outfall) 47 0.090a 3.4b 0.98 – 1.18 USGS (code 

62075 and 62823)

Surface Water (μg/L) 

Freshwater 94 NDf 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 0.02a 2.3b 0.12 – 1.08

USGS NWIS 

(code 62075 and 62823)

Groundwater (well) 1548 0.04a 0.35b 0.32 USGS NWIS (code 62075)

Sediment (μg/kg dw)

Freshwater >59 NDe 388 1.1 - 144c

USGS (2008), Klecka (2010), 

Sapozhnikova (2010), Chase 

(2012), Reiner et al. (2011), Peck

et al. (2006)

Bottom material (Lake, 

Resevoir, Impoundment, 

Stream) 

606 17.68a 212.87b 67.99-87.46

USGS NWIS (code 63209)

Biosolids/Sludge (μg/kg dw) >58 13 427,000 13 - 177,000c 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),

Wastewater treated n/a <0.33 5.69 --- Chase et al. (2012) 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) (μg/kg ww)

Earthworm (tissue conc.) 9 NDg 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 

Mink (wet weight) 4 2.2 5.3 3.7 Kannan (2005)

Bird (wet weight) 5 1.9 4.2 2.3 - 4.0 

ND not detected 

n/a not available 

a

 5th percentile value 

b

 95th 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 dw 
3.2.2.1 Wastewater 

In published studies, measured concentrations of HHCB 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 type 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 polycyclic musk concentrations more rapidly than 

anaerobically digested sludge (Smyth et al., 2008). Reported values from a Canadian study 

confirmed that effluent values were process dependent and concentrations were seasonally 

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

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, with 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 data points) were 0.98-1.18 μg/L. The 

highest values were measured at outfall sites (95th percentile: 3.4 μg/L). The mean effluent 

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

was greater at outfall sites. 

Very few studies have reported measuring metabolites of HHCB other than HHCB-lactone, the 

apparent principal degradation product of HHCB. However, measured values of HHCB-lactone 

in the influent (maximum 1.15 μg/L) and effluent (maximum 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 those 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 reported concentrations were ND levels (reported 

detection range = 0.038-2.16 μg/L) at locations upstream from WWTPs; higher concentrations 

of HHCB were found downstream from the WWTPs. A maximum 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 significant role in the observed surface water 

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

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

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

percentile groundwater concentrations at well sites and surface water concentrations at 

streams and lakes, reservoirs, and impoundment sites were ≤0.35 μg/L (Table 3-1; see also: 

Appendix G, Table G-3 and Figures G-2 and G-3). Lower concentrations were consistently 

recorded in surface water samples collected from streams 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 median 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 site. 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 upstream of the same WWTPs (average = 37 

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

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

Reported sediment concentrations in the EU (EC, 2008) and Asia (Lee et al., 2014) were similar 

to those in the US, with higher values found in Berlin, Germany and in the sediment of 

contaminated brooks in Hessen, Germany. 

Over 600 bottom material measurements (collected from 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 

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

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

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

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

3.2.2.4 Biosolids and Soil 

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

3-1; see also: Appendix G, Table G-4). These values were dependent on a number of factors, 

including: season, date of collection, location, population served, WWTP operations (e.g., 

municipal or industrial receiving waste stream, water volume, treatment type) and preparation 

methodologies prior to land application. Because these factors cannot be readily separated 
from all values within the reported data set, it was not possible to determine which single 

treatment variable had the largest impact on the reported environmental concentrations. 

However, EPA observed that the US values for measurements of biosolids were similar to those 

obtained in Ontario, Canada and likewise showed differences 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 amended soils (Table 3-2; see also: Appendix 

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

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

3.2.2.5 Biota 

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

reporting (e.g., lipid weight, tissue weight, or wet weight) (Table 3-1; see also: Appendix G, 

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

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

measurements). On a lipid weight basis, measured values across various fish species and 

locations ranged from <1 to 51.1 μg/kg, whereas levels in wild caught and farm raised shrimp 

were 330 and 424 μg/kg, respectively (Appendix G, Table G-6). Mean tissue weight 

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

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

and trophic levels. The reported values were relatively 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 USGS dataset includes values that are reported 

as less than the USGS laboratory reporting level10 (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 reported with a “less than” remark code for 

samples in which the analyte was not detected. “Estimated” remark codes are noted for all 

values falling outside the calibration range because of increased measurement uncertainty or 

values below the LT-MDL determined using information-rich methods11

. 

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

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

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

 

10 The LRL for water sampling was 0.5 μg/L for sampling dated 7/16/2001 – 9/30/2009 and was updated to 0.05 

μg/L for samples dating from 10/1/2009 to the present. 

11 The USGS defines information rich methods as “Classified as organic methods that use either mass spectrometric 

or photodiode array ultraviolet/visible spectroscopic detection. These methods have qualifying information that 

allows enhanced analyte identification.” 
was greater than 3.0, EPA/OPPT replaced values recorded as “less than LRL” or “estimated” 

with the LRL divided by two (EPA, 1994). This practice presents a conservative low end value 

protecting against false negative values; therefore, it should be noted that these values do not 

necessarily represent quantitative measured concentrations and are biased towards the LRL. 

High quality monitoring data (greater than 6800 data points) were available from the USGS 

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

effluent to a lesser degree (47 data points) as shown in Table 3-1. 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 exposure 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 environment. However, these data reflect 

discrete locations and times, and the extent to which they are representative of the overall 

distribution of HHCB is not known. 

3.3 ECOLOGICAL HAZARD ASSESSMENT 

The environmental hazard assessment is based on previous hazard assessments including Balk 

and Ford (1999a, b); EU RAR (EC, 2008); HERA Project Report (HERA, 2004a); and Robust 

Summaries submitted under the US EPA HPV program (IFRA, 2003). In addition, a literature 

search was performed to identify peer-reviewed articles on ecotoxicity published between 

2007 and May 2012. The search terms included freshwater and saltwater fish, aquatic 

invertebrates, and aquatic plants; pelagic and benthic organisms; acute and chronic sediment 

toxicity in freshwater and saltwater and terrestrial toxicity to soil organisms, birds, and 

mammals. The test species, test conditions, toxicity endpoints, statistical significance, and 

strengths/limitations of the study were summarized and evaluated for data quality. Data quality 

inclusion criteria included: use of appropriate analytical and test controls, identification of test 

substance and test organism, stated exposure duration time and administration route, a clear 

description of the effect endpoints, and transparent reporting of effect concentrations. 

Guideline studies as well as studies using other protocols were included if they met data quality 

criteria. Specific criteria for exclusion are: studies that included HHCB as part of a mixture in 

wastewater effluent or surface water and studies described only in abstract form or in a 

language other than English. 

Application of uncertainty factors based on established EPA/OPPT methods (EPA, 2012f; 2013) 

were used to calculate lower bound effect levels (referred to as the concentration of concern; 

COC) that would likely encompass more sensitive species not specifically represented by the 

available experimental data. Uncertainty factors are included in the COC calculation to account 

for differences in inter- and intraspecies variability, as well as laboratory-to-field variability. 
These uncertainty factors are dependent upon the availability 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 conducted under TSCA, since the data available 

for most industrial chemicals is limited (Ahlers et al., 2008). 

A summary of the available ecotoxicity data for HHCB that were deemed adequate for 

consideration in this assessment are provided in tables and the studies selected for use in 

calculating risk quotients are described in more detail 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 embryos, an EC50s of 0.39 mg/L and an LC50 of 

>0.67, were reported, respectively. These data were not included in EPA’s consideration 

because they were reported in a secondary source and sufficient study details were not 

provided in the report or found during the search for information performed by EPA/OPPT. The 

zebrafish study appears to be described in the EU RAR (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 from the available acute toxicity studies to 

calculate an RQ because Daphnia were the most sensitive species for acute toxic effects to 

HHCB. Although the lower end of the effects concentration range for freshwater mussel is 

lower than the Daphnid effects concentration, the authors 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 three times higher than the Daphnid effects 

concentration. 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. Daphnid is a 

more representative species for this assessment because the available monitoring data used for 

estimating exposure is largely freshwater. Furthermore, the study is reliable and demonstrates 

the acute effects (i.e. survival/immobilization) using appropriate, reproducible protocols. 

The Daphnia magna 48-hour EC50 of 0.282 mg/L was divided by an assessment factor 

(uncertainty factor [UF]) of 5 for invertebrates, as per established EPA/OPPT methods (EPA, 

2012f; 2013), to give an acute concentration of concern (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)

EndpointConcentration

(mg/L) 

Chemical 

Analysis Effect References

Fish - Freshwater

Bluegill sunfish 

Lepomis macrochirus Fresh

OECD TG 

204/Flowthrough 96-hra

 LC50 0.452 Measured Mortality Wüthrich 

(1996)c 

Aquatic Invertebrates - Freshwater

Water flea (Daphnia 

magna)

Fresh

OECD TG 

202/Semistatic 48-hra EC50 0.282 Measured Immobilization Wüthrich 

(1996)c

Freshwater mussel 

(Lampsilis cardium)

Fresh Staticb

48-hr 

glochidia LC50 0.999 - >1.75

Measured

Mortality 

Gooding et al. 

(2006) 96-hr 

juvenile EC50 0.153 - 0.831 Growth 

Aquatic Invertebrates - Saltwater

Estuarine Copepods 

(Nitocra spinipes)

Brackish SIS 1991/

Static 96-hr LC50 1.9 Measured Mortality Breitholtz 

et al. (2003) 

Marine copepod 

(Acartia tonsa)

Salt ISO 1997/

Static 48-hr LC50 0.47 Nominal Mortality Wollenberger 

et al. (2003) 

Aquatic Plants - Freshwater

Green algae 

(Pseudokirchneriella 

subcapitata)

Fresh OECD TG 

201/Static 72-hr EC50

>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 organisms.

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 acceptable quality are summarized in Table 3-3. 

Chronic toxicity values for rainbow trout (Oncorhynchus mykiss) and zebrafish (Danio rerio) with 

reported 21-day EC50-repro and 21-day LC50 of 0.282 mg/L and 0.452, respectively are reported. 

However, these data were not included in EPA’s consideration 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 information 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 organic matter and surfaces may affect 

exposure concentrations. For example, the measured concentrations at the end of two chronic 
copepod tests (Breitholtz et al., 2003; Wollenberger et al., 2003) were only 2 to 19 percent of 

the nominal concentration. Volatilization or sorption to organic material likely accounts for the 

disappearance of HHCB from the water phase. HHCB has been reported to have a half-life of 

hours to days in river water tests using radiolabeled 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 organic 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 otherwise 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 summarized 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 calculated 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 reconsidered this selection. The marine 

copepod (Acartia tonsa) study by Wollenberger et al. (2003) was considered 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 study and due to other test design 

limitations described in the EU RAR (EC, 2008). Although 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 used for estimating exposures is largely 

freshwater. Furthermore, the study is reliable and demonstrates the chronic effects (i.e., 

survival and growth) using appropriate, reproducible protocols. 

The study was conducted using a flow-through system, wherein fathead minnow eggs (<24 

hours old) were exposed to nominal concentrations ranging from 0.0125 to 0.2 mg/L for 36 

days. Corresponding mean measured concentrations (measured via HPLC or GC 13 times during 

the test) were: 0.0091, 0.019, 0.037, 0.068 and 0.140 mg/L. Egg hatchability was not affected at 

any concentration. HHCB did have an effect on larval 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, underdeveloped and displaying erratic swimming 

behaviors. The authors identified 0.068 mg/L as the NOEC for the study (based on survival and 

growth); the LOEC would be 0.140 mg/L. 

EPA/OPPT calculated at a Maximum Acceptable Toxicant Concentration (MATC) effect 

concentration of 0 0.097 mg/L for survival. To derive a chronic COC, the MATC was divided by 

an assessment factor (UF) of 10, according to established 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 

EC50 0.293 Immobilization

Aquatic Invertebrates - Saltwater 

Marine 

copepods 

(Acartia tonsa) 

Salt 

OECD TG 

Draft Invert 

Life Cycle/ 

Static 

6-day 

EC10 0.044 

Measured Development Bjornestad 

(2007)b

EC50 0.131 

NOEC 0.038 

LOEC 0.075 

5-day EC10 0.037 Nominal Development Wollenberger

EC50 0.059 et al. (2003) 

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 

(Pseudokirchneriella 

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 organisms. 

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 including amphipods, midges, oligochaete 

worms, polychaete worms, and mud snails are summarized 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 (Pennak, 1978). The study was conducted using 

established EPA and OECD test guidelines for measuring 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 New Zealand mud snail, it is an invasive species in 

the US and currently is not ubiquitously distributed 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 marine organisms appears, based on 

available data, to be less than an order of magnitude, 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., survival 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 exposed 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 (measured 16.3 mg/kg). The reported LOEC was 35 

mg/kg (measured, 16.3 mg/kg) and NOEC was 14.5 mg/kg (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 7.1b

Measured Growth Egeler 

(2004)c

LOEC 16.3b

MATC 10.8 

Midge 

(Chironomus 

riparius)

Fresh OECD 218/

Renewal 28-day 

NOEC 200b

Measured Emergence Egeler & 

Gilberg 

(2004a)c

LOEC 400b

EC50 335b Development 

Oligochaete 

(Lumbriculus 

variegatus)

Fresh OECD 218/

Renewal 28-day 

NOEC 16.2 

Measured Reproduction Egeler & Gilberg 

(2004b)c

LOEC 36.5 

EC50 44.5 Reproduction 

Saltwater 

New Zealand 

mud snails 

(Potamopyrgus 

antipodarum)

Fresh/ 

Brackish Renewala

 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 Renewala

 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-dwelling 

organisms. 

a

 Test guideline not reported 

b

 Concentration adjusted based on measured concentrations (recovery) 

c

 as reported in EC, 2008 

EPA/OPPT calculated at a Maximum Acceptable Toxicant 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 assessment 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 considered for this assessment is summarized in 

Table 3-5. 

Invertebrates 

The earthworm (Eisenia fetida) was the most sensitive terrestrial (soil) invertebrate species 

tested (Chen et al., 2011a). In a two part test, Chen et al. (2011a) observed the lethality of 

earthworms exposed to HHCB in a 14-day study and the chronic effects in a 28-day study. In the 

acute study, adult earthworms were exposed for 14 days to concentrations of 0, 100, 140, 196, 

274.4, 384.2, 537.8, 752.9, 1,054.1, and 1,475.8 mg HHCB (99 percent purity)/kg (air-dried soil) 

in acetone (Chen et al., 2011a). Observations were taken on days 7 and 14. No mortalities 
occurred at the lowest concentrations of 100 and 140 mg/kg. However, there was 100% 

mortality at ≥ 1,054.1 mg HHCB /kg. The 7-day LC50 was reported as 489 mg HHCB/kg and the 

14-day LC50 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. After 28-days of exposure to HHCB there was a 

significant decrease in growth rate of earthworms which occurred at 100 mg HHCB /kg. 

Additionally, a significant decrease in reproduction rate and cocoon production in earthworms 

exposed to 50 and 100 mg/kg were noted. A LOEC of 50 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 invertebrates, EPA/OPPT did not calculate a COC 

for the earthworm. 

Plants 

For plants, two studies (An et al., 2009 and Chen et al., 2010; summarized in Table 3-5) were 

located that evaluated the toxicity of HHCB to wheat 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 toxicity data for terrestrial plants coupled 

with only limited soil concentration data and external expert advice, does not support assessing 

risks to terrestrial plants and hence, COCs were not 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 LC50 32.60 

μg/cm2

 not reported Mortality Chen et al. 

(2011b) 48-hr LC50 11.87 

72-hr LC50 6.14

OECD 

207/222*/

Not 

reported

7-day LC50 489 

mg/kg not reported

Mortality 

Chen et al. 

(2011a) 

14-day LC50 392.4 

28-day 

NOEC 30

LOEC 50 Reproduction 

MATC 38.7 

Nematode 

(Caenorhabditis elegans)

Not 

reported

24-hr LC50 >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 EC50

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 

LC50 846.6 

EC50 2,123 

EC50 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 Concentration 
3.3.5 Conclusions of Environmental Hazard Assessment 

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

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

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

aquatic, soil, and sediment-dwelling organisms are summarized in Table 3-6. 

Table 3-6. Concentrations of Concern (COCs) for Environmental Toxicity

Environmental Toxicity Concentration of Concern (COC) 

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 characterization was to determine whether there are risks 

to the aquatic and benthic environments from the measured levels of HHCB found in 

wastewater, surface water, or sediment. The data for environmental monitoring and toxicity 

were used in this risk assessment to address the following key questions: 

1. Does acute exposure to levels of HHCB measured in wastewater 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 was 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 the same as the concentration that causes 

effects. If the RQ is above 1, the exposure is greater than the effect concentration. If the RQ is 
below 1, the exposure is less than the effect concentration. The Concentrations of Concern 

(COCs) for aquatic and benthic organisms shown in Table 3-6 and the environmental 

concentrations shown in Table 3-7 were used to calculate 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 insufficient 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 to be continuous based on constant 

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

supported by the fact that HHCB is frequently measured in surface water downstream from 

WWTPs even though 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 risk 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 and sediment, and also for the mean and 95th

percentile for surface water at outfall sites and bottom 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 aquatic concentrations for HHCB that 

would be experienced acutely in streams that receive wastewater. 

Table 3-7. Environmental Concentrations Used to Calculate RQs

Maximum 

Published 

Concentration

Range of Mean 

Published 

Concentrations

95th 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 reflect the range of concentrations of HHCB 

measured in the environment. In addition to the maximum published surface water values, the 

acute and chronic RQs for aquatic organisms in surface water were also calculated from the 95th

percentile USGS NWIS data using the concentrations measured in surface water at outfall sites. 

Though the concentration of HHCB in effluent discharge may be diluted in surface water 

downstream from the WWTP outfall, the calculated RQs for chronic risk to aquatic organisms 

are considered conservative estimates based on actual surface water concentrations. 
Table 3-8. Calculated Risk Quotients (RQs) for HHCB

Calculated RQs 

Maximum 

Published 

Concentration

Range of Mean 

Published 

Concentrations

95th

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 organisms was 0.006 to 0.04 (Table 3-8). The 

highest RQ of 0.04 was calculated using the 95th 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 organisms 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 organisms was 0.04 to 0.23 (Table 3-8). The 

highest RQ of 0.23 was calculated using the 95th 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 organisms in Table 3-8 range from 4 to 25 times or 

an order of magnitude below 1. 

The maximum concentration for surface water as measured at an outfall site from the USGS 

dataset is 4.4 μg/L, a value well below both the chronic and acute concentration of concern for 

aquatic organisms. 

The range of RQs for sediment-dwelling organisms was 0.001 to 0.36 (Table 3-8). The highest 

RQ of 0.36 was calculated using the maximum published level of HHCB in sediment of 388 

μg/kg dw (Table 3-1). This HHCB sediment concentration is from a single sample taken from 

the upper Hudson River near Troy, NY (Reiner, 2011) and a second sample (from the same 

study) taken near Albany, NY was 351 μg/kg. Both of these values are above the 95th percentile 

value from the USGS monitoring data for bottom material. It should be noted that of the more 

than 600 bottom material samples available in the USGS dataset, a single value was identified 

as greater than the COC of 1080 μg/kg dw. This value is presumed to represent a “hotspot” as 

>99.9% of the remaining values were below the COC for 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 calculate 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 adequate ecological toxicity data set. 

However, a number of limitations in the approach create uncertainties associated with this 

assessment. 

3.4.2.1 Representativeness of Exposure Concentrations 

Available monitoring data for HHCB in water, sediment, and soil were used in the assessment. 

A robust dataset of over 6800 sampling measurements were available from the USGS NWIS for 

wastewater effluent, surface water, and bottom material along with over 500 sampling 

measurements from the scientific literature. The data from both published studies and the 

USGS NWIS database were used because the mean values for surface water and sediment were 

within the same order of magnitude. The concentrations reported in the literature and in the 

USGS NWIS data are assumed to represent reasonable estimates, however the data may not 

reflect the actual distribution of HHCB in the environment. Calculation of the RQ based on the 

maximum value, the 95th percentile, and the highest mean value was performed in an effort to 

capture the full range of concentrations in the environment. 

3.4.2.2 Variability in Environmental Concentrations 

Although calculation of the RQ based on the maximum reported value from the literature, the 

95th percentile USGS NWIS value for surface water at effluent sites, and the highest mean value 

was performed in an effort to capture the full range of concentrations observed in this data set, 

higher values may exist in the environment. Variation of use practices across the population 

and differences in sewage treatment plant inputs (industrial, commercial or residential), 

removal efficiencies and receiving streams, may affect the final concentrations present in any 

single location or point in time. Unless actual environmental concentrations of HHCB in surface 

water or sediment are greater than one or two orders of magnitude higher than expected, 

additional measurements or estimates will not affect the risk calculation. 

Use of all available data obscures trends such as effects 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 particular 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 depending upon other factors such as input and 

receiving environment. Well-designed studies that follow 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 concentrations in any given media or location. 
3.4.2.3 Anaerobic Degradation 

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

sludge volume and organic load before disposal of the concentrated material. Removal of HHCB 

under anaerobic treatment conditions has been addressed in several studies. Xue et al. (2010) 

reported poor overall removal of HHCB in a full-scale anaerobic/anoxic/aerobic system 

combined with membrane bioreactor for municipal wastewater reclamation, and little removal 

was attributed to biodegradation. In contrast Carballa et al. (2007) reported good removal of 

HHCB (60-80%) in mesophilic and thermophilic anaerobic 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. Therefore, no direct evidence of anaerobic 

(methanogenic) biodegradation of HHCB exists, and this 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 studies are cited in the section on fate in water, 

and they may appear at first glance to be in conflict, however the amount of volatilization in lab 

studies versus that in the environment is highly dependent 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 and metabolites is very limited. The 

measurement of specific isomers of HHCB were not described in the collected studies or USGS 

data set reviewed for this assessment and only a very 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 be needed to refine this assessment and 

provide a fuller understanding of if or how specific HHCB isomers or metabolites contribute to 

the observed HHCB behavior in environmental media and their impact on organisms present in 

water, sediment and soil. 

3.4.2.6 Deriving Concentrations of Concern from Single Species Tests 

Ecological hazard of industrial chemicals is routinely evaluated using a single species and within 

a controlled laboratory setting in the form of a toxicity test. There is uncertainty associated with 

extrapolating these single-species laboratory test results to concentrations intended to be 

protective of all species and environments. Application 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 concentration of concern; COC) that would likely 

encompass more sensitive species not specifically represented by the available experimental 

data. Uncertainty factors are included in the COC calculation 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 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 conducted under TSCA, since the data available for most 

industrial chemicals is limited (Ahlers et al., 2008). 

3.4.2.7 Assessment of Risk to Terrestrial Invertebrates or Plants 

Two studies with six total samples reported HHCB concentrations in soil following land 

application of either biosolids or wastewater effluent. The locations are limited to two sites and 

may not representative of overall US soil concentrations given differences in sludge application 

practices, WWTP contaminant removal, or seasonal factors. Measured concentrations of HHCB 

in non-amended soil were not available for comparison. The limited data available on measured 

levels of HHCB in biosolids or wastewater-amended soil, prevented quantitative assessment of 

risk to terrestrial invertebrates and terrestrial plants. 

3.4.3 Conclusions of Risk Characterization 

RQs were calculated based on measured values for the maximum published levels and range of 

means from published studies of surface water and sediment, and also for the mean and 95th

percentile for surface water at outfall sites and bottom material from the USGS NWIS data in an 

effort to capture the spectrum of environmental concentrations. Uncertainty factors were used 

to account for inter-species and lab-to-field extrapolation of hazards for organisms exposed 

HHCB in the ambient environment. Risk estimates for terrestrial environments were not 

performed due to very limited monitoring and toxicity 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 commercial products. Scenarios of interest included 

release of HHCB in wastewater to surface water or application of biosolids to land. The 

assessment endpoints were acute and chronic toxicity to aquatic organisms and chronic toxicity 

to sediment organisms. The risk assessment considered only direct exposure to aquatic 

organisms and sediment-dwelling organisms. 

Acute and chronic RQs indicated that maximum measured 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 environmental concentrations for the maximum 

published values or the more than 600 sediment measurements from the USGS. Low risk 

concerns were identified for both acute and chronic toxicity to aquatic organisms and for 

sediment-dwelling organisms. Therefore, unless environmental concentrations increase by 

factor of 1 to 2 orders of magnitude, the assessment does not indicate risk concerns. 

Insufficient data were available to calculate chronic 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 EU RAR on HHCB (EC, 2008). Other reviews 

are also available (OECD, 2009; EPA, 2003; HERA, 2004a; 2004b). This review is not intended to 

be an exhaustive discussion of the toxicity of HHCB, but rather a summary of the available data. 

The reader is referred to the original documents for more detailed information. 

A-1 Human Hazard Characterization 

A-1-1 Toxicokinetics (Absorption, Distribution, Metabolism, 

Excretion) 

Available toxicokinetics data indicate that HHCB has poor dermal absorption and is extensively 

metabolized and excreted, with no evidence of significant bioaccumulation. An in vivo human 

study reported approximately 0.1 percent dermal absorption following exposure to HHCB 

(Hawkins et al., 1996a; Ford et al., 1999; as cited in EC, 2008), while an in vitro dermal 

absorption study with human epidermal membranes showed 5.2 percent absorption (Green 

and Brian, 2001; as cited in EC, 2008). Human biomonitoring data show that HHCB has been 

detected in blood, milk, and adipose tissue following dermal exposures (see Table A-2). Data in 

humans after oral and inhalation exposures, including half-life data, are not available. In 

animals, dermal exposures in rats showed 16 percent absorption, with distribution to the 

stomach, liver, fat, plasma, adrenal glands, kidneys, and thyroid, with excretion primarily in the 

feces (Ford et al., 1999; Hawkins et al., 1995; as cited in EC, 2008). Intravenous studies in rats 

and pigs indicated that HHCB is rapidly distributed and excreted primarily in the feces in the rat 

and in the urine in pigs (Hawkins et al., 1997; Hawkins, 1997; as cited in EC, 2008). No evidence 

of accumulation in the feces or urine was seen in either the rat or pig. Apparent half-lives of 

elimination in the blood were longer in the rat than in the pig, with clearance from fat and skin 

considerably slower. HHCB and metabolites were found in the milk of pregnant and lactating 

rats exposed via the oral route (Hawkins et al., 1996a; as cited in EC, 2008). Toxicokinetics data

in animals after oral and inhalation exposures are not available. 

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; as cited in EC, 2008). Acute toxicity studies 

by the inhalation route are not available. The acute toxicity of HHCB is low via the oral route in 

rats and low via the dermal route in rabbits, with LD50 values >3,000 mg/kg-bw. 
A-1-3 Subchronic/Repeated-Dose Toxicity 

An oral range-finding study in rats with mean achieved intakes up to 829 mg/kg-bw/day 

reported signs of liver toxicity at 347 mg/kg-bw/day (lowest-observed-adverse effect 

concentration [LOAEL] (Api and Ford, 1999; Hopkins et al., 1996; as cited in EC, 2008); however, 

the follow-up 13-week repeated-dose oral (dietary) toxicity study showed no adverse, 

treatment-related effects on any parameters measured. The no-observed-adverse-effect level 

(NOAEL) for systemic toxicity was 150 mg/kg-bw/day, the highest dose tested (Api and Ford, 

1999; Hopkins et al., 1996; as cited in EC, 2008). In that same study, no effects on organ weight 

or histopathology were reported in the reproductive organs at any dose level. Repeated-dose 

subchronic toxicity studies by the inhalation and dermal routes were available, but were 

considered to be of limited value because they were not conducted according to guidelines or 

GLPs or they used fragrance mixtures in which HHCB was only present at low levels. 

A-1-4 Reproductive Toxicity and Fertility 

No combined repeated-dose/reproductive/developmental or multigenerational reproductive 

toxicity studies on HHCB are available. Information on reproductive organs is available from the 

repeated-dose dietary toxicity study in rats; this study did not show any signs of toxicity to the 

reproductive organs examined at doses as high as ~150 mg/kg-bw/day. 

A modification to the ICH Guideline on Detection of Toxicity to Reproduction for Medicinal 

Products (http://www.ich.org/products/guidelines/safety/article/safety-guidelines.html) was 

conducted in rats in order to assess the potential for adverse effects to the neonate following 

exposure to HHCB through nursing. In this comprehensive study, the basis for the dose 

selection for this study was a pharmacokinetic analysis by Hawkins et al. (1996b) (as cited in EC, 

2008) that determined the oral doses required to produce levels in the milk of lactating rats 

similar to those reported in human milk and several orders of magnitude higher. At oral doses 

of 2 and 20 mg/kg-bw/day, HHCB was found in the milk of lactating dams at 2.28 and 32.8 

mg/L, respectively (see Human Biomonitoring below for comparison to levels in human breast 

milk). Five groups of pregnant Sprague-Dawley rats (28/dose) were administered 0, 2, 6, or 20 

mg/kg-bw/day HHCB (>95 percent purity) by gavage from gestation day (GD) 14 through 

weaning on lactation day 21 (Ford and Bottomley, 1997; Jones et al., 1996; as cited in EC, 2008). 

Dams were allowed to litter and rear their offspring to weaning. F1 offspring were exposed to 

HHCB in utero from GD 14 through lactation via the dam’s milk, and exposures ceased after 

weaning in the F1 offspring and in the F2 generation. Since milk consumption was not measured 

in the lactation study, actual intake of HHCB could not be determined. 

After parturition, the offspring were counted, sexed, weighed, and examined for external 

abnormalities. On postnatal day 4, the pups were weighed and the litters were culled to four 

males and four females. During the pre-weaning period, pups were examined for several 

developmental milestones, including surface righting reflex, startle reflex, air righting reflex, 

and pupil reflex. Selected offspring from these litters (24/sex/dose) were allowed to reach 

sexual maturity and were then assessed for behavioral changes in motor coordination and 
balance, activity, and avoidance. When the F1 generation offspring reached 84 days of age, they 

were mated (avoiding sibling pairings) and the reproductive capacity was assessed by 

examining the females before and after mating to determine time of pregnancy, estrous 

cyclicity, pre-coital time, pregnancy rates, and duration of gestation. F2 generation offspring 

were examined for abnormalities at parturition and then periodically until study termination on 

postnatal day 21. No adverse, treatment-related effects were reported in the dams or in either 

generation of offspring for any parameters assessed up to and including 20 mg/kg-bw/day. The 

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 developmental toxicity (Christian et al., 1997; Christian

et al., 1999; as cited in EC, 2008), four groups of 25 pregnant Sprague-Dawley rats were 

exposed to 0, 100, 250, 500, and 1,000 mg/kg-bw/day HHCB (purity not specified) by gavage on 

GDs 7 to 17. Three dams were found dead at 1,000 mg/kg-bw/day. Maternal body weights 

were reduced (approximately 10 percent) in the three highest dosage levels. No other effects 

were reported in the dams. No signs of gross fetal malformations were reported; however, fetal 

body weights were reduced at the two highest doses. 

Based on the results of the dose range-finding study described above, pregnant SpragueDawley rats (25/dose) were administered 0, 50, 150, and 500 mg/kg-bw/day HHCB (>95 percent 

purity) by gavage on GDs 7 to 17 (Christian et al., 1997; Christian et al., 1999; as cited in EC, 

2008). The dams were observed for clinical signs of toxicity. Food and water consumption and 

body weights were recorded at various intervals. On GD 20, the dams were sacrificed and a 

gross necropsy was performed. The numbers of corpora lutea, implantation sites, resorption 

sites, live and dead fetuses, and live pup body weights were recorded. Fetuses were examined 

for sex and underwent a gross examination. One half of the fetuses was examined for visceral 

anomalies and the other half was examined for skeletal abnormalities. Signs of maternal 

toxicity in the dams consisted of significant increases in salivation, urine-stained fur, red or 

brown substance on the forepaws, and alopecia beginning at 150 mg/kg-bw/day, with 

statistical significance at 500 mg/kg-bw/day. At ≥150 mg/kg-bw/day, statistically significant 

reductions in food consumption and maternal body gain were observed, with the most severe 

reductions in body weight occurring during GDs 7 to 10. After completion of the dosing period 

(GDs 18 to 20), a significant rebound in food consumption and body weight gains occurred, with 

values reported as being comparable to controls. Signs of developmental toxicity were confined 

to the highest dose of 500 mg/kg-bw/day. Significant reductions in offspring body weight were 

reported at the high dose; however, according to the study authors, when severity, dose 

relationships, and historical ranges were taken into consideration, these body weight 

reductions were not definitively considered to be treatment-related. The number of litters or 

fetuses with morphological alterations (malformations and variations) reported at the high 

dose was not statistically significantly increased when compared to the control group. 

Malformations that did occur were few in number and were considered spontaneous in origin, 

except for the vertebral/rib malformations reported in only three fetuses that were not litter 
mates. No other parameters were reported to be affected by treatment at any dose level. The 

LOAEL for maternal toxicity was 150 mg/kg-bw/day, based on increases in clinical signs of 

toxicity and decreases in body weight, and the NOAEL was 50 mg/kg-bw/day. Although there is 

uncertainty concerning the effects reported in the offspring, a conservative LOAEL for 

developmental toxicity of 500 mg/kg-bw/day was concluded and the NOAEL was 150 mg/kgbw/day. 

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 aberrations in mammalian cells (San et al., 1994; 

Api and San, 1999; Mersch-Sundermann et al., 1998a; 1998b; Kevekordes et al., 1997; Curry 

and Putman, 1995; Gudi and Ritter, 1997; San and Sly, 1994; Kevekordes et al., 1998; as cited in 

EC, 2008). 

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, subchronic, 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. No data on respiratory tract irritation were 

available. HHCB was considered to be a minimal eye irritant in rabbits and a possible photoirritant in rabbits and guinea pigs, but not in humans (Guillaume et al., 1973a; 1973b; 

Rubenkoenig and Ede, 1964; Lindstrum et al., 1978a; 1978b; 1978c; Harrison and Stolman, 

1986; Gabriel and Mark, 1987; Folk and Dammers, 1987; Shanahan and Alworth, 1987; Haynes, 

1984; 1985; 1986; Sauer, 1980; Levenstein, 1973a; 1973b; 1975a; 1975b; Wolven and 

Levenstein, 1963; Basketter, 1996; Parish, 1988; as cited in EC, 2008). 

In a series of in vitro assays with HHCB using different protocols designed to test for potential 

endocrine disruption (i.e., estrogenic and/or anti-estrogenic activity), HHCB showed weak 

estrogenic and anti-estrogenic activity in vitro, dependent on the estrogen receptor type 

(Seinen et al., 1999; Bitsch et al., 2002; Schreurs et al., 2002; Schreurs et al., 2004; Schreurs et 

al., 2005a; Schreurs et al., 2005b; Gomez et al., 2005; as cited in EC, 2008). Marginal repressing 

effects were also found in vitro on the androgen and progesterone receptor. However, no 

estrogenic effects were seen in the in vivo uterotrophic assay. 

Possible anti-estrogenic effects in zebrafish were assessed (Schreurs et al., 2004; as cited in EC, 

2008). Transgenic zebrafish were exposed to 0.01, 0.1, 1, and 10 μM HHCB with and without 

estradiol (E2). The highest concentration of 10 μM HHCB was toxic. Concentrations of 0.01, 0.1, 

and 1 μM HHCB were not estrogenic. Dose-dependent antagonistic effects were reported at 0.1 

and 1 μM levels.
Table_Apx A-1. Summary of Human Health Hazard Information

Endpoints HHCB CASRN (1222-05-5) Referen

(as cited in E

Acute oral toxicity

LD50 (mg/kg-bw) >3,000 

Minner and Foster (1977); M

Acute dermal toxicity

LD50 (mg/kg-bw) >3,250 

Minner and Foster (1977); M

Acute inhalation toxicity

LC50 (mg/L) 

No data

Repeated-dose toxicity 

Oral (mg/kg-bw/day) Dermal (mg/kg-bw/day) Inhalation (mg/L/day)

NOAEL = 150 (highest dose tested) 

No adequate data No adequate data 

Api and Ford (1999); Hopkin

Reproductive toxicity

Oral (mg/kg-bw/day) Maternal NOAEL/LOAEL Reproductive NOAEL/LOAEL Developmental NOAEL/LOAEL 

NOAEL = 20 (highest dose tested) NOAEL = 20 (highest dose tested) NOAEL = 20 (highest dose tested) 

Ford and Bottomley (1997); 

Developmental toxicity Oral (mg/kg-bw/day), rats 

Maternal NOAEL/LOAEL Developmental NOAEL/LOAEL 

NOAEL = 50/LOAEL = 150 

NOAEL = 150/LOAEL = 500 

Christian et al. (1997); Chris

Genetic toxicity, gene mutation

In vitro 

Genetic toxicity, gene mutation 

In vivo 

Negative Negative 

San et al. (1994); Api and SaSundermann et al. (1998a); al. (1998b) 

Genetic toxicity, chromosomal aberrations

In vitro 

Genetic toxicity, chromosomal aberrations 

In vivo

Negative Negative 

Api and San (1999); Kevekoand Putman (1995); Gudi anSly (1994); Kevekordes et al
Table_Apx A-1. Summary of Human Health Hazard Information

Endpoints HHCB CASRN (1222-05-5) Referen

(as cited in E

Additional informationCorrosivity/skin-eye Irritation/sensitization 

Respiratory tract irritation Carcinogenicity Endocrine disruption In vitro In vivo

Negative 

No data No data 

Weak 

Negative 

See 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 numbers of subjects, limited geographic 

representation, high variability, and incomplete details on methodology. 

Several small studies have reported concentrations of HHCB in limited numbers of the general 

population in Asia and Europe. Only two small studies have reported exposure levels in the US. 

To date, HHCB has been measured in adipose tissue, blood, breast milk, and umbilical cord 

blood. Most of these studies reported data for <100 samples and several of them did not 

provide data tables or details on the results of the data. However, HHCB was detected in a high 

majority of the samples collected in the studies reported here. Therefore, it can be assumed 

that exposure is widespread. 

Both US studies took place in 2004 in the northeastern part of the US. A study on adipose tissue 

in 49 residents of New York City undergoing liposuction reported a mean concentration of 178 

ng/g lipid weight (lw) (range: 12 to 798 ng/g lw) (Kannan et al., 2005), which falls between the 

reported concentrations of HHCB in adipose tissue from two other available studies (Moon et 

al., 2012; Schiavone et al., 2010). A study of 43 Korean women reported a mean of 81 ng/g lw in 

adipose tissue (range: 28 to 211 ng/g lw) (Moon et al., 2012), while a small study in Italy 

reported a mean of 361 ng/g lw (range: 28 to 211 ng/g lw) (Schiavone et al., 2010). 

Several small studies on concentrations of HHCB in breast milk have been conducted. In the US, 

a mean of 227 ng/g lw was reported for 39 samples collected in Massachusetts (Reiner et al., 

2007b). The levels reported in this small study are much higher than those reported in other 

countries to date (see Table A-2). However, in order to compare the results, the timing of the 

collection of the samples must be considered (length of time after birth), and it is not clear from 

the report when the US samples were collected. 

No studies in the literature reported blood concentrations of HHCB in the US. A collection of 

matched maternal and cord blood samples in the Netherlands indicated high detection and 

correlation between maternal and cord blood serum. Concentrations reported in studies in the 

Netherlands, Austria, and China are so variable that they cannot and should not be compared 

(see Table A-2). Little data were provided in these studies, which were primarily done in 2005. 

Given that HHCB is sequestered in the fat, blood may not be an appropriate matrix for 

biomonitoring. 

High variance in measurements may be a result of the analytical methods for measuring HHCB 

in human matrices as they are still being developed and modified because the sample sizes are 

very small, the results should not be extrapolated to larger populations. Also, data from other 

countries may not represent those of the US since exposure patterns may vary greatly between 

countries, especially for personal care products such as those that contain HHCB. 
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 been extensively assessed by several other 

reliable entities (EC, 2008; OECD, 2009; EPA, 2003; HERA, 2004a; 2004b). Each of these welldocumented, independent reviews concluded that there was no significant risk to human 

health from exposure to HHCB as used in household cleaning products. Moreover, the overall 

conclusions of the EU RAR (EC, 2008) were that there was no need for further information 

and/or testing and no need for risk reduction measures beyond those already being applied. 

The EPA/OPPT has thoroughly reviewed these other assessments 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 populations for relevant endpoints: workers

(exposures from handling during production and dilution, compounding, formulation, and 

professional cleaning); consumers (exposures from a wide variety of consumer products such as 

perfumes, creams, toiletries, soaps, shampoos, and household and laundry cleaning products); 

and humans exposed via the environment (exposures via food and water including fish, root 

crops, and mother’s milk). 

Dermal absorption of 16 and 5.2 percent in rats and humans, respectively, was used. Since 

toxicokinetic data of HHCB following oral and inhalation absorption are not available in humans 

or animals, intermediate defaults of 50 percent for oral absorption and 100 percent absorption 

from inhalation exposures were used. 
For general systemic effects following repeated exposures, the oral NOAEL of 150 mg/kgbw/day from the 13-week subchronic toxicity study in rats was used as a point of departure. 

Since NOAELs are not available following dermal or inhalation exposures, the administered 

dose from the oral 13-week toxicity study was converted into an internal dose (body burden) by 

taking into account different absorption factors. Based on an oral absorption value of 

50 percent in humans and animals, an internal NOAEL of ≥75 mg/kg-bw/day was used, with 

100 percent absorption assumed for inhalation exposures and 5.2 percent absorption for 

dermal exposures. 

For developmental/reproductive toxicity, the NOAEL of 20 mg/kg-bw/day (the highest dose 

tested) from the peri/post-natal oral (gavage) toxicity study was used as a point of departure 

(Ford and Bottomley, 1997; Jones et al., 1996; as cited in EC, 2008). Based on an oral absorption 

rate of 50 percent, an internal NOAEL of ≥10 mg/kg-bw/day was used, with 100 percent 

absorption assumed for inhalation exposures and 5.2 percent absorption assumed for dermal 

exposures. 

Worst-case estimates were based upon combined (simultaneous) dermal and inhalation 

exposures for compounding workers only (the only scenario relevant for combined exposures). 

A-3-2 Risk to Workers 

Oral exposures in workers were assumed to be mitigated by personal hygiene measures, and 

therefore, only risks following dermal and inhalation exposures in the workplace were assessed. 

For acute toxicity following dermal exposures, high LD50 values (>3,000 mg/kg-bw) indicated no 

concern for workers. Although acute toxicity studies by the inhalation route are not available, 

low-level, short-term exposures in workers, combined with the low acute toxicity by the oral 

route, suggested no concern for workers for acute toxicity by the inhalation route. 

Given that HHCB showed no potential for skin/eye irritation, corrosivity, or sensitization/ 

photosensitization, it was concluded that there was no concern for local effects in workers 

following repeated exposures. Likewise, since there is a lack of skin and eye irritation potential, 

no significant respiratory tract irritation potential was expected. 

For both the dermal and inhalation exposure scenarios for the general systemic toxicity 

endpoint, a minimal margin of safety (MOS) of 100 was used for comparison. The MOS was 

based on an interspecies factor of 10 (4 for metabolic size differences and 2.5 for remaining 

differences), an intraspecies factor of 5, and a factor of 2 for extrapolation from semi-chronic to 

chronic exposure. For each inhalation/dermal exposure scenario, the calculated MOS was well 

above the minimal MOS of 100 (≥2,600). Combined exposure routes for total body burdens for 

skin contact and inhalation for the compounding scenario resulted in MOS values well above 

100 (≥2,000). As a result, it was concluded that for workers under these exposure scenarios for 

general systemic toxicity, there is, at present, no need for further information and/or testing 

and no need for risk reduction measures beyond those already being applied. 
Given that HHCB was not found to be mutagenic in a wide array of studies, it was concluded 

that for mutagenicity for workers under these exposure scenarios, there is, at present, no need 

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

already being applied. Although there are no carcinogenicity data for HHCB, it was concluded 

that there was no concern for workers under these exposure scenarios for carcinogenicity 

based on negative mutagenicity data and the lack of any apparent structural alerts that would 

raise a concern. 

 

For both the dermal and inhalation exposure scenarios for the developmental/reproductive 

toxicity endpoint, a minimal MOS of 50 was used for comparison. The MOS was based on an 

interspecies factor of 10 (4 for metabolic size differences and 2.5 for remaining differences) and 

an intraspecies factor of 5. For dermal, inhalation, and combined exposure scenarios, the 

calculated MOS was above 50 (≥261). As a result, it was concluded that for workers under these 

exposure scenarios for developmental/reproductive toxicity, there is, at present, no need for 

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

already being applied. 

 

A-3-3 Risk to Consumers 

The main route of exposure for consumers was assumed to be dermal, with some inhalation 

exposures and no oral exposures. 

Given that HHCB showed no potential for skin/eye irritation, corrosivity, or sensitization/ 

photosensitization, it was concluded that there was no concern for local effects in consumers 

following repeated exposures. Since there is a lack of skin and eye irritation potential, no 

significant respiratory tract irritation potential was expected. 

 

For general systemic effects following repeated exposures, a MOS of 200 (based on a factor of 

10 for intraspecies differences, 4*2.5 for interspecies differences, 2 for duration extrapolation, 

and 1 for dose-response) was used. The calculated MOS for dermal exposure was well above 

200 (≥1,400). Therefore, it was concluded that there was no concern for general systemic 

toxicity in consumers following repeated dermal exposures. 

Given that HHCB was not found to be mutagenic in a wide array of studies, it was concluded 

that for mutagenicity for consumers, there is, at present, no need for further information 

and/or testing and no need for risk reduction measures beyond those already being applied. 

Likewise, although there are no carcinogenicity data for HHCB, it was concluded that there was 

no concern for consumers under these exposure scenarios for carcinogenicity based on 

negative mutagenicity data and the lack of any apparent structural alerts that would raise a 

concern. 

For the dermal exposure scenario for the developmental/reproductive toxicity endpoint, a 

minimal MOS of 100 was used for comparison. The MOS was based on an intraspecies factor of 
10, an interspecies species factor of 10 (4 for metabolic size differences and 2.5 for remaining 

differences), and factor of 1 for dose-response. The calculated MOS was above 100 (MOS ≥189). 

It was concluded that for consumers with a dermal exposure scenario for developmental/ 

reproductive toxicity, there is, at present, no need for further information and/or testing and no 

need for risk reduction measures beyond those already being applied. 

A-3-4 Risk to Humans Exposed Indirectly via the Environment 

Exposure to HHCB by the inhalation route (via air) was considered to be negligible. The main 

route of exposure to humans in the environment was oral (via fish and root crops). Exposures 

were based on local and regional daily intake estimates following production. A separate risk 

characterization for breast-fed babies was conducted for exposure via mother’s milk (see next 

section for more information). 

For general systemic effects following repeated exposures, an internal NOAEL of ≥75 mg/kgbw/day and a MOS of 200 (based on a factor of 10 for intraspecies differences, 4*2.5 for 

interspecies differences, 2 for duration extrapolation, and 1 for dose-response) were used. The 

calculated MOS for exposure via food was >3E+4 for the local production scenario and 

>8E+5 for the regional production scenario. Therefore, it was concluded that for the local and 

regional exposure scenarios, there was no concern for general systemic toxicity in humans 

exposed indirectly via the environment. 

As with the other populations, since HHCB was not found to be mutagenic in a wide array of 

studies, it was concluded that for mutagenicity, there is, at present, no need for further 

information and/or testing and no need for risk reduction measures beyond those already 

being applied. Likewise, although there are no carcinogenicity data for HHCB, it was concluded 

that there was no concern for carcinogenicity based on negative mutagenicity data and the lack 

of any apparent structural alerts that would raise a concern. 

For the oral exposure scenario for the developmental/reproductive toxicity endpoint, an internal 

NOAEL of ≥10 mg/kg-bw/day and a MOS of 100 were used. The MOS was based on an 

intraspecies factor of 10, an interspecies species factor of 10 (4 for metabolic size differences 

and 2.5 for remaining differences), and factor of 1 for dose-response (based on the lack of effect 

at the highest dose tested). The calculated MOS values for local and regional oral exposure 

scenarios were 3,846 and 10,000, respectively, indicating no concern for developmental/ 

reproductive toxicity in humans exposed indirectly via the environment. 

A-3-5 Assessment of Risk for Breast-Fed Babies Exposed via 

Mother’s Milk 

The presence of HHCB in human milk was considered to be the result of indirect environmental 

exposures from a variety of sources, including maternal exposure to consumer products; intake 

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

dose level in dams (20 mg/kg-bw/day, the highest dose tested from the oral peri/postnatal 

developmental toxicity study), the measurement of levels of HHCB in the milk of dams 

(17.6 and 5.0 μg/mL at four or eight hours post dosing, respectively), with the maximum level of 

HHCB found in human milk samples (1,316 μg/kg fat), the EU RAR (EC, 2008) concluded that 

pups in the high-dose group were estimated to be exposed to levels approximately 100 to 

360 times the maximum level found in human milk samples. 

The amount of milk consumed by infants and rat pups was compared. Assuming 50 percent 

absorption of the ingested HHCB, the average daily uptakes (ADUs) for the breast-feeding infant 

for 0 to 3 and 3 to 12 months, as well as the average daily milk consumption for the rat pup, 

were estimated. A comparison of these two estimates of uptake showed a difference of 

approximately three orders of magnitude between the levels of HHCB exposure in the rat (in 

which no adverse effects were found) and the human infant exposure. 

Total internal (worst-case) combined exposures were estimated and compared to the internal 

NOAELs for general systemic toxicity following repeated exposures and for developmental/ 

reproductive toxicity (≥75 and ≥10 mg/kg-bw/day, respectively) in order to calculate MOS 

values. The worst-case combined exposure was estimated from the sum of the worst-case 

estimates from three populations: (1) dermal and inhalation exposures for scenario 2 for 

compounding workers; (2) dermal and inhalation exposures for consumers; and (3) oral and 

inhalation, locally via the environment. For general systemic effects following repeated 

exposures, a MOS of 100 was used (based on 4*2.5 for metabolic size and other differences, a 

factor of 5 for intraspecies differences, and a factor of 2 for semichronic to chronic exposure 

extrapolation). The calculated MOS was well above 100 (798). For the developmental/

reproductive toxicity endpoint, a MOS of 50 (based on 4*2.5 for interspecies species differences 

and 5 for intraspecies differences) was used. The calculated MOS was above 50 (≥106). 

As a result, the EU concluded that, overall, there was no concern for breast-fed babies exposed 

indirectly via the environment and no need for further information and/or testing or risk 

reduction measures beyond those already being applied. 

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

Health 

Overall, adequate screening-level animal toxicity data are available to characterize the human 

health hazard for HHCB. Toxicokinetics (by the dermal route), acute, repeated-dose, and 

developmental toxicity data are available to characterize the human health hazard of HHCB. 

Although no multigenerational reproductive toxicity studies on HHCB are available, information 

on developmental and reproductive toxicity was obtained from both the repeated-dose dietary 

toxicity study (reproductive organ data) and the peri/post-natal reproductive toxicity study with 

modified exposures. Several assays testing for endocrine disruption, genotoxicity, and 

irritation/sensitization (including several studies in humans) are also available. The database is 
incomplete for toxicokinetic data in animals by the oral and inhalation routes; acute toxicity 

data by the inhalation route; repeated-dose toxicity data by the inhalation and dermal routes; 

and chronic toxicity/carcinogenicity. Data in humans on most toxicity endpoints are not 

available. 

One area of uncertainty concerns the effects reported in offspring in the prenatal 

developmental toxicity study in rats (Christian et al., 1997; Christian et al., 1999; as cited in EC, 

2008). When severity, dose relationships, and historical ranges were taken into consideration, 

the reductions in pup body weight were not definitively considered by the study authors to be 

treatment-related. Additionally, fetal malformations observed were reported in only three 

fetuses from separate litters. Even though there is uncertainty surrounding these endpoints, 

the study authors, as well as other reliable entities, have concluded a conservative LOAEL for 

developmental toxicity of 500 mg/kg-bw/day from this study. 

Another area of uncertainty concerns the period of dosing in the peri/post-natal reproductive 

toxicity study (Ford and Bottomley, 1997; Jones et al., 1996; as cited in EC, 2008). This study 

was designed to determine effects of HHCB on the neonate when exposed through nursing. 

Exposures in the F1 offspring occurred in utero from GD 14 through lactation; there were no 

exposures to HHCB beginning from weaning in the F1 offspring throughout the F2 generation. 

Dosing of pregnant animals should typically include the entire period of organogenesis, which in 

the case of the rat is GDs 6 to 15. Although dosing in this study began towards the end of 

organogenesis, it was considered to be adequate for screening-level purposes in order to 

characterize developmental toxicity (including endpoints such as pup weight, pup survival, and 

postnatal death) and reproductive toxicity (reproductive performance) as well as a battery of 

developmental neurotoxicity tests. Other reliable entities such as the 2008 EU RAR on HHCB 

(EC, 2008) and the 2009 SIDS Initial Assessment Profile (OECD, 2009) also concluded that this 

study was adequate for assessing hazard for these endpoints. 

Human biomonitoring data indicate that HHCB is present in milk, fat, and blood, but the studies 

in the US are preliminary and of limited value for characterizing exposures. Specifically for the 

US, there are few biomonitoring studies in limited locations with small numbers of samples. 

HHCB has not been measured as part of the National Health and Nutrition Environmental 

Survey (NHANES), so the data are not representative. Exposures in the US are assumed to be 

similar to those in Europe, but it is unknown how manufacture and use of products containing 

HHCB may differ between regions or cultures since use of fragrances, personal care, and 

cleaning conventions may vary. 

A-5 Conclusions of Human Health Assessment 

The available toxicokinetic data indicate that HHCB is poorly absorbed through the skin; 

toxicokinetic data for the oral and inhalation routes are not available. Limited pharmacokinetic 

data have reported HHCB metabolites in the milk of pregnant and lactating rats. Human 

biomonitoring data have reported HHCB in milk, fat, and blood, but the studies are preliminary 
and of limited value for characterizing exposures. The acute toxicity by the oral and dermal 

routes is low. Signs of systemic toxicity following repeated exposures have not been reported. 

None to weak estrogenic and anti-estrogenic activity has been demonstrated. HHCB is not 

mutagenic, corrosive, irritating, or sensitizing; however, minimal eye irritation and possible 

signs of photoirritation have been reported. Carcinogenicity data are not available. 

HHCB was initially selected for review based on a moderate hazard concern for developmental 

toxicity and a high potential for exposure. However, following further review of the 

developmental and reproductive toxicity data, and taking into account several lines of 

evidence, the conclusion is an overall low concern for developmental toxicity. The uncertainty 

surrounding the pup body weights seen at 500 mg/kg-bw/day from the prenatal developmental 

toxicity study, the occurrence of only three fetuses from separate litters exhibiting fetal 

malformations at 500 mg/kg-bw/day from this same study, and the lack of fetal morphological 

changes observed in the pilot studies, even at maternally lethal dosages (up to 1,000 mg/kgbw/day) lead to the conclusion that 500 mg/kg-bw/day is a conservative LOAEL. No effects 

were reported in fetuses in the perinatal toxicity study with lactational exposures at doses 

similar to those reported in human milk and several orders of magnitude higher (20 mg/kgbw/day). Overall, a low concern for developmental toxicity is indicated by the data. 

Additionally, the human health hazards and risks of HHCB have been extensively reviewed and 

assessed by several other reliable entities, most recently the EU, and it has been concluded that 

the overall concern for human health hazards, including that for developmental toxicity, is low. 

The review of human health hazard studies, biomonitoring studies in the US and elsewhere, 

and the EU RAR showed an overall low risk concern for human health, including the risk for 

developmental toxicity, from current use of HHCB. 
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 (Ullmann, 2003; Zviely, 2002). First, a 

cycloaddition reaction of .alpha.-methyl styrene and 2-methyl-2-butene (i.e., amylene) is 

performed under acidic conditions to obtain 1,1,2,3,3-pentamethylindane (1). Second, the 

pentamethylindane (1) is hydroxyalkylated with propylene oxide in a Friedel-Crafts reaction 

using aluminum chloride as a catalyst. Third, the ring closure of the resulting 1,1,2,3,3-

pentamethyl-5-(-hydroxyisopropyl)indane (2) to 1,3,4,6,7,8-hexahydro-4,6,6,7,8,8-

hexamethylcyclopenta-γ-benzopyran (HHCB; Galaxolide) is accomplished with

paraformaldehyde and a lower aliphatic alcohol via the acetal or with paraformaldehyde and a 

carboxylic acid anhydride via the acylate. 

1) Ullmann's Encyclopedia of Industrial Chemistry (2003). 6th ed. Vol 1: Federal Republic of 

Germany: Wiley-VCH Verlag GmbH & Co. 2003 to Present, p. V14 145. 

2) Zviely M; Kirk-Othmer Encyclopedia of Chemical Technology. (2005). NY, NY: John Wiley 

& Sons; Aroma Chemicals. Online Posting Date: Jan 25, 2002. 
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

H3C

CH3

CH3

CH3

H3C

O

CH3 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)- 

N
Table_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-

hexamethylN
Table_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-ethyl1,3,4,6,7,8-hexahydro-4,6,8,8-

tetramethylN

15 78448-49-4 Cyclopenta[g]-2-benzopyran, 8-ethyl1,3,4,6,7,8-hexahydro-4,6,6,8-

tetramethylN

*This compound was evaluated in this assessment.

 
Appendix D CDR DATA FOR HHCB 

The information in Tables D-1 through D-4 is from EPA’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 Informationa

Production volume (aggregated) 3,126,728 lbs

Maximum concentration (at manufacture or import site) > 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 industrial 

manufacturing, processing and use workers 

Not completely reported.

c

Was industrial processing or use information reported? Yes

Was commercial or consumer use information reported? Yes

a

 More detailed information was not publically available on the CDR website because it was considered to be CBI. 

b The total number of sites reported by two submitters is in the range of less than 65 to 123 but one submitter 

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 either 

reported that the number of sites is “not known or reasonably ascertainable” or else did not report any 

information. 

Table_Apx D-2. CDR HHCB Industrial Use Information

Type of Processing Industrial Sector (Based on NAICS) 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 agents
Table_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 reasonably 

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 releases of HHCB to the environment from 

industrial sites of all stages of the HHCB life cycle and estimates of releases from consumer and 

commercial use of end-use products. 

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

• Manufacturing of the chemical substance, 

• Compounding of fragrance oils containing the chemical substance, 

• Blending of fragrance oils containing the chemical substance into commercial and 

consumer products (e.g., perfumes, cosmetics, soaps, detergents), 

• Use of these commercial and consumer products. 

The release estimates are summarized in Table E-1 and 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 estimates 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 prepare fragrance oils or fragrance 

compounds within the United States, based on data collected 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 release 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 follows. 

E-1 Estimated Release from Manufacture and Import 

As discussed in Chapter 2, HHCB is not currently manufactured in the US (IFRA, 2012a). For this 

release assessment, four companies are assumed to import HHCB to the US at five sites (see 

Appendix D, Table D-4). Releases are not expected to result from import activities, but may 

occur at import sites if HHCB is also diluted and compounded onsite after import, as further 

discussed below. 
Table_Apx E-1. Summary of Estimated Environmental Releases

Release Activity Release Media 

Release Factor 

(Amount Released 

per Amount Used, 

Percent) 

Number 

of Sites

Number of 

Release Days 

at Each Site 

(Days/Year) 

Daily Release

(kg/Site-Day)

Combined Annual 

Releases of All Sites

(kg/Year) 

Import and Compounding of Fragrance Oils Unloading 100 percent of HHCB at 75 °C from bulk containers into dilution tank 

Air NA 5-49 2-18a 9.02 × 10-3 0.812-0.884 E

1H

Transferring 65 percent of (diluted) HHCB at 75 °C from dilution tank into totes for distribution to customers or indoor vessels for compounding

Air NA 5-49 26-251a 3.22 × 10-4- 6.45 × 10-4 0.405-0.821 E

1H

Cleaning of bulk containers Water 0.07-0.2b 5-49 2-18a 13.2-37.8c 1,190-3,400d E

R1

Cleaning compounding equipment 

Water 0.07-1b 5-49 250 0.097-13.6 1,190-17,000 E

R1
Table_Apx E-1. Summary of Estimated Environmental Releases

Release Activity Release Media 

Release Factor 

(Amount Released 

per Amount Used, 

Percent) 

Number 

of Sites

Number of 

Release Days 

at Each Site 

(Days/Year) 

Daily Release

(kg/Site-Day)

Combined Annual 

Releases of All Sites

(kg/Year) 

Blending of Fragrance Oils into Products Cleaning of transport containers Water, 

incineration, land 

0.3-3e 950 250 0.021-0.215 5,100-51,000 E

P(M

Cleaning of process vessels Water, 

incineration, land 

2 950 250 0.143 34,000 E

R1

Conveying, mixing, and packaging powder products 

Air, water, incineration, land 

3.83 950 250 0.274 65,110 4

fwpe0(

Use of End Products Disposal of products Water 90.0-93.7 Not 

known 

250 Not known 1,530,000-

1,592,900 

Ear

Total Releases Water 90.1-94.9 NA 2-250 NA 1,532,380-

1,613,300 

Air NA NA 2-251 NA 1.22-1.71 Water, incineration, land (uncertain) 

2.3-5 NA 250 NA 39,100-85,000 

Air, water, incineration, land (uncertain) 

3.83 NA 250 NA 65,110 
Table_Apx E-1. Summary of Estimated Environmental Releases

Release Activity Release Media 

Release Factor 

(Amount Released 

per Amount Used, 

Percent) 

Number 

of Sites

Number of 

Release Days 

at Each Site 

(Days/Year) 

Daily Release

(kg/Site-Day)

Combined Annual 

Releases of All Sites

(kg/Year) 

NOTES: 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-aThe number of unloading days, which is equal to the number of release days, is calculated by dividing the average HHCB throughput 

volume, assuming that one container is unloaded per site per day. Bulk containers are assumed to be 5,000 gallons and totes are ass

The calculation for the cleaning of bulk containers is presented as a sample calculation of this variable: 

 Low-end of range of number of unloading days (or release days) per site: 

HHCB throughput per site = (1,700,000 kg/year ÷ 49 sites) ÷ (1 g/cm3 × kg/1,000 g × cm3/0.0002642 gallons) = 9,166 gallons 

 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 

 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 unit amount used, is equal for this release activity to the residual amount in a transp

equipment. This residual amount is expressed as a fraction of container or equipment volume and has the following range of values: 

(1) container unloading and equipment cleaning by gravity drain: 0.07 percent (central tendency) and 0.2 percent (high-end)(2) equipment cleaning by pumping: 1 percent (conservative). 

cThe calculation of this range of releases per site is presented as a sample calculation of this variable: 

 Low-end of range of amount released per site-day:  5,000 gallons/container-site-day × 0.0007 residual gallons/container gallon × (1 g/cm3 × kg/1,000 g × cm3/0.0002642 gallons

UpperHigh-end of range of amount released per site-day = 5,000 gallons/container-site-day × 0.002 residual gallons/container gallon × (1 g/cm3 × kg/1,000 g × cm3/0.0002642 gallons) 

dThe calculation of this range of combined releases of all sites per year is presented as a sample calculation of this variable: 

 Lower-end of range of amount released per year = 13.2 kg released /site-day × 5 sites × 18 days of release = 1,190 kg/year 

Upper-end of range of amount released per year = 37.8 kg released/site-day × 5 sites × 18 days of release = 3,400 kg/year 

eResidual amount as a fraction of drum volume: (1) containers unloaded by pumping: 3 percent (high-end) and 2.5 percent (central t

unloaded 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 processed is not known. Therefore, the 

number of sites was estimated using two alternative approaches, resulting in an estimated 

range of per-site release from compounding sites. EPA/OPPT expects that imported HHCB may 

be delivered to compounding sites. Compounding sites are classified under North American 

Industry Classification System (NAICS) code 32562, Toilet Preparation Manufacturing, and 

specifically under the Perfume Oil Mixtures and Blends Manufacturing subcategory. This NAICS 

code is not mentioned in the 2006 non-CBI IUR data on HHCB; therefore, EPA/OPPT assumed 

that HHCB compounding may occur only at the five import sites. This approach resulted in a 

high estimate of per-site release. Alternatively, the number of sites was estimated to be equal 

to the number of IFRA member companies. The estimated HHCB volume used to compound 

fragrance oils was collected through surveys of IFRA member companies, which represent the 

vast majority of fragrance volume produced in the US. IFRA North America currently has 49 

member companies (IFRA, 2012d). If each member company owns at least one site in the US, 

then each site may compound HHCB for fragrance oils. Alternately, the 2002 Economic Census 

estimates 33 companies under this NAICS subcategory (USCB, 2004). The 2007 Economic 

Census does not provide a detailed breakdown of NAICS 32562 by subcategory, but shows that 

the total number of companies within NAICS 32562 increased by four percent from 2002 to 

2007. EPA/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 subcategory in 2007.12 Therefore, the number of 

compounding sites is estimated to include 5 to 49 sites. The geographic distributions of all sites 

under NAICS 32562 are shown in Table E-2. 

In summary, compounding could occur at these sites under two scenarios: (1) after import, 

HHCB is diluted and compounded at the five import sites; or (2) HHCB is delivered to and 

compounded at the 49 compounding sites identified in the Census. During compounding, 

EPA/OPPT assumes that imported HHCB is unloaded into large outdoor storage tanks for 

dilution to a 65 percent solution and is subsequently transferred to an indoor tank in the 

compounding facility for further processing. Typically, HHCB is present in the compounded 

fragrance oils at two to four percent (HERA, 2004a). 

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

5.47 × 10-4 mmHg (see Table 2-1). Because HHCB is a highly viscous liquid at concentrations 

≥65 percent, concentrated HHCB may be heated to become less viscous during transfer 

activities, during which HHCB may have sufficient vapor pressure to volatilize (EC, 2008). The 

unloading and transfer are expected to occur at temperatures between 25 and 75 °C (EC, 2008). 

Therefore, there is potential for releases to air due to volatilization from unloading and transfer 

 

12 Because the number of establishments under this subcategory is not known, and most companies under 

NAICS 32562 are small companies with no more than one establishment, the 34 companies are assumed to 

represent 34 establishments (sites). 
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

a250 to 499 employees. 

b500 to 999 employees. 

c100 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 vessels. The environmental release media 

may include water or incineration, depending on the method of cleaning (IFRA, 2012c). 

No quantitative release information was found for the compounding scenarios. Compounding 

releases were assessed using US EPA release models related to cleaning and transfer activities 

(Fehrenbacher and Hummel, 1996; EPA, 1988; 1991; 1992). 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 compounding sites, assuming that one container 

is unloaded per site per day and using a default unloading rate of approximately 200 gallons per 

minute. EPA/OPPT estimated the number of release days per year for equipment cleaning by 

assuming that cleaning occurs after each batch, there is one batch per site per day, and there 

are 250 days per year operation at each site. 

E-3 Estimated Release from Blending of Fragrance Oils 

After compounding, fragrance oils containing HHCB are blended with other materials to 

formulate commercial and consumer products. EPA/OPPT assumed that HHCB may be present 

in the formulated products at a maximum of 0.9 percent (HERA, 2004b). The 2006 non-CBI IUR 

reported a range of 100 to 999 industrial processing and use sites specifically associated with 

HHCB. The estimate of 100 to 999 sites includes both compounding and blending sites. Because 

there are approximately 49 compounding sites, EPA/OPPT estimated that there may be up to 

950 blending sites (999 sites – 49 sites = 950 sites). Blending sites are classified under NAICS 

codes 32562, 325611, and 325612. A geographic distribution of these sites, many of which are 

located in California, New York, and Texas, is provided in Tables E-3 and E-4. Note that the data 

include sites within the NAICS industry sectors, and may include sites that do not specifically 

handle HHCB. 
Table_Apx E-3. Geographic Distribution for Facilities under NAICS 325611 Soap and Other 

Detergent Manufacturing 

State Number of 

Establishments 

Establishments with ≥20 

Employees 

Number of 

Employees 

Texas 72 11 1,111

California 69 11 851

Illinois 43 14 2,123

Ohio 39 15 3,075

New Jersey 34 16 1,787

Pennsylvania 32 8 900

Florida 31 5 —a

Missouri 31 8 1,553

Georgia 28 10 862

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

a250 to 499 employees.

b500 to 999 employees. 

c100 to 249 employees. 

d1,000 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

a1,000 to 2,499 employees.

b100 to 249 employees. 

c500 to 999 employees. 

Source: USCB (2007c). 

During the blending of fragrance oils into commercial and consumer products, releases may 

result from the cleaning of transport containers and process vessels. Additionally, releases may 
occur as a result of conveying, mixing, and packaging of powder products containing HHCB. No 

quantitative release information was found for the scenario of blending of fragrance oils; 

therefore, these releases were estimated using the OECD Emission Scenario Document (ESD) on 

the Blending of Fragrance Oils into Commercial and Consumer Products (OECD, 2010). The ESD 

assumes that fragrance oils may be transported to blending sites in drums. Further, the ESD 

indicates a potential for dust losses if the operation involves the blending of solid commercial 

and consumer products (e.g., powder detergent) (OECD, 2010). Because the volume of HHCB 

formulated into solid and powdered products is not known, EPA/OPPT conservatively assumes 

that the entire volume is formulated into powder products, 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 not 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 Consumer 

Products 

No quantitative release information was found for the use of commercial and consumer 

products containing HHCB. EPA/OPPT assumed that the use of all end-use products containing 

HHCB results in down-the-drain release of HHCB to water. The basis for this assumption is the 

information on the type of end-use products that fragrance oils are used in, as summarized in 

Table 2-7. All of the types of end-use products listed with the possible exception of “other” and 

“fine fragrances” are products whose use is likely to result in down-the-drain release of these 

products, and these products together represent 89% of the use volume of HHCB. It is noted 

that the data in Table 2-7 is neither specific to HHCB nor to the use of HHCB in the US. 

However, a 100% release to water from the use of end-use products has also been assumed in 

other HHCB risk assessments (EC, 2008 and HERA, 2004b). This assumption is deemed to be 

adequate for the purpose of this release assessment as well. These releases, combined with 

upstream losses from compounding and blending, account for 100 percent of the HHCB use 

volume. Wastewater containing detergent and cleaning products is discharged to the sewer 

and routed to industrial and municipal wastewater treatment facilities. The release estimates 

represent the quantity released prior to treatment (e.g., wastewater influent). Because HHCB 

has uses in many different types of products and an accurate estimate of the number of 

commercial sites that may use these products is not available, release calculations for use of 

commercial and consumer products are presented as combined releases from all sites on an 

annual basis. 

 
Appendix F ADDITIONAL STUDIES 

F-1 Endocrine Mechanisms and Molecular Pathways 

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

pathways and are briefly summarized here. 

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

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

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

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

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

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

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

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

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

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

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

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

was detected in a transgenic zebrafish assay, but anti-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 locations (Simmons et al., 2010). A study of the 

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

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

(Schnell et al., 2009). 

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

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

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

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

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

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

after a recovery period in clean seawater (Luckenbach and Epel, 2005).
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 concentration of HHCB in influent is 

dependent on the source of waste received (municipal, commercial, industrial) and the 

population served by the WWTP. The concentrations of fragrance materials in effluent, 

however, are dependent upon the ability of these compounds to be eliminated during 

treatment as well as WWTP size, type, and process. 

In the draft report of the Science Advisory Panel for Chemicals of Emerging Concern in 

California’s Aquatic Ecosystems, the maximum concentration of HHCB in wastewater effluent 

reported from within the state of California was 2.78 μg/L (Anderson et al., 2012). The 

maximum aqueous concentration of HHCB in treated municipal wastewater effluent discharged 

to the ocean was 2.5 μg/L. Details regarding the extraction methods and individual QA/QC 

procedures for the specific studies were not provided in this summary document. 

Quarterly samples of wastewater influent and effluent were taken over a one year period at 

two WWTPs in Texas, and synthetic musk fragrance concentration was measured (Chase et al., 

2012). HHCB was consistently one of the more abundant synthetic musk fragrances present 

throughout quarterly sampling, though seven other musk 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 extraction methods, the influent concentration 

of HHCB was as high as 5.7 μg/L and the effluent concentration 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 quantitation limit was 0.040 μg/L. Where 

synthetic musk fragrances were detected in blank samples, the amount present was below the 

calculated method detection limit. 

A state-wide survey in Oregon of trace metals and organic chemicals in municipal effluent was 

published in 2012 (Hope et al., 2012). Oregon's Senate Bill 737, enacted in 2007, required the 

state's 52 largest municipal WWTPs and water pollution control facilities to collect effluent 

samples in 2010 and analyze them for persistent organic pollutants. These facilities are located 

state-wide and represent a variety of treatment types, service population sizes, geographic 

areas, and flow conditions. HHCB was detected in 2/102 (the reported limit of quantitation was 

10 μg/L) samples with a median value of 12.5 μg/L and a maximum value of 13.5 μg/L. The 

samples were filtered and analyzed using mass spectrometry 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 were determined at two WWTPs, one located 

in a rural area in Kentucky and the other from an urban site in Georgia (Horii et al., 2007). 

These two WWTPs treat primarily domestic and commercial wastewater. Concentrations of 

HHCB in influent from both WWTPs were 3 to 60 times higher than those in effluent. Mean 

concentrations of HHCB in influent of the Kentucky plant (2.499 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 differ significantly between the two sites 

(0.044 and 0.055 μg/L, respectively). Liquid-liquid extraction was performed on the wastewater 

samples and the concentration of HHCB was determined by GC/MS. Procedural blanks analyzed 

with the samples as a check for contamination during analysis did not reveal the presence of the 

target polycyclic musks. The limit of quantification 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 identical treatment processes and serve cities 

of moderate population size (approximately 100,000 people) (Reiner et al., 2007a). Both 

WWTPs receive domestic and commercial discharge, and although one also receives industrial 

discharge, the measured concentration of HHCB did not differ. HHCB was found in wastewater 

samples with influent and effluent concentrations in the ranges of 4.76 to 12.7 and 0.010 to 

0.098 μg/L, respectively. Liquid-liquid extraction was performed on the wastewater 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 the limit of quantitation, 10 ng/L, was set to 

be 10 times the standard deviation found in the blanks. The average recoveries for HHCB were 

85 ± 4.3%. 

WWTP effluent was sampled at two sites over a two-year period: Taylors Falls WWTP at Taylors 

Falls, Minnesota and St. Croix Falls WWTP at St. Croix Falls, Wisconsin (USGS, 2011). The sites 

were chosen because the WWTPs discharge into the St. Croix River upstream from endangered 

mussel populations. Mean concentrations of HHCB over the sampling period were 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 disposable, 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 surface water from the confluence of three 

municipal sewage treatment effluent streams in the state of Nevada, for a period of 7 to 

12 months. HHCB 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 receives only sewage effluent and rarely runoff). 

Twice filtered water was drawn by and passed through a peristaltic or diaphragm pump and 

finally through a cartridge containing a 1:1 polymethyl methacrylate:polystyrene crossed linked 

with 50% divinylbenzene sorbent. 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 STP effluent and 99% for nanopure water. 

Smyth et al. (2008) measured the HHCB concentration in the influent and effluent over a oneyear period at six different WWTPs, located in Ontario Canada, employing four different 

treatment processes. Influent concentrations were as high as 40.3 μg/L and effluent 

concentrations were as high as 3.73 μg/L. Samples were 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 concentrations measured in wastewater influent 

and 1-2 orders of magnitude below concentrations measured 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 produced the lowest effluent concentration, 

although the authors noted that process temperature may have influenced the removal 

efficiency. Three WWTPs that utilize conventional activated sludge processes were studied, 

with only small differences in effluent concentrations observed. Overall, effluent HHCB 

concentrations were an order of magnitude less than the influent HHCB concentrations, as 

summarized in Table G-1. 

A limited number of studies have evaluated the differences between different types of 

wastewater treatment and their ability to remove HHCB. A study by Simonich et al. (2002) of 

six wastewater treatment processes (activated sludge, carousel, oxidation ditch, trickling filter, 

rotating biological contactor, and lagoon) in the US and Europe concluded that overall removal 

(primary and secondary treatment) of fragrance materials ranged from 58.6-99.9%, with the 

lowest removal occurring at carousel plants. The highest 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 aeration, oxidation ditch, lagoon) similarly 

found that lagoon treatment yielded the highest removal (>95%). Both authors suggested that 

the long retention times associated with lagoon treatment 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 removal 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) 

aNot 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, are used in many consumer products, these 

compounds enter WWTPs through down-the-drain practices. The ability of fragrance materials 

to be eliminated during treatment depends on a variety of factors inherent to the WWTP, such 

as WWTP size and type (municipal or industrial), processes utilized to treat waste, and 

populations served (rural versus urban). Consequently, effluents from WWTPs are a primary 

source of fragrance materials into the environment. These compounds are found in surface 

water and groundwater located near wastewater discharge areas, with peak environmental 

concentrations typically occurring near effluent discharge points (Heberer et al., 1999; Peck et 

al., 2006). Although the concentration of HHCB in wastewater and surface water may vary, 

HHCB is continuously present and constitutes a constant exposure source. 

A science advisory panel convened by the Southern California Coastal Water Research Project 

was charged to identify potential sources of chemicals of emerging concern, evaluate their fate 

and effects, and ultimately provide guidance for developing monitoring programs that assess 

those chemicals with the highest potential to cause effects in the state’s receiving waters 

(Anderson et al., 2012). All data were blank censored to ensure that the reported compounds 

were in the sample at the time of collection and not 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 (freshwater) waterway; Scenario 2, a coastal 

embayment that receives both WWTP effluent and stormwater discharge; and Scenario 3, 

offshore ocean discharge of WWTP effluent. The maximum values reported for each of these 

scenarios are recorded in Tables G-1 and G-2. 

Water samples were taken from the North Side water reclamation plant effluent and the North 

Shore Channel of the North Branch of the Chicago River, both located in the Chicago, Illinois 

metropolitan area (Barber et al., 2011). This plant collects sewage from residential (96 percent) 

and commercial/industrial (3.4 percent) sources. Following treatment, the water reclamation 

plant effluent is discharged to the North Shore Channel. There were no statistical differences in 

concentrations of HHCB in these two locations, indicating minimal in-stream dilution or 

transformation. The mean concentration was reported as 1.6 μg/L. Glass fiber filtered samples 

were extracted using polystyrene divinylbenzene solid phase extraction and eluted with 

methylene chloride. Field blanks, laboratory blanks and distilled water matrix spikes were 

analyzed for qualtiy assurance, however, specific details regarding laboratory reporting limits 

for the water samples were not provided. 

HHCB was detected in the surface waters of the Canyon Lakes System and North Fork Brazos 

River near Lubbock, Texas in 2010 (Chase et al., 2012). Reported concentrations in the lake 

ranged from ND to trace amounts, whereas concentrations in the river were appreciably higher, 

ranging from 0.077 to 0.794 μg/L. The low lake concentrations 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 of WWTP effluent. The semi permeable 

membrane device used for sampling was dialyzed into hexane and reduced in volume prior to 
analysis by GC/MS. Recoveries were reported to be consistently over 50%. The method 

detection limit was 0.001 μg/L and the method quantitation limit was 0.005 μg/L. Where 

synthetic musk fragrances were detected in blank samples, the amount present was below the 

calculated method detection limit. 

Peck and Hornbuckle (2004) characterized the concentration of several synthetic musk 

fragrances in western Lake Michigan with water samples that were collected from 1999 to 

2001. A lake-wide annual mass balance analysis showed that WWTP discharge was the major 

source of the synthetic musks, whereas major loss mechanisms were outflow and volatilization. 

The average measured concentration of HHCB over this time period was 4.7 ng/L. The resin and 

glass fiber filters used in water sampling were extracted with acetone/hexane. Separations 

were performed with hexane, dichloromethane/hexane and methanol followed by analysis 

using GC/MS techniques. The average recovery of fluoranthene-d10 was 56±34%. Field and 

laboratory blanks were utilized with the average sample to blank ratio for the dissolved–phase 

listed as 7.9. The limit of detection was not reported. 

Water concentrations were measured in 2006 for HHCB from three locations along the upper 

Hudson River, with mean concentrations for HHCB ranging from 0.00395 to 0.0251 μg/L (Reiner 

and Kannan, 2011). The highest concentration was found near Albany, New York, which 

corresponded to the location with the highest population. 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 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 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. Croix National Scenic Riverway in Minnesota and 

Wisconsin. Samples of treated wastewater effluent from two WWTPs, located in Taylors Falls, 

Minnesota and St. Croix Falls, Wisconsin were collected from 2007 to 2008, with concentrations 

up to 1.33 μg/L; however HHCB in surface water sampled from the St. Croix River at two 

locations in MN was not detected. Semi-permeable membrane 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 HHCB was listed as “0.5e” where e denotes 

variable performance during initial method validation. 

Klecka et al., 2010 in a report prepared to a multi-board work group, summarized studies on 

chemicals of emerging concern that could pose threats 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 regarding detection limits and analytical 

methodologies was not provided as the report provided summary statistics only. 
Concentrations of HHCB were detected in low concentrations in Lake Mead, with a mean value 

of 0.00036 ± 0.0003 μg/L in 2001, although this waterbody is downstream of the confluence of 

three WWTPs in Las Vegas, Nevada (Osemwengie and Gerstenberger, 2004). The low 

concentrations of HHCB, compared to the concentrations found in the municipal sewage 

effluent, may be due in part to the dilution of the discharge of the combined sewage treatment 

plant effluents, as the effluent constitutes only 1.37 percent of the total lake volume. Twice 

filtered water was drawn by and passed through a peristaltic or diaphragm pump and finally 

through a cartridge containing a 1:1 polymethyl methacrylate: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 STP effluent and 99% for nanopure water. 

Passive sampling devices were deployed within two nearshore study zones along the north 

shore of Lake Ontario, Canada near the municipalities of Pickering and Ajax, Ontario, and Port 

Hope, Ontario (Helm et al., 2012). Time-weighted average dissolved water concentrations were 

estimated based on chemical amounts sequestered in the device. Organic wastewater 

chemicals, like HHCB, were greatest in the vicinity of the Pickering-Ajax WWTP, and then 

dramatically declined. Calculated sampling device-based concentrations of HHCB ranged from 

approximately 0.0002 to 0.010 μg/L near Pickering-Ajax and from 0.1 to 0.3 ng/L near Port 

Hope. Recovery of HHCB was >79% as determined by extracting passive sampling devices 

spiked with the target compound. Measured concentrations 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 liquid–liquid extraction. The measured HHCB 

concentration within the bay of Hamilton (highly influenced 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 effects 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 Channel 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 95th%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) 

aNot 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 fragrances are known to be present in wastewater 

effluent and the aquatic environment and have been detected in downstream surface water 

sediment of streams and lakes. Aquatic biota may be subsequently exposed to HHCB from 

suspended sediment intake. 
Chase et al. (2012) measured the concentration of HHCB in the sediment of a playa lake, a noneffluent impacted study site located in proximity to an area treated by a WWTP. The primary 

water sources are rain and urban runoff, as these playas are integral for storm water retention. 

In this study, sampling was performed in the summer and fall timeframes of 2010. The 

measured concentrations of HHCB were 2.13 and 1.43 μg/kg dw in the summer and fall of 2010, 

respectively, as shown in Table G-3. Samples were extracted 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 distribution of organic waste contaminants in the 

Tinkers Creek watershed and two other tributaries to 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 flow in Tinkers Creek and its tributaries 

(sometimes ≥80 percent). HHCB was detected in the streambed sediment in 29 percent of 

upstream sites and 86 percent of downstream sites. The range of values is 20 to 390 μg/kg dw 

with an estimated median value of 60 μg/kg dw. The reported upstream concentrations were 

<60 μg/kg dw, whereas the downstream concentrations ranged from 20 (estimated value) to 

390 μg/kg dw with three of the seven downstream samples having measured values >200 μg/kg 

dw. Sediment samples were extracted using a pressurized 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 and July 2006 at three locations along the 

upper Hudson River: Troy, Albany, and Catskill, New York (Reiner and Kannan, 2011). WWTPs 

in each of these New York cities discharge effluents into the Hudson River. HHCB was detected 

in all sediments sampled. Concentrations of HHCB were 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 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 sediment was 

5 μg/kg. The authors noted that these sediment concentrations were similar to concentrations 

that have been measured in Germany (Fromme et al., 2001; Dsikowitzky et al., 2002). 

Sediment samples from three tidal tributaries of the Chesapeake Bay, US (Magothy, Corsica, 

and Rhode Rivers) were collected in 2007 (Sapozhnikova et al., 2010). The Rhode River is the 

only one that receives effluent from a WWTP. Concentrations of HHCB detected in sediments 

were as high as 9.2 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 average HHCB concentration was 1.1±2.2 

μg/kg dw. The Magothy River is the most urbanized river among the three tributaries studied 

and was reflected in the concentration of HHCB found in the sampling. Thirty-nine total 

samples were collected, but details regarding the number of samples for each location were not 
provided. Samples were extracted with a 1:1 mixture of methylene chloride and acetone prior 

to GC/MS analysis. Reagent blanks and replicate samples were analyzed with each batch of 

samples and blank subtracted as needed. HHCB recoveries 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 were measured in August of 2003 (Peck et 

al., 2006). The concentration in the Lake Erie core sample were blank corrected using the 

average mass of each compound found in four method blanks. Sample extraction was 

performed with dichloromethane and exchanged into hexane. 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 difference may be attributed to multiple factors, 

including distance from principal sources (e.g., wastewater outfalls) and differences in loss 

mechanisms (e.g., photolysis, outflow, or volatilization). The concentration of HHCB over time 

was extrapolated from the collected sediment core samples and the HHCB concentration was 

correlated to US fragrance consumption from 1980 to 2006. The authors concluded that HHCB 

concentration found in the sediment samples reflected the trend of fragrance consumption 

over this time period. 

Table_Apx G-3. Measured Sediment Concentrations at Locations in the 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 20a-<60, 37 average

20a-390, 144 mean

Upstream of 7 WWTPs

Downstream of 7 WWTPs 

USGS (2008); 

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 efficiencies ranging from as low as 29.9 percent 
to as high as 99.7 percent (typical values were 70 to 80 percent) depending upon the type of 

wastewater treatment (Simonich et al., 2002; Deblonde et al., 2011; Smyth et al., 2008). A 

fraction of HHCB is expected to preferentially partition to organic-rich biosolids during 

wastewater treatment based on the high measured log KOW values of 5.3 to 5.9 (see Table 2-1) 

and the estimated KOC values of 4.1 to 4.3 (EPI Suite v4.11, see Chapter 2, Section C, 

Environmental Fate). These solids, produced through the process of wastewater treatment, are 

subsequently disposed of by either landfill or incineration. They may also be utilized for land 

application to fertilize plants or to improve the quality of soil, and may then become available 

for plant and terrestrial biota uptake. 

In a study by Kinney et al. (2006), the presence and concentration of organic waste 

contaminants was determined in solid materials produced during wastewater treatment. Nine 

different biosolid products from municipal WWTPs in seven different states were analyzed. Six 

of these products were intended for commercial, homeowner, and municipal use and are 

available to the public for purchase. Three of the products are used in agriculture. HHCB was 

detected in all nine biosolids. HHCB was extracted from the biosolid using accelerated solvent 

extraction and measured by GC/MS. Each sample was composited from bulk material and 

samples were analyzed in triplicate. All samples had similar treatment (secondary activated 

sludge and chlorine disinfection); however, the preparation and applied form of the solid 

varied. The concentration of HHCB varied, with one site sample measuring as high as 

177,000 μg/kg dw. All other measured HHCB concentrations were below 4,000 μg/kg dw. As 

this high value was more than 40 times higher than the next highest measured concentration, 

and more than four orders of magnitude higher than the lowest reported concentration, the 

authors considered it as an “outlier.” Excluding this value, the reported median value for HHCB 

was 1,461 μg/kg dw. Extractions were performed using 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 quantifications; however HHCB was not 

detected in any of the blanks hence no contamination was indicated. The method detection 

limit was noted as 16.5 μg/kg and the recoveries ranged 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 delineated in this study. 

A second study by Kinney et al. (2008) was performed to examine bioaccumulation in 

earthworms from soil obtained from three Midwest agricultural fields, which had been 

amended with biosolids or swine manure. The biosolid source had a concentration of 

427,000 μg/kg dw, although details regarding this source were not provided. 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 detection limit for 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 weekly schedule for 19 weeks (Buyuksonmez 

and Sekeroglu, 2005). Samples were taken from the return activated sludge line after purging 
the stagnant sludge from the sampling line. The presence of HHCB was determined using 

GC/MS following extraction (sequentially with methylene 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 not provided. 

In a study by DiFrancesco et al. (2004), digested sludge was collected from two sites over a 

period of two years. Anaerobically digested and dewatered sludge was obtained from two 

activated sludge plants in Delaware: Georgetown WWTP and Wilmington WWTP. Background 

concentrations of fragrance materials in anaerobically digested sludge from the municipal 

WWTPs were determined in July 2000 and again in March (Georgetown) and October 

(Wilmington) 2002. Sludge was dewatered using a gravity separator at Georgetown WWTP and 

a belt filter press at Wilmington WWTP. Both WWTPs added polymers to assist in sludge 

dewatering, and the Georgetown facility also used aluminum chloride for phosphorus removal. 

Samples were extracted using accelerated solvent extraction (dichloromethane) and were 

analyzed using GC/MS. The HHCB concentrations ranged from 21,800 to 86,000 μg/kg dw. The 

overall concentrations in this limited sampling were higher in samples collected in 2000 and at 

the Georgetown site. Table G-4 includes a summary of measured concentrations of HHCB in 

biosolids. 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 are 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 discharge into the Grand River, Ontario, Canada 

were used as sampling sites with four sampling events from November 2003 to July 2004. 

Extraction was performed using supercritical fluid (5% acetone in hexane) or microwave 

assisted techniques (acetone and hexane), measurement was performed using GC/MS. Method 

blanks were employed to confirm the absence of contamination. 3 out of 15 blanks showed 

concentrations of HHCB that were 2-3 orders of magnitude below concentrations measured in 

sludge samples. The method detection limit was reported 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 values in Canada (Lee et al., 2003) and the 

United Kingdom (Stevens et al., 2003). 

Yang and Metcalfe (2006) also studied biosolids from WWTPs in Ontario, Canada. Samples were 

collected from the Peterborough WWTP over a one-year period, and the HHCB concentrations 

were measured in the raw, return activated sludge, and digested biosolids. Concentrations of 

HHCB did not vary significantly for any one sampling type over the one-year study. However, 

there was a trend for the synthetic musks to accumulate in the return activated sludge and to 

finally concentrate in the digested biosolids. The annual average concentrations of HHCB were 

3,309.9 μg/kg dw in the return activated sludge and 6,788.4 μg/kg dw in the digested biosolid. 

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 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 samples; the specific value for HHCB was not 

reported. 

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

Location Collection

Year N Concentrationμg/kg dw Preparation/Comments Reference 

WI 5/2003

11/2003 

3

3 

1,400 (average)

1,470 (average) 

Rotary kiln dried Kinney et al. 

(2006) 

CO 5/2003

11/2003 

3

3 

3,700 (average)

2,670 (average) 

Composted + plant material

TX 8/2003

11/2003 

3

3 

1,100 (average)

977 (average)

Composted + yard material

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

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

KS 9/2003

11/2003 

3

3 

2,370 (average)

1,970 (average) 

Air dried and turned

AZ 6/2003

1/2004 

3

3 

1,300 (average)

697 (average) 

biosolid

AZ 6/2003

1/2004 

3

3 

1,100 (average)

767 (average) 

Dewatered biosolid + polymer

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

digestion + polymer 

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

(2008) 

San Diego, CA 2005 —a 7,840 (average) Return activated sludge line Buyuksonmez and 

Sekeroglu (2005) 

Wilmington, 

DE 

7/2000

10/2002 

—a

7 

43,000b

21,800 (± 4,300) 

Anaerobically digested sludge DiFrancesco et al. 

(2004) 

Georgetown, 

DE 

7/2000

3/2002 

—a

3 

86,000b

37,600 (± 4,500) 

Anaerobically digested sludge

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 

Smyth 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 Concentrationμg/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) 

aNot reported. 

bValue was not reported as mean, average, or median. 

Studies summarized in Table G-4 reported analytical reliability using accelerated solvent 

extraction or supercritical fluid extraction followed by GC/MS techniques. Studies of HHCB in 

biosolids are somewhat limited, however, because the studies that detected and measured 

HHCB reported concentrations in the range of <100 to >100,000 μg/kg dw. The measured 

concentration depended on a variety of factors, including WWTP operations such as receiving 

waste stream (municipal or industrial), date of collection, location, population served, water 

volume, treatment type, and preparation methodologies prior to land application. Because 

these factors cannot be independently separated from values within the reported data set, it is 

not possible to assign sensitivity to a single factor. Comparison with values reported for studies 

in Ontario, Canada are within the reported range for those from the US and also similarly show 

differences depending upon location, treatment process, and season. 

G-5 Measured Concentrations in Soil 

Limited data were available that described concentrations 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 wastewater effluent or amendment of soils with 

biosolids resulted in measurable concentrations of HHCB. The high frequency of HHCB 

detection in biosolids and wastewater suggests that they may be an important organic waste 

contaminant source to terrestrial environments, particularly when applied to soil. Table G-5 

summarizes the measured concentrations in HHCB in soil. 
In 2008, Kinney et al. (2008) reported HHCB concentrations in agricultural soil amended with 

biosolids. Three sites were studied. Site 1 (minimally affected site) was a non-irrigated soybean 

field that had not been amended with either human or livestock waste for at least the previous 

seven years. Soil and earthworm samples were collected from this field in June and September 

of 2005. Site 2 (biosolid amended site) was a no-till, non-irrigated soybean field receiving 

biosolid as a fertilizer for the first time. Site 3 (manure-amended site) was a non-irrigated 

cornfield receiving liquid swine manure as an organic fertilizer. While HHCB was not detected at 

site 3, HHCB was detected at site 1 in June 2005 at measurable levels, up to 633 μg/kg dw. 

HHCB concentrations were highest at site 2 and ranged from 1,050 to 2,770 μg/kg dw. 

Extractions were performed with a 70:30 acetonitrile: 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 detection limit for HHCB was reported as 

12.5 μg/kg dry weight. 

In a study by Chase et al. (2012), soil core samples were collected quarterly over a two-year 

period at sites outside the city limits of Lubbock, Texas to determine the amount of sorption 

after land application of treated wastewater and groundwater recharge. The wastewater at this 

site is treated through primary and secondary treatment processes for removal of suspended 

solids and organic carbon through different operating flow streams at the WWTP. Some of the 

treated effluent is released to a secondary effluent reservoir, which is then applied via center 

pivot irrigation at a land application site. Two sites were studied both inside and outside the 

pivot irrigation system; however, small differences in measured HHCB concentrations were 

observed between the two locations. Three sample depths were studied: 0 to 6, 12 to 18, and 

24 to 30 inches below the surface. Over the time period of the study, at locations inside the 

irrigation system, concentrations ranged from <0.33 to 1.98 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 ranged 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 value, concentrations were below 1.98 μg/kg dw 

over the time period of the study. Soil concentrations were found to be lowest during the 

summer months, possibly due to volatilization. Extractions were performed with 1:2 

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. Groundwater was also sampled 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) Reference

Midwest, US

Agricultural fields

2005 Minimally affected ND-633

Kinney et 

al. (2008) Biosolid amended 1,050-2,770

Manure amended ND

Lubbock, TX

Land application 

site 

2009-

2010 

0-6 inches 12-18 inches 24-30 inches

Chase et al.

(2012) Inside pivot irrigation <0.33-1.98 <0.33-1.25 ND-0.90

Outside pivot irrigation <0.33-5.69 <0.33-0.77 <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 fields 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 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 0.2 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, transformation, and leaching), and organic 

matter content on long-term concentrations. Soil studies, although clearly limited, do indicate 

that land application of treated wastewater effluent or biosolids results in detectable quantities 

of HHCB in soil and hence represents a potential route through which HHCB may enter 

terrestrial ecosystems. 

G-6 Measured Concentrations in Biota 

The liver, fillet, and fat of many aquatic organisms, including fish, shrimp, zebra mussels, and 

aquatic mammals and birds, have been sampled from US waters in a variety of locations 

(Table G-6). These data confirm the widespread occurrence of HHCB in aquatic media and 

subsequent exposure to wildlife. 

Six geographical locations in various parts of the US were selected as sampling sites for five 

effluent-dominated rivers receiving discharge from WWTPs located in Chicago, Illinois; Dallas, 

Texas; Orlando, Florida; Phoenix, Arizona; and West Chester, Pennsylvania; and one reference 

site on the Gila River, New Mexico (Ramirez et al., 2009).The reference site was expected to be 

removed from anthropogenic point sources; therefore, no accumulation of HHCB was expected 
or detected in fish collected from this site. A total of 18 to 24 adult fish of the same resident 

species were collected from each sampling location during late summer and fall of 2006. Fish 

sampled at each site were divided into six composites, each containing three or four adult fish 

of similar size. The composites revealed the presence of HHCB at maximum concentrations 

ranging from 300 to 2,100 μg/kg tissue. Lipid determinations were made gravimetrically and 

extractions were performed with 1:1 dichloromethane-hexane. Each analytical batch contained 

one blank, at least one continuing calibration verification 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 concentration. 

Concentrations of polycyclic musks, including HHCB, in fish were collected in 2006 from Troy, 

Albany, and Catskill along the upper Hudson River in the eastern region of the state of New 

York (Reiner and Kannan, 2011). There are 148 WWTPs that discharge treated wastewater into 

the Hudson River, which flows southward from its headwaters in the Adirondack Mountains 

and ultimately empties into the Atlantic Ocean at New York City. The measured levels of HHCB 

varied across fish species, with a range of <1 to 39 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 analyzed for the presence of synthetic musks. 

HHCB was detected in all samples, with concentrations ranging from 48 to 683 μg/kg lipid 

weight (average 199 μg/kg lipid weight) in farmed shrimp and 66 to 762 μg/kg lipid (334 ng/g 

lipid weight) in wild shrimp (HHCB in US wild shrimp max 330 μg/kg lipid weight, n=6). Farmraised shrimp from Indonesia (n = 1) showed the highest 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 significant; however, the presence of HHCB in 

shrimp tissues, both farmed and wild caught, indicates the widespread exposure of aquatic 

biota to HHCB. Lipid content was determined gravimetrically and measured as 

dichloromethane extractible non-volatiles. Reagent blanks and replicate samples were 

analyzed with each batch of samples and blank subtracted as needed. HHCB recoveries were 

71% and recoveries in blank samples spiked with musk 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 summary of human biomonitoring studies). 

Among the wildlife species analyzed, tissue concentrations 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. Sample 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%. Procedural blanks were analyzed with every 

set of six samples to check for laboratory contamination 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 found 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 one 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 laboratory blanks were used to check for 

contamination. The method reporting limit for HHCB was not specified. 

Kinney et al. (2008) measured HHCB concentrations in earthworm tissue obtained from 

amended soils where biosolids or liquid swine manure was applied. These values were 

compared to values reported for worms collected from a soybean field, which had not been 

amended with human or livestock waste for at least the previous seven years. HHCB 

concentrations in worms 30 days post application (3,340 μg/kg ww) were an order of 

magnitude higher than the levels measured after 156 days (131 μg/kg ww). Tissue 

concentrations in worms from the manure amended site were 49 μg/kg after 139 days. 

Extractions were performed with a 70:30 acetonitrile: 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 earthworm 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 

61a

; NDb

3,340c; 131d

NDe; 49f

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 <1-3.2g 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 mean 
Table_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 

<1-3.2h

Smallmouth bass (liver) 2003 3 Effley Falls Reservoir 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 summarized in Appendix A, Section A-2

ND Not detected, method detection level = 12.5 μg/kg 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 concentrations. 

h 83 percent of samples contained detectable concentrations.
G-7 USGS National Water Quality Information System Data

Monitoring data collected by the USGS NWQL and available from the NWIS database up to May 

2012 were obtained for environmental concentrations of HHCB within the US (USGS, 2012). 

The following categories of HHCB data were available from the USGS NWIS: water, filtered 

recoverable; water unfiltered, recoverable; and solids recoverable, dw. These data were 

grouped by medium type and site type, as shown in Table G-8. The Medium and Site Code 

definitions are provided in Table G-9. 

The data compiled herein include values that are reported as less than the USGS LRL; values 

that are between the LRL and the LT-MDL; and values that are below the LT-MDL (Oblinger 

Childress et al., 1999). Similar categories of data were available for both types of water 

samples. However, for unfiltered water from effluent/stream, surface water/outfall, and 

groundwater/well sites, significantly fewer than 10 data points were available, and were thus 

not incorporated into the summary plots. 

For this assessment, data contained in three NWIS parameter codes (i.e., 62075, 62823, and 

63209) for HHCB were included. The LRL for water sampling was 0.5 μg/L for sampling dated 

7/16/2001 to 9/30/2009. The LRL was updated to 0.05 μg/L for samples dating from 10/1/2009 

to the present based on a re-evaluation of the LT-MDL by the USGS. Interim reporting levels 

were recorded for data collected from solid samples, based on USGS practices for data 

interpretation. For monitoring data sets where the geometric standard deviation was <3.0, 

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

square root of two, as per the EPA OPPT guidance document (EPA, 1994). Likewise, where the 

geometric standard deviation was >3.0, values recorded as “less than LRL” or “estimated” were 

replaced by the LRL divided by two (EPA, 1994). These values are summarized in Table G-7 

below. It should be noted, therefore, that resulting low-end values are biased toward the LRL. 

This practice presents a conservative low-end value, which protects against false negative 

values. As such, values do not necessarily represent quantitative measured concentrations. 

Table_Apx G-7. Summary of Substituted Values (μg/L) for Water Samples

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 data 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, variously dated). Data reporting procedures were 

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

For each of the data groupings (Table G-8), box plots were generated with the calculated mean, 

median, 1st (Q1) and 3rd (Q3) quartile, and 5th and 95th percentile values as shown in Figures G-1 

to G-3. A summary of these calculated values is presented in Table G-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 95thPercentile 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 reporting level 
Table_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 facility or wastewater treatment 

plant, or from combined sources, such as combined-sewer overflows or tile 

drainage systems. 

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

not include soil moisture or interstitial water. 

Bottom material A mixture of mineral and organic matter that compose the top bed deposits 

(usually the first few inches) underlying a body of water. 

Site Codes 

Outfall A site where water or wastewater is returned to a surface-water body (e.g., the 

point where wastewater is returned to a stream). Typically, the discharge end of 

an effluent pipe. 

Lake, reservoir, 

impoundment 

An inland body of standing fresh or saline water that is generally too deep to 

permit submerged aquatic vegetation to take root across the entire body (Cf. 

wetland). This site type includes an expanded part of a river, a reservoir behind a 

dam, and a natural or excavated depression containing a water body without 

surface-water inlet and/or outlet. 

Well A hole or shaft constructed in the earth intended to be used to locate, sample, or 

develop groundwater, oil, gas, or some other subsurface material. The diameter 

of a well is typically much smaller than the depth. Wells are also used to 

artificially recharge groundwater or to pressurize oil and gas production zones. 

Additional information about specific kinds of wells should be recorded under the 

secondary site types or the Use of Site field. Underground waste-disposal wells 

should be classified as waste-injection wells. 

Stream A body of running water moving under gravity flow in a defined channel. The 

channel may be entirely natural, or altered by engineering practices through 

straightening, dredging, and (or) lining. An entirely artificial channel should be 

qualified with the "canal" or "ditch" secondary site type. 

Figure G-1 shows box plots for filtered water from three different mediums and associated 

sites. HHCB concentrations are highest from effluent at outfall or stream sites and from surface 

water at outfall sites with higher values at outfall sites, although the range of values varied 

from ND to >3 μg/L. The mean effluent concentrations were comparable, but the range of 

concentrations was greater at outfall sites. This may be due to the limited data that were 

available for effluent/stream sites, such that data as summarized here encompasses only 10 

data points collected from four individual sites. Mean values for filtered water were <1.5 μg/L 

for all medium and site types. 

Groundwater concentrations at well sites and surface water concentrations at streams and 

lakes, reservoirs, and impoundment sites ranged from ND to <0.5 μg/L. For effluent/outfall 
filtered/non-filtered samples, the HHCB concentrations were higher in unfiltered samples 

(Figure G-2). This may be attributed to the presence of HHCB in suspended solids. Again, 

significantly lower concentrations were consistently recorded in filtered surface water samples 

collected from streams and lakes, reservoirs, and impoundment sites—areas that are 

unaffected or nominally affected by wastewater treatment outfall. 

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

Filtered 
Figure 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 Solids 
The majority of the data available from the USGS NWIS data set for solids was from bottom 

material. Bottom material data were predominantly divided between two types of sites: 

lake/reservoir/impoundment and streams (Figure G-3). The 95th percentile concentration range 

at lake/reservoir/impoundment sites was 213 μg/kg, and the mean concentration was 

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

concentration was 68 μg/kg. The range of data was similar for both site types, with the Q3 

value for lake/reservoir/impoundment sites slightly higher than that observed for stream sites, 

though it is unclear what this may be attributed to.