United States                      EPA Document # 740‐Z1‐4001

Environmental Protection        August 2014

Agency                                                                             Office of Chemical Safety and

                                                                                      Pollution Prevention    

______________________________________________________________________________

TSCA Work Plan Chemical Risk Assessment

Antimony Trioxide

CASRN:  1309‐64‐4

August 2014

   
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TABLE OF CONTENTS

TABLE OF CONTENTS ............................................................................................................... 2

LIST OF TABLES ........................................................................................................................ 4

LIST OF APPENDIX TABLES ....................................................................................................... 4

LIST OF FIGURES ...................................................................................................................... 5

AUTHORS / CONTRIBUTORS / ACKNOWLEDGEMENTS / REVIEWERS ....................................... 6

GLOSSARY OF TERMS AND ABBREVIATIONS............................................................................ 8

EXECUTIVE SUMMARY........................................................................................................... 10

1 BACKGROUND AND SCOPE............................................................................................. 12

1.1 INTRODUCTION..........................................................................................................................................12

1.2 PROBLEM FORMULATION..........................................................................................................................13

1.2.1 ATO Uses ................................................................................................................................................13

1.2.2 Regulation of ATO ..................................................................................................................................14

1.2.3 Human Exposure Potential.....................................................................................................................14

1.2.4 Ecological Exposure Potential.................................................................................................................15

1.2.5 Ecological Hazard Assessment ...............................................................................................................15

1.3 CONCEPTUAL MODEL FOR ENVIRONMENTAL ASSESSMENT OF ATO.........................................................16

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

2.3 ENVIRONMENTAL FATE..............................................................................................................................23

2.3.1 Environmental Persistence....................................................................................................................23

2.3.1.1 Fate in Water.................................................................................................................................................... 23

2.3.1.2 Fate in Sediment............................................................................................................................................... 24

2.3.1.3 Fate in Soil ........................................................................................................................................................ 25

2.3.1.4 Fate in Air ......................................................................................................................................................... 26

2.3.2 Environmental Bioavailability, Transformation and Bioaccumulation..............................................27

2.3.2.1 Bioavailability ................................................................................................................................................... 27

2.3.2.2 Transformation................................................................................................................................................. 29

2.3.2.3 Bioaccumulation............................................................................................................................................... 30

2.3.3 Conclusion of Environmental Fate ........................................................................................................31

3 ENVIRONMENTAL ASSESSMENT ..................................................................................... 32

3.1 INTRODUCTION..........................................................................................................................................32

3.1.1 Environmental Exposure Assessment.....................................................................................................32

3.1.2 Toxics Release Inventory.........................................................................................................................32

3.1.3 Measured Concentrations in the Environment.....................................................................................36

3.1.3.1 Compilation of Environmental Monitoring Data from USGS‐NWIS Database.................................................. 36

3.1.3.2 Compilation of Environmental Monitoring Data from the STORET Database .................................................. 37

3.1.3.3 Location of Monitoring Sites Relative to TRI Facilities...................................................................................... 39
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3.1.4 Estimated Environmental Releases.......................................................................................................39

3.1.4.1 EFAST Modeling................................................................................................................................................ 39

3.1.4.2 Criteria for Selection of TRI Facilities................................................................................................................ 40

3.1.4.3 Release Scenarios used for Exposure Characterization.................................................................................... 40

3.2 ECOLOGICAL HAZARD ASSESSMENT...........................................................................................................42

3.2.1 Acute Toxicity to Aquatic Organisms....................................................................................................43

3.2.2 Chronic Toxicity to Aquatic Organisms.................................................................................................45

3.2.3 Toxicity to Sediment‐Dwelling Organisms..............................................................................................47

3.2.4 Chronic Toxicity to Terrestrial Organisms...............................................................................................48

3.2.5 Conclusions of Environmental Hazard Assessment..............................................................................50

3.3 ENVIRONMENTAL RISK CHARACTERIZATION .............................................................................................50

3.3.1 Summary of TRI Modeling Results........................................................................................................51

3.3.2 Summary of Environmental Monitoring Data......................................................................................54

3.3.3 Conclusions of Risk Characterization....................................................................................................54

3.3.4 Uncertainty and Data Limitations........................................................................................................55

3.3.4.1 Limitations of the Exposure and Release Data ................................................................................................. 55

3.3.4.2 E‐FAST2 Model Uncertainties........................................................................................................................... 56

3.3.4.3 Limitations of the Hazard Data......................................................................................................................... 56

3.4 CONCLUSIONS ............................................................................................................................................57

REFERENCES .......................................................................................................................... 58

APPENDICES .......................................................................................................................... 70

Appendix A HUMAN HEALTH HAZARD SUMMARY .............................................................. 71

A‐1 TOXICOKINETICS .......................................................................................................................................71

A‐2 ACUTE TOXICITY........................................................................................................................................72

A‐2‐1 Irritation ...............................................................................................................................................72

A‐3 SUBCHRONIC TOXICITY.............................................................................................................................72

A‐4 REPRODUCTIVE TOXICITY .........................................................................................................................73

A‐5 DEVELOPMENTAL TOXICITY......................................................................................................................73

A‐6 MUTAGENICITY.........................................................................................................................................74

A‐7 CARCINOGENICITY ....................................................................................................................................74

A‐8 SYNOPSIS ..................................................................................................................................................75

A‐9 PREVIOUS ASSESSMENTS..........................................................................................................................76

A‐9‐1 Environment Canada/Health Canada: Screening Assessment for the Challenge – ATO (EC/HC,

2010) 76

A‐9‐2 EU RAR – Diantimony Trioxide (CASRN 1309‐64‐4; EC, 2008)...........................................................77

A‐9‐3 OECD Assessment for Antimony Trioxide (OECD, 2008)......................................................................77

A‐9‐3‐1 Environment................................................................................................................................................... 77

A‐9‐3‐2 Human	Health ................................................................................................................................................ 77

Appendix B EFAST2 MODEL INPUTS USED WITH TRI DATA.................................................. 78

Appendix C SUMMARY OF USGS AND STORET DATA .......................................................... 85
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LIST OF TABLES

Table 2‐1. Physical‐Chemical Properties....................................................................................... 21

Table 2‐2. US Manufacturers and Importers................................................................................ 22

Table 2‐3 Major End‐Uses of Antimony in the US in 2011 ........................................................... 22

Table 2‐4. Antimony Bioconcentration in Terrestrial Plants ........................................................ 31

Table 2‐5. Bioconcentration and Bioaccumulation of Antimony in Aquatic Species................... 31

Table 3‐1 Summary of 2010 TRI Releases of Antimony Compounds ........................................... 34

Table 3‐2 Summary of E‐FAST2 Model Results for Facilities Reporting TRI Releases, with

Assumptions for Daily Release Amountsa..................................................................................... 41

Table 3‐3 Aquatic Toxicity Data for Sb2O3 and SbCl3 – Acute Toxicity.......................................... 45

Table 3‐4 Aquatic Toxicity Data for Sb2O3 and SbCl3 – Chronic Toxicity....................................... 46

Table 3‐5. Chronic Toxicity Data for Sediment‐Dwelling Invertebrates....................................... 48

Table 3‐6. Chronic Toxicity Data for Soil‐dwelling Invertebrates................................................. 49

Table 3‐7. Concentrations of Concern for Environmental Toxicity .............................................. 50

Table 3‐8. Summary of E‐FAST2 Results for Facilities Reporting TRI Water Releases.................. 52

LIST OF APPENDIX TABLES

Table_Apx C‐1 Summary of All USGS‐NWIS and STORET Data for Exceedances of the

Concentrations of Concern per Media Type................................................................................. 85
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LIST OF FIGURES

Figure 1‐1 Conceptual Model ....................................................................................................... 18

Figure 2‐1 Representative Chemical Structure of Senarmontite (CASRN 12412‐52‐1)................ 20

Figure 2‐2 Representative Chemical Structure of Valentinite (CASRN 1317‐98‐2)...................... 20

Figure 3‐1 Geographic Locations of TRI Facilities and Environmental Monitoring Data.............. 38

   
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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 Work Plan Chemical Risk Assessment for Antimony Trioxide (ATO) 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 US EPA.   

EPA Assessment Team

   

Lead:   Sharon Oxendine, OPPT/RAD

Team Members:

Kay Austin, OPPT/RAD

Christina Cinalli, OPPT/RAD

Rick Fehir, OPPT/Chemistry, Economics & Sustainable Strategies Division (CESSD)

Kirsten Hesla, OPPT/CCD

Jay Jon, OPPT/RAD

Wen‐Hsiung Lee, OPPT/RAD   

Timothy Lehman, OPPT/CESSD

Sara Pollack, OPPT/RAD

Emma Lavoie, OPPT/ CESSD

Management Lead:  Stanley Barone, OPPT/RAD

   

Portions of this document were developed with support from:

Contractor name     (contract number)    Task      

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

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

Versar        (EP‐W‐10‐005)    Modeling, Monitoring

SRC, Inc.       (EP‐W‐09‐027)    Environmental Fate
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External Peer Review

EPA/OPPT released a peer review plan in August of 2012 and draft risk assessment and charge

questions for peer review and 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 ATO.  As an

influential scientific product, the draft risk assessment was peer reviewed in accordance with

US EPA’s peer review guidance. The peer review panel performed its functions by web

conference and teleconference between November 13, 2013 and January 6, 2014.  The panel

consisted of the following individuals:

Dr. Anne Fairbrother (Chair)  

Exponent

Dr. Bruce Hope

CH2M HILL

Dr. David Ostrach

Ostrach Consulting

Dr. Larry Kapustka

Independent Consultant

   

     

Dr. Wayne Landis

Huxley College of the Environment

Dr, Erik Smolders  

Catholic University

Dr. Maria Sepulveda

Purdue University

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.
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GLOSSARY OF TERMS AND ABBREVIATIONS

ATO    Antimony Trioxide

BAF    Bioaccumulation Factor

BCF    Bioconcentration Factor

°C    Degree Celsius

CASRN   Chemical Abstract Service Registry Number

CBI    Confidential Business Information

cm3

    Centimeters Cubed  

COC    Concentration of Concern

d    Day

DMR    Discharge Monitoring Report  

dw    Dry Weight

EC50    Half Maximal Effective Concentration

EC    European Commission

E‐FAST2 Exposure and Fate Assessment Screening Tool, version 2

EPA    Environmental Protection Agency

EU RAR   European Union Risk Assessment Report

MATC    Maximum Allowable Toxicant Concentration

IARC    International Agency for Research on Cancer

IUR    Inventory Update Rule  

kg    Kilogram  

kg/yr    Kilogram(s) per year

lb    Pound

LC50    Half Maximal Lethal Concentration  

LOEC    Lowest‐Observed‐Effect Concentration

mg    Milligram

mg/kg    Milligram(s) per kilogram  

mg/kg‐bw Milligram(s) per kilogram body weight

mg/L    Milligram(s) per liter

mg/m3   Milligram(s) per cubic meter

μg/kg    Microgram(s) per kilogram

μg/L    Microgram(s) per liter  

MMAD   Mass Median Aerodynamic Diameter

NAICS    North American Industry Classification System

NHANES National Health and Nutrition Examination Survey

NOAEC   No‐Observed‐Adverse‐Effect Concentration  

NOEC    No‐Observed‐Effect Concentration

NPDES   National Pollution Discharge Elimination System

NWIS    National Water Information System  

OCSPP   Office of Chemical Safety and Pollution Prevention

OECD    Organization for Economic Co‐operation and Development

OPPT    Office of Pollution Prevention and Toxics  

OSHA    Occupational Safety and Health Administration
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PDM    Probabilistic Dilution Model  

PET    Polyethylene Terephthalate

PNEC    Predicted No‐Effect Concentration

PM    Particulate Matter  

ppb    Parts Per Billion

PVC    Polyvinyl Chloride

REACH   Registration, Evaluation, Authorization and Restriction of Chemicals

Sb    Antimony

Sb3+    Trivalent antimony

Sb5+    Pentavalent antimony

Sb(OH)3 Antimonite  

Sb(OH)6

‐ Antimonate

SbCl3    Antimony trichloride

Sb2O3    Antimony trioxide

Sb2O5    Antimony pentoxide

STORET STOrage and RETrieval Data Warehouse  

TRI    Toxics Release Inventory

TSCA    Toxic Substances Control Act

US    United States

USGS    United States Geological Survey

WWTP   Waste Water Treatment Plant

WHO    World Health Organization

ww    Wet Weight

7Q10 Refers to seven consecutive days of lowest streamflow over a 10‐year period

based on USGS.

   
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EXECUTIVE SUMMARY

As a part of a 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. The category, “Antimony and Related

Compounds” was assessed as part of the work plan.  

ATO is the most widely used antimony compound, accounting for roughly 80 percent of primary

(i.e., mined) antimony consumption in the US (USGS, 2011). ATO is mainly used as a flame

retardant synergist for halogenated flame retardants used in textiles and plastics (Freedman et

al., 2003); approximately 25 million pounds (lbs) of ATO were consumed for this end‐use

application in 2011 (USGS, 2011). Available US Geological Survey (USGS) data indicated no

operational antimony mines or primary antimony production in the US in 2011, except for one

company in Montana that produced antimony metal and oxide from a foreign feedstock (USGS,

2012a).  

Data from EPA’s Toxics Release Inventory (TRI) for the 2010 reporting cycle indicated that

approximately 970,000 lbs of antimony compounds were released to land and roughly 15,700

lbs were released to water by industries that use or produce ATO‐containing flame retardants

(TRI, 2010). On‐site air releases totaled approximately 12,000 lbs, of which roughly two‐thirds

were from stack releases. A significant portion of the remaining fugitive air releases are likely

related to the combustion of fossil fuels; however, the precise quantity cannot be determined

from TRI data.

This assessment focused on ecological risks from the use of ATO as a synergist in halogenated

flame retardants. Based on a review of the available data regarding antimony concentrations in

food and environmental media, and biomonitoring data obtained from the latest National

Health and Nutrition Examination Survey (NHANES), general population exposure to antimony

is expected to be low (CDC, 2009; 2012). Because food and water are the primary sources of

general population exposure, and the less toxic (i.e., pentavalent) form of antimony

predominates in these media, significant human health risks are not anticipated. This

conclusion is supported by recent risk assessments completed for ATO in Canada and Europe.

The US National Toxicology Program is currently conducting a two‐year carcinogenicity study on

ATO; EPA will review these results when they become available.  

The remaining area of potential concern, for effects on ecological organisms, was selected as

the focus of the current assessment. This decision was based on the Agency’s review of

available ecotoxicity information and reported industrial releases of Antimony and Related

Compounds in association with the manufacture and use of ATO‐containing flame retardants.  

                                                       1

 http://www.epa.gov/oppt/existingchemicals/pubs/workplans.html 
Page 11 of 87

To assess ecological risks, EPA/OPPT derived hazard benchmarks (i.e., concentrations of

concern; COC) for water‐ and sediment‐dwelling organisms, which were compared to measured

or predicted antimony concentrations in these media. Media concentrations were obtained

from two independent environmental monitoring datasets, the US Geological Survey‐National

Water Information System (water samples from 1972 to 2012; sediment samples from 1987 to

2011; and soil samples from 2005 to 2010; USGS, 2012b) and EPA’s Storage and Retrieval

(STORET) Data Warehouse (water and sediment samples from 2002 to 2010; EPA, 2012c). In

addition, site‐specific modeling was conducted for selected TRI facilities (i.e., those with

identified use or production of ATO‐containing flame retardants) using EPA’s Exposure and Fate

Assessment Screening Tool, version 2 (E‐FAST2; EPA, 2010a) to estimate surface water

concentrations. The measured and modeled concentrations used in this assessment are

assumed to represent reasonable estimates of environmental exposures.

To evaluate the potential for adverse ecological effects, antimony concentrations in the

ambient environment were compared to the hazard benchmarks (i.e., concentrations of

concern; COC) identified to be protective of aquatic and benthic organisms. Under this

paradigm, ecological risks were not found in association with ATO use as a flame retardant

synergist. Using the E‐FAST2 model, only one exceedance of the COC was predicted following

acute exposure to aquatic organisms, and the available monitoring data showed no

exceedances of the COC for aquatic, benthic or soil‐dwelling organisms within the last three

years (2009 to 2012).  The available hazard and exposure information was inadequate to

support a quantitative evaluation of risks to terrestrial organisms.

This assessment provides insights regarding ecological risks that may be expected in association

with ATO use as a synergist in halogenated flame retardants. EPA/OPPT identified key sources

of uncertainty related to limitations in the available hazard and exposure information, and

conservative assumptions incorporated in the E‐FAST2 model estimates of antimony surface

water concentrations. Although these uncertainties may limit data interpretation, EPA/OPPT

has high confidence in its minimal risk finding based on the following considerations:   

• The use of release data for antimony compounds and conservative assumptions

incorporated in the EFAST‐2 model estimates likely overestimate ATO exposure potential.  

• Environmental monitoring data obtained during the last three years show no exceedances

of the hazard benchmarks identified to be protective of sensitive ecological species.

   
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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). After gathering input from stakeholders, EPA

developed criteria used for identifying chemicals for further assessment. 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 over the next several years. In the prioritization process, ATO was identified for

assessment based on: (1) high potential for exposure due to high production volume and

widespread use in consumer products; (2) expected releases to aquatic, benthic, and terrestrial

environments; (3) high persistence and low‐to‐moderate bioaccumulation potential; (4)

biomonitoring data indicating its occurrence in water, air and soil; and (5) adverse effects

reported in humans and laboratory animals.   

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 ATO. The information presented in the risk assessment may be of assistance to

other Federal, State, and local agencies as well as members of the general public who are

interested in the risks of ATO. EPA/OPPT relied on Agency guidance to develop this the risk

assessment (US EPA, 1998 and US EPA 2014a). EPA followed the risk assessment paradigm

beginning with planning and scoping and problem formulation.  

The initial step in EPA’s risk assessment development process, which is distinct from the initial

prioritization exercise, is scoping and problem formulation. During this step 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, a more robust review

of the factors that influence initial prioritization may result in refinement – either

addition/expansion or removal/contraction – of specific hazard or exposure concerns previously

identified using the prioritization methodology.

ATO is the most widely used antimony compound, accounting for roughly 80 percent of primary

(i.e., mined) antimony consumption in the US (USGS, 2011). ATO is mainly used as a flame
Page 13 of 87

retardant synergist for halogenated flame retardants used in textiles and plastics (Freedman et

al., 2003); approximately 25 million pounds (lbs) of ATO were consumed for this end‐use

application in 2011 (USGS, 2011). Available US Geological Survey (USGS) data indicated no

operational antimony mines or primary antimony production in the US in 2011, except for one

company in Montana that produced antimony metal and oxide from a foreign feedstock (USGS,

2012a).  

Antimony compounds are subject to a number of regulations for various environmental media.

Examples of programs where EPA has identified priority actions for this chemical category

include:

 Listed as a Hazardous Air Pollutant (HAP) under the Clean Air Act 112(b)  

 National Primary Drinking Water Regulation Maximum Contaminant Level established

under the Safe Drinking Water Act (SDWA)

 Designated a priority pollutant under the Clean Water Act; effluent limitation guidelines

established under CWA Section 304

 Annual reporting of chemical releases to the Toxics Release Inventory (TRI) under

Section 313 of The Emergency Planning and Community Right‐to‐Know Act (EPCRA).  

 Hazardous Waste designation and land disposal restrictions established under the

Resource Conservation and Recovery Act (RCRA) and The Comprehensive Environmental

Response, Compensation, and Liability Act (CERCLA)

 Listing on the TSCA Inventory and reporting requirements established under the Toxic

Substances Control Act (TSCA)  

1.2 PROBLEM FORMULATION		

The objective of OPPT’s TSCA Work Plan Chemicals assessments is to use readily available data

to evaluate the human health and/or environmental risks associated with chemical exposures

incurred as a result of TSCA‐related uses. EPA/OPPT approached problem formulation by

reviewing other recently performed assessments and searching published literature from 1978

to 2012 using EPA’s ECOTOX database to identify additional information which was

subsequently retrieved and evaluated for data adequacy. The literature review included ATO

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 International Antimony Association, EPA’s publicly available Inventory Update

Reporting (IUR), Chemical Data Reporting (CDR) and Storage and Retrieval (STORET) databases,

and the US Geological Survey National Water Information Service (NWIS). Data was acceptable

if it met standard adequacy criteria which varies according to the type of information (data

adequacy criteria are described in later sections of this report).
Page 14 of 87

1.2.1 ATO Uses

Global antimony demand is largely driven by the production of antimony compounds used in

various consumer products such as medicines, polymers, resins, and textiles. ATO is considered

to be the most economically relevant and environmentally prevalent antimony compound,

accounting for approximately 80 percent of global antimony consumption; a smaller portion

(roughly 20 percent) of global antimony consumption is associated with metallic uses (Roskill,

2011). Based on this market information, ATO was selected as the focus of the current

assessment.  

ATO is primarily used as a flame retardant synergist. ATO is also used as a condensation catalyst

for the manufacture of polyethylene terephthalate (PET), a large‐volume commodity plastic

used in bottles, films, and various food contact applications. This use scenario has been

evaluated for safety and is regulated by the US Food and Drug Administration and other

international authorities (EC/HC, 2010; EC, 2008; European Food Safety Authority, 2004);

therefore, ATO use in PET plastics will not be further evaluated here.  

1.2.2 Regulation of ATO

Antimony compounds are subject to a number of regulations for various environmental media.

Examples of programs where EPA has identified priority actions for this category include:

 Listed as a Hazardous Air Pollutant (HAP) under the Clean Air Act 112(b)  

 National Primary Drinking Water Regulation Maximum Contaminant Level established

under the Safe Drinking Water Act (SDWA)

 Designated a priority pollutant under the Clean Water Act; effluent limitation guidelines

established under CWA Section 304

 Annual reporting of chemical releases to the Toxics Release Inventory (TRI) under

Section 313 of The Emergency Planning and Community Right‐to‐Know Act (EPCRA).  

 Hazardous Waste designation and land disposal restrictions established under the

Resource Conservation and Recovery Act (RCRA) and The Comprehensive Environmental

Response, Compensation, and Liability Act (CERCLA)

 Listing on the TSCA Inventory and reporting requirements established under the Toxic

Substances Control Act (TSCA)  

1.2.3 Human Exposure Potential		

Based on a review of the available data regarding antimony concentrations in food and

environmental media, and biomonitoring data obtained from the latest National Health and

Nutrition Examination Survey (NHANES), general population exposure to antimony is expected

to be low (CDC, 2009, 2012). Because food and water are considered to be the primary sources

of general population exposure, and the less toxic (i.e., pentavalent) form of antimony generally

predominates in these media, significant human health risks are not expected at ambient levels
Page 15 of 87

to which individuals may be exposed in the environment. This conclusion is supported by

previous ATO risk assessments conducted in Canada and Europe (EC/HC, 2010; EC, 2008).

Additional information regarding the human health effects that may be associated with ATO

exposure can be found in Appendix A. The US National Toxicology Program is currently

conducting a two‐year carcinogenicity study on ATO; EPA will review these results when they

become available.

1.2.4 Ecological Exposure Potential

Industry reported water releases associated with the production or use of ATO as a synergist in

halogenated flame retardants are possible. EPA’s 2010 TRI data can be sorted by industry

codes, using the North American Industry Classification System (NAICS) codes, to identify a

subset of facilities that indicate production, processing or use of ATO‐containing flame

retardants. Because ATO is not specifically listed on the TRI, releases reported under the

broader category of antimony compounds were used as a surrogate to evaluate the potential

for aquatic exposures. Many different types of antimony compounds, and sources for their

environmental release are known, so use of this data is therefore expected to provide a

conservative estimate of ATO exposure potential.  

Queries of the 2010 TRI were used to obtain information on antimony water releases for the 36

TRI facilities identified under the scope of this assessment (i.e., those indicating production,

processing or use of ATO‐containing flame retardants). Water releases reported by the

identified TRI facilities could be modeled using EPA’s Exposure and Fate Assessment Screening

Tool version 2 (E‐FAST2) modeling program to estimate antimony surface water concentrations.  

Environmental monitoring information obtained from the USGS National Water Information

System (NWIS; USGS, 2012b) and EPA STORET (EPA, 2012c) databases include antimony

concentrations in the ambient environment, which can also be used to evaluate environmental

exposures to antimony. Data collection specifically centered on those states in which the

selected TRI facilities are located may provide an approach to linking environmental antimony

concentrations to ATO use. This USGS dataset is considered robust based on the presumption

that consistent internal methodologies were utilized to derive monitoring information of

sufficient quality for exposure characterization, as it was 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, 1999).   

1.2.5 Ecological Hazard Assessment		

For hazard identification, key ecological hazard reports were obtained for critical review and

evaluation of data adequacy. Details regarding the data adequacy criteria used for the

ecological hazard assessment are included in Chapter 3.  
Page 16 of 87

Due to ATO data limitations, antimony trichloride and antimony sulfate were considered as

supporting chemicals to characterize hazards to aquatic and sediment‐dwelling organisms.

Antimony trichloride tends to dissolve more readily than ATO and is therefore expected to

provide a conservative estimate of ecological hazard potential.  

Risks to soil/terrestrial organisms were not evaluated due to deficiencies in the available hazard

and exposure data; however, ATO is not expected to present a major concern for soil‐dwelling

organisms, as it exhibits limited bioaccessibility in this environmental compartment. During

problem formulation, EPA/OPPT summarized the following information, which may be used in a

weight‐of‐evidence approach to support this assumption:  

- There is considerable evidence for the natural attenuation of antimony in soils over

time. Field studies of weathered or ‘aged’ soils show significant decreases in

bioaccessibility when compared to ‘antimony spiked’ soils typically used for laboratory

studies. This phenomenon, and the use of more soluble antimony compounds for

laboratory testing, may overestimate potential toxicity to soil‐dwelling organisms.

- The limited bioaccessibility of antimony in soil may be explained by its association with

organic or inorganic soil constituents. Soil conditions that promote precipitation or

adsorption are expected to decrease bioaccessibility to soil‐dwelling organisms.

- Sequential extraction procedures have shown that antimony often occurs in the

‘residual’ or ‘refractory’ soil fraction. Largely on the basis of such results, it has been

established, and is generally accepted, that antimony is essentially immobile in soils.

- Because antimony exhibits limited mobility, and low solubility in soil (i.e., limited

bioaccessibility), the amounts that may be released to this environmental compartment

are not expected to present significant concerns for soil‐dwelling organisms.  

For the above reasons EPA/OPPT did not conduct a quantitative evaluation of risks to the

terrestrial compartment.    

1.3 CONCEPTUAL MODEL FOR ENVIRONMENTAL ASSESSMENT

OF ATO

The Conceptual Model depicted in Figure 1‐1 identifies environmental releases and critical

exposure pathways associated with ATO use as a flame retardant synergist. Based on the

information currently available, EPA/OPPT identified releases to water as a primary exposure

pathway for aquatic and benthic organisms. Aqueous exposures that occur throughout the full

extent of the water column, including surface sediments, were considered to be the critical

exposure pathway for ATO environmental releases.  
Page 17 of 87

During the wastewater treatment processes, 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, ATO enters the environment via effluents and is

diluted to varying degrees in streams, lakes, estuaries, and oceans where aquatic and benthic

organisms may be exposed.  

Because of its physical chemical properties, ATO partitions to solid phases (sludge) during

wastewater treatment processing and with further treatment, can become concentrated in

biosolids. This organic carbon‐rich material may be disposed of via landfill, incinerated, or

utilized for land amendments. Wastewater Treatment Plants (WWTPs) need to dispose of

waste sludge continuously, and a portion of this material is applied to agricultural land each

year as biosolids.  ATO released to land has a long half‐life in soil, and may be available for

uptake by plants and soil invertebrates; however, ATO is expected to exhibit limited

bioaccessibility in soil due to chemical interactions with soil matrix constituents that limit

exposure potential.   

Surface waters and sediments were identified as primary exposure pathways because

ecological organisms are most likely to encounter ATO in these media. Soils and subsurface

sediments are not expected to provide a significant route of exposure for ecological organisms;

therefore, these environmental compartments were not evaluated. The ATO assessment did

not assess future risks that could occur if environmental conditions change.

Assessment Questions:

1. Do ATO concentrations measured in US surface waters pose risks for adverse effects in

aquatic organisms?

2. Do ATO concentrations measured in US sediments pose risks for adverse effects in

benthic organisms?

  
Page 18 of 87

Figure 1‐1. Conceptual Model* There are many potential sources of antimony releases that may occur directly to environmental media, or indirectly through WWT**Risks are expected to be low for this compartment due to limited bioaccessibility to soil‐dwelling organisms.
Page 19 of 87

1.4 ANALYSIS PLAN FOR ENVIRONMENTAL ASSESSMENT		

The objective of this assessment is to determine the media and pathways through which ATO

may adversely impact ecological organisms.  

Although information specifically related to ATO was limited, the data identified for antimony

and related compounds was considered sufficient for an ecological risk evaluation. Risks to

aquatic and sediment‐dwelling (i.e., benthic) organisms were selected as the focus of this

assessment, based on the available hazard and exposure information. The ecological

assessment endpoints evaluated for ATO include acute and chronic toxicity to aquatic and

benthic organisms. Risks to terrestrial/soil‐dwelling organisms were not evaluated due to

deficiencies in the available hazard and exposure data.   

A characterization of environmental hazards, as indicated by reported ecotoxicity benchmarks

(concentrations of concern; COC), and environmental exposures, as indicated by reported

industrial releases and environmental monitoring data, were used to evaluate risks to water‐ 

and sediment‐dwelling organisms.  

EPA/OPPT identified a subset of TRI facilities (i.e., those reporting production or use of ATO as a

flame retardant synergist) for exposure characterization.  Due to the availability of a sufficient

quantity of environmental monitoring data for the US, EPA/OPPT also collected information on

ambient antimony concentrations for those states in which the selected TRI facilities are

located.  

Hazard data obtained from relevant and reliable ecotoxicological studies were used to calculate

conservative (acute and chronic) COCs for aquatic and benthic organisms.  

To assess ecological risks, COCs for water‐ and sediment‐dwelling organisms are compared to

the measured or modelled antimony concentrations in these media using a risk quotient (RQ)

approach. For risk assessments that are based on an RQ approach, the comparison of exposure

concentrations to COCs has a binary outcome: either the environmental concentrations exceed

the COC, or they do not. A risk quotient greater than one (RQ > 1) suggests a potential for

adverse ecological effects; however, it is generally presumed that an RQ at or below one

indicates a low potential for adverse effects to ecological organisms.  
Page 20 of 87

2 SOURCES AND ENVIRONMENTAL FATE

2.1 INTRODUCTION

ATO enters the environment from natural (e.g., weathering of rocks, aerial transport of volatile

compounds) and anthropogenic sources (e.g., combustion, mining, smelting activities). In order

to assess ecological risks, a firm understanding of the primary uses, physicochemical properties,

and ultimate fate in the environment is necessary.  

2.1.1 Physical and Chemical Properties		

Antimony (Chemical Abstracts Service Registry Number [CASRN] 7440‐36‐0) is a naturally

occurring metalloid that belongs to Group 15 of the periodic table; it can exist in +5, +3, 0, or     

‐3 oxidation states, with +5 (i.e., pentavalent) and +3 (i.e., trivalent) being most common.

ATO is a dimorphic crystalline solid that exists in cubic (i.e., senarmontite; CASRN 12412‐52‐1)

and orthorhombic configurations (i.e., valentinite; CASRN 1317‐98‐2). Cubic senarmontite is

thought to be stable below 570 °C, whereas valentinite (i.e., infinite double chains) is formed

above this temperature (Freedman et al., 2003); the cubic form is shown in Figure 2‐1; the

orthorhombic form is shown in Figure 2‐2. Because ATO may exist in either molecular

arrangement and there are no exposure or hazard data that specifically differentiate between

the two, both were addressed as ATO (CASRN 1309‐64‐4) in this assessment.

Figure 2‐1. Representative Chemical Structure of Senarmontite (CASRN 12412‐52‐1)

Figure 2‐2. Representative Chemical Structure of Valentinite (CASRN 1317‐98‐2)
Page 21 of 87

ATO exhibits negligible vapor pressure at 25°C and tends to be soluble in hydrochloric acid,

tartaric acid, and caustic alkalies; it is insoluble in organic solvents and only slightly soluble in

water (Patnaik, 2003). ATO solubility (when determined in reconstituted standard water; ISO

6341 medium) was 2.76 mg/L after a seven‐day equilibration period at 22.2 °C and slightly

alkaline (pH 8) conditions (EC, 2008); typical physicochemical properties of ATO are shown in

Table 2‐1.

Table 2‐1. Physical‐Chemical Properties  

Property Value Comment

Physical state at 25 °C White solid   

Particle size   0.2‐13.89 μma   

Estimated vapor pressure at 25 °C <1 × 10‐6 torr   

Water solubility at 22.2 °C 2.76 mg/L at pH 8b Value determined after 7 days In  

reconstituted standard water,  

loading = 100 mg/L .   

Water solubility at 20 °C 19.7 mg/L at pH 5c

25.6 mg/L at pH7c

28.7 mg/L at pH 9c

Value determined after 24 hours in

distilled water.  

Melting point 656 °Cd   

Boiling point 1425 °Cd   

a

Weidenfeller (2005) as cited in EC (2008). b

LISEC (2002) as cited in EC (2008). c

Umweltanalytik GmbH (1993) as cited in EC (2008). d

Lide (2000) as cited in EC (2008).

2.2 PRODUCTION AND USES

Several types of antimony compounds are known to exist in nature; however, the mineral

stibnite is considered to be the primary source of processed antimony in commerce. In terms of

the total production volume and potential end‐use applications, ATO is the most widely used

antimony compound.  

No primary (i.e., mined) antimony production occurred in the US in 2011, with the exception of

one company that produced antimony metal and oxide from a foreign feedstock (USGS, 2012a).

Approximately 87 percent of the ATO consumed in the US between 2007 and 2011 was

imported from other countries (USGS, 2012a). Approximately 240 million pounds (lbs) of ATO

was produced globally in 2005, with the majority of production taking place in China (The

Chemical Company, 2012). Total combined production of ATO in the US and Mexico was

estimated at approximately 52.8 million lbs in 2005; however, the total amount of ATO

consumed in the US during this time period was estimated at roughly 70 million lbs

(International Antimony Association, 2010).
Page 22 of 87

Currently, Chemtura Corporation and US Antimony Corporation are the two main ATO

producers in the US (Roskill, 2011; USGS, 2006). While neither is listed as a manufacturer on the

2006 TSCA Inventory Update Reporting (IUR) rule, their respective websites indicated that they

manufacture and sell ATO products (Chemtura, 2012; USAC, 2012). The ATO national

production volume for 2012 was 5.4 million pounds (EPA, 2012). Table 2‐2 provides a list of

companies that manufacture or import ATO in the US.

Table 2‐2. US Manufacturers and Importers  

Company City State Manufacture Import

Arkema Inc. Philadelphia PA CBI CBI

Chemson Inc. Philadelphia PA No Yes

E. I. du Pont de Nemours and Company Wilmington DE No Yes

Eastman Kodak Company Rochester NY No Yes

Ferro Corporation Washington PA No Yes

General Electric Company

Mt. Vernon IN No Yes

Bridgeville PA No Yes

Washington WV No Yes

ICC Chemical Corporation New York NY No Yes

LG Chem America, Inc. Englewood Cliffs NJ No Yes

PolyOne Corporation Avon Lake OH No Yes

Sojitz Corporation of America New York NY No Yes

United Mineral & Chemical Corp. Lyndhurst NJ No Yes

CBI = confidential business information.

Source:  EPA (2010b)

ATO is primarily used as a synergist with halogenated flame retardants applied to textiles and

plastics; other applications include plastic stabilizer, paint pigment, catalyst precursor, fining

agent or antisolarant (Freedman et al., 2003). Additional uses for ATO are shown in Table 2‐3.

Table 2‐3. Major End‐Uses of Antimony in the US in 2011

End‐Use Market Share (%)

Flame retardant synergist 36

Transportation (including batteries) 23

Chemicals 16

Ceramics and glass 12

Other 13

Source:  USGS (2012a)
Page 23 of 87

ATO has no inherent flame retarding capacity (EPA, 2012a); however, when used as a flame

retardant synergist, it decreases the amount of halogen needed to impart acceptable flame

resistance (USGS, 2004; Weil and  Levchik, 2009). ATO interacts with chloride or bromide to

form flame quenching antimony trichloride or antimony tribromide (Weil and  Levchik, 2009). In

many instances, ATO is added to (rather than incorporated in) the actual polymer matrix

(AkzoNobel, 2012; Albemarle, 2000; ChemEurope, 2004), occurring as a separate, non‐reactive

(i.e., non‐covalently bonded) phase. ATO may be found in a number of flame‐retarded

products, including electrical equipment, wires, automotive parts, building materials, furniture,

carpets, mattress covers, plastics, and textiles (EC/HC, 2010). The average ATO content in flame

retardant formulations used to treat plastics generally ranges from 25 to 33 percent by weight,

whereas the relative abundance of ATO in the final (flame‐retarded) product can range from 1.5

to 12 percent (USGS, 2004). The demand for ATO in halogenated flame retardant applications

appears to be decreasing; approximately 25 million lbs of ATO were consumed by this market

sector in 2011, as compared to approximately 50 million lbs in 1999.

2.3 ENVIRONMENTAL FATE

ATO may be released to air, water, sediment and soil where it is slowly transformed into

Sb(OH)3 (antimonite), which may be further oxidized to Sb(OH)6

‐ (antimonate) under aerobic

conditions in the environment. Atmospheric releases may be dissolved in rain water, or

adsorbed to particulate matter and deposited to soil and surface waters via wet and dry

deposition (EC, 2008). Environmental fate processes can increase or decrease antimony

mobility in these environmental compartments, thereby impacting its availability to ecological

organisms. Processes that influence fate in water, sediment and soil systems are briefly

described below.

2.3.1 Environmental Persistence		

Though not particularly abundant, antimony compounds are naturally occurring in the earth’s

crust. ATO is an inorganic metal oxide; therefore, a fate analysis based on the octanol‐water

partition coefficient or the soil organic carbon‐water partition coefficient is not applicable.

Typical mass balance fugacity modeling is also inapplicable to ATO (or other metal ions)

released upon dissolution, because these substances exert zero partial pressure and fugacity in

air (Diamond et al., 1992). ATO is persistent in all media, because the antimony ions issued

upon dissolution cannot be degraded; however, antimony can be transformed between

different chemical species via biotic and abiotic processes which can impact bioaccessibility and

potential toxicity to ecological organisms (EC, 2008).  

2.3.1.1 Fate in Water		

Industrial effluents released to the environment may contain antimony as a result of losses

from various operating procedures, such as material handling (e.g., textile washing, etc.) and/or

washing of equipment. Most of this material will end up in industrial or municipal wastewater
Page 24 of 87

treatment plants, where significant amounts may be deposited in biosolids as a result of

alkalinization processes. Biosolids containing antimony may be incinerated, landfilled or applied

to land, with potential impacts on terrestrial and aquatic compartments due to leaching and

runoff (EC/HC, 2010).  

The fate and transport of dissolved and particulate forms of antimony are influenced by various

physical, chemical, and biological processes (Filella et al., 2002a, 2002b; Filella et al., 2007).

Fluctuations in flow parameters such as advection, or dispersion, and colloidal interactions

(e.g., settling, resuspension) can directly impact bulk movement of materials throughout the

water column. In addition, chemical interactions between aqueous constituents can impact

transport via shifts in thermodynamic equilibria that drive phase associations (e.g., adsorption,

dissolution), or biological transformations (e.g., methylation, mineralization) which can affect

environmental mobility and potential toxicity in aquatic systems.

Antimony concentrations in unpolluted natural waters are low, usually on the order of one part

per billion (ppb) or less; levels detected in ocean surface waters tend to be even lower (Filella et

al., 2002a). In a review of the published literature on antimony speciation in natural waters,

Filella et al. (2009) reported that relatively small amounts of Sb3+ were detected in oxygenated

waters. This finding has been corroborated by other authors, such as Garbos et al. (2000), who

determined that the majority of antimony detected in natural waters (i.e., 96 to 99 percent)

occurs as Sb5+. Cutter (1992) evaluated the rate of oxidation of Sb3+ to Sb5+ by measuring the

depth profiles for antimony species in the upper 100 meters of the Black Sea. In this study, no

Sb3+ was detected at surface levels; however, a gradual increase in the Sb3+ concentration (in

association with a decline in Sb5+ levels) was observed at depths beyond 60 meters. An

estimated pseudo‐first‐order rate constant of 0.008 day‐1 was calculated for the oxidation of

Sb3+ to Sb5+, corresponding to a residence time of approximately 125 days for Sb3+ in natural

waters.   

Despite prevailing thermodynamic theories, Sb5+ has been detected in anaerobic environments

in association with the following potential transport processes:  movement associated with

detritus sinking from aerobic waters, advection of surface waters containing Sb5+ and

biotransformation of Sb3+ by marine organisms. These transport mechanisms assume a slow

rate of conversion (i.e., reduction) to Sb3+ in the absence of oxygen; the rate constant for this

reaction was estimated at approximately 1.1 × 10‐6 days‐1 (Cutter, 1992). Atmospheric

deposition and biotic or abiotic reduction have been proposed as possible explanations for the

occurrence of Sb3+ in aerobic surface waters (ATSDR, 1992).   

2.3.1.2 Fate in Sediment

Antimony is mobile upon release to aquatic compartments (Filella et al., 2002a, 2002b);

however, it exhibits a strong tendency for adsorption to suspended particles and sediment,

such that a significant portion of dissolved antimony will ultimately end up in soil and sediment.

This may occur through adsorption to suspended particles as they settle out of solution
Page 25 of 87

(Hamilton‐Taylor et al., 1984), or precipitation reactions; a portion of the undissolved parent

material may also occur in sediment (EC, 2008).  

The extent to which antimony partitions between solid and solution phases is influenced by its

chemical speciation, which may be impacted by a variety of environmental processes (e.g., acid

neutralization, secondary mineral precipitation). Dissolved antimony concentrations are mainly

controlled by adsorption to (or precipitation of) metal oxide containing particulates, and

colloids of mineral secondary phases (e.g., amorphous ferric‐hydroxide) or other high surface

area materials (e.g., bacteria, fungi, organic matter). For example, acid neutralization via

dissolution of carbonate or dilution with surface waters may reduce the solubility of metal ions,

such as Fe3+, resulting in precipitation of secondary minerals that may serve as a reservoir for

antimony in sediment (Nordstrom and  Alpers, 1999). Colloidal and fine‐grained sorbent

materials may transport antimony for long distances before settling in the streambed (Ritchie

et al., 2013).   

The log sediment‐water partition coefficient for antimony ranges from 2.5 to 4.8 (EC/HC, 2010);

suggesting a tendency for adsorption; however, under certain circumstances antimony

adsorption to sedimentary organic matter may be reversed by changes in pH or redox

conditions. For example, Sb5+ adsorption to humic acids (as the soluble antimonate anion

Sb(OH)6

‐

), which is maximal under acidic conditions (Filella et al., 2002a; Tighe et al., 2005), may

be reversed under alkaline conditions, resulting in release of Sb5+ back into the water column.

In contrast, the neutral antimonite species (Sb(OH)3, which is the most stable form of Sb3+

under anaerobic (reducing) conditions, will adsorb to humic acids and metal oxides across a

wide range of pH values, and is therefore expected to be largely immobile in the environment

(Leuz et al., 2006; Scheinsost et al., 2006).   

2.3.1.3 Fate in Soil

Given its negligible vapor pressure and limited water solubility, ATO released to the terrestrial

compartment is expected to remain in soil; however, leaching and subsequent releases to local

surface waters (via run off) may occur as a result of seasonal rainfall or snow (EC/HC, 2010).

Subsequent exposures (i.e., those incurred as a result of leaching from ATO containing

materials) are unlikely to reach levels that would present a major concern for soil‐dwelling

organisms due to complex interactions with soil constituents that effectively limit antimony

mobility (bioaccessibility).  

Ambient antimony concentrations in US soils are low, ranging from 0.15 to 1.5 mg kg‐1 (USGS,

2014); however, elevated concentrations have been reported in association with anthropogenic

activities such as mining, smelting and the combustion of fossil fuels (ATSDR, 1992; Filella et al.,

2002a, 2002b). Although high levels (up to 700 mg kg‐1

) have been reported in association with

mining activities, antimony released to soil was determined to be biologically unavailable over a

wide range of environmentally relevant pH values (Flynn et al., 2003). Antimony bioaccessibility

to soil‐dwelling organisms may be characterized in part, by its distribution between solid and
Page 26 of 87

solution phases (i.e., soil‐water partition coefficient kd for soil); however, this parameter can

vary widely due to the site‐specific nature of phase associations. An evaluation of the published

literature showed significant variation in the available partition coefficients for antimony in soil

(EC, 2008); reported values ranged from 60 to 1316 (log kd = 1.8 to 3.2) in sandy soils (pH 5.7)

and 94 to 962 (log kd = 2.0 to 2.98) in heavy clay soils (pH 7.8). The observed variability in values

reported for different sites may result from numerous of factors including differences in media

type and loading rate. Key limitations associated with the use of kd values to assess

bioaccessibility to soil‐dwelling organisms include the following:

‐ The nature and stability (i.e., permanence) of binding interactions are poorly defined

‐ A lack of site‐specific information precludes quantitative evaluations of bioavailability  

‐ The environmental factors that influence adsorption/desorption reactions are dynamic

Antimony mobility in soils is primarily limited by adsorption to hydrous oxides of iron,

manganese, and aluminum (HSDB, 2012), which are fairly ubiquitous in the environment.

Because antimony exhibits a strong tendency for adsorption to metal oxides at environmentally

relevant pH levels, antimony is expected to exhibit limited mobility in this environmental

compartment (Leuz et al., 2006). Studies suggest that over time, antimony released to soil may

become sequestered (i.e., removed from the dissolved phase) by weathering or ‘aging’

phenomena. This natural attenuation mechanism likely reflects biotic and/or abiotic surface

charge interactions which induce formation of mineral aggregates of varying stability depending

on the strength of the chemical association (e.g., electrostatic, ionic, or covalent bond energy).

Site‐specific parameters (e.g., pH, matrix composition), and mineralogical properties (i.e., soil

type, predominate mineral phase) will likely influence the relative stability of the controlling

secondary mineral phase (Leuz et al., 2006).  

2.3.1.4 Fate in Air

Antimony is often released into the atmosphere as a result of high‐temperature anthropogenic

processes, such as metal smelting and fuel combustion, which represent major sources of

antimony air releases. As a non‐gaseous element with negligible vapor pressure, ATO is emitted

primarily as fine particulate matter (PM) which may be removed from the atmosphere by

gravitational settling or wet and dry deposition, depending on particle size. Larger particles are

generally removed from the atmosphere within hours, whereas smaller particles, such as those

generated via combustion processes, require a longer time period (on the order of weeks). In a

model comparing particle size to volatility, antimony exhibited an average atmospheric half‐life

of approximately 3.2 days (ATSDR, 1992).   

Regardless of the original form, combustion conditions (e.g., high temperature and excess

oxygen) are expected to release antimony oxides to the atmosphere; it is assumed that

combustion produces the trivalent form which may be further oxidized to the pentavalent form

upon contact with moisture and oxygen in the environment (EC, 2008). Ashes generated from

combustion processes are generally landfilled; however, small amounts may be released to air
Page 27 of 87

despite use of pollution control devices (e.g., bag filters, wet scrubbers) and subsequently

deposited on soil, surface water and sediments.

   

A study of wet and dry deposition over an eight‐week period in an area that was presumably far

from release sources found that 87 percent of antimony deposited from air was dissolved in

rain; 11 percent occurred as fine PM in rain, and approximately two percent was removed from

the air via dry deposition of fine PM (Stössel and Michaelis, 1986; as cited in EC, 2008). Airborne

sources of antimony releases to aquatic and terrestrial systems seem to be important even for

environmental systems that are located far from the original point of release; however, the

available data do not allow firm conclusions to be drawn (Filella et al., 2009).

2.3.2 Environmental Bioavailability, Transformation and Bioaccumulation

Antimony occurs in water, soil and sediment, where it can transition between various organic

and inorganic forms. The potential for ecological exposure varies according to the inherent

capacity and changing ability of various environmental compartments to adsorb, transform, and

release antimony in forms that can impact ecological organisms.  

2.3.2.1 Bioavailability

Changes in antimony bioavailability are largely a function of environmental processes that

effectively increase or decrease its mobility, thereby making it more or less accessible to

receptor organisms. For environmental risk assessments involving soils and sediments, the

ability of matrix constituents to sequester or transform antimony will ultimately determine how

much of the ambient exposure concentration is capable of affecting biological activity. Site‐

specific chemical and physical conditions can greatly influence the chemical form (i.e.,

speciation) of antimony and thus the degree to which it adsorbs to matrix constituents and is

thereby removed from solution.  

The main purpose for bioavailability considerations is to provide a more accurate or ‘realistic’

characterization of exposure potential. Qualitative evidence can assist in determining how

much of the total antimony concentration is actually available for uptake, and this is likely to

provide a better indication of the potential for adverse ecological effects. Biochemical

processes (e.g., ion exchange, methylation) can change the physicochemical form, or speciation

of antimony, thereby impacting its accessibility to ecological organisms. For the purposes of this

discussion, bioaccessibility is defined as the amount of antimony that an organism can extract

from exposure media, whereas bioavailability reflects the amount that is taken up from the gut

contents into the general circulation.  

Adsorption, dissolution and precipitation reactions are primary determinants of antimony

availability in soils. Antimony can occur in the solid or aqueous (i.e., soil solution) phase; in

solution, it can exist either as the hydrated free ion or be complexed with various organic or

inorganic ligands. Conversely it may be sequestered by various adsorption or precipitation
Page 28 of 87

reactions that effectively remove it from solution via formation of relatively inert secondary

mineral phases (Leuz et al., 2006). Dissolved antimony is generally more available for a variety

of processes, including uptake and transport; therefore, soil pH is a primary determinant of

availability to soil‐dwelling organisms, as it directly influences solubility. Acidic conditions and

high humic acid content (> 5 to 10 percent) generally decrease antimony mobility in soil

(Alloway, 2005).  

Dissolution kinetics represent a major controlling factor for antimony availability in soil, due to

the formation of stable secondary mineral phases. Both the extent to which these mineral

species occur in a particular soil and their solubility in various biological fluids will influence

bioavailability to ecological organisms. Previous studies have demonstrated slow dissolution

kinetics for antimony in soil and limited bioaccessibility to soil‐dwelling organisms (Denys et al.,

2012; Oorts et al., 2008).   

Studies conducted using soluble antimony compounds freshly mixed with soils or sediments

generally do not show significant reductions in bioavailability, and will not provide an accurate

representation of the influence exerted by natural modifying factors that can dramatically

affect exposure potential. In general, soil conditions that promote precipitation or adsorption

tend to reduce mobility in the environment and bioaccessibility to ecological organisms.

Because this information is largely site‐specific, extrapolation across different geographic

regions is not feasible; however, as a first approximation, assumptions regarding expected

impacts of various physicochemical parameters (e.g., soil pH, matrix constituents), can be used

in a weight‐of‐evidence approach to support the assumption of limited availability in soil. The

following information is intended as a guideline; soils at specific sites may not conform to all of

the general trends discussed here.  

- Both the form of antimony (speciation) and the amount of time it has been present in

soil or sediment can influence its bioaccessibility to ecological organisms; antimony

generally exhibits limited solubility and slow dissolution kinetics in soil.  

- Site‐specific parameters can influence soil weathering (‘aging’) reactions which further

decrease antimony solubility. Limited solubility and bioaccessibility to ecological

organisms is expected in soils that contain calcium, sulfide or metal oxides (which are

ubiquitous in the environment) due to increased adsorption potential.

- Antimony exhibits increased mobility under alkaline conditions, and in soils containing

elevated (> than 5 to 10 percent) fulvic acid content; however, these soils often contain

relatively insoluble sulfides (due to the actions of sulfate‐reducing bacteria in seasonally

flooded soils) that can effectively remove antimony from the soil solution phase.  

- Rainfall, bacterial activity and degradation of organic matter can decrease soil pH, which

generally increases antimony adsorption to soil particles. Since most US soils tend to be

acidic to slightly alkaline, and antimony generally exhibits limited solubility and slow
Page 29 of 87

dissolution kinetics in soil, it is expected to exhibit limited bioaccessibility to soil‐

dwelling organisms.  

2.3.2.2 Transformation

Transformations of antimony have been reported in the presence of organic matter and

microorganisms. For example, humic acids have been shown to oxidize Sb3+ to Sb5+ (Buschmann

et al., 2005; Steely et al., 2007) and various microorganisms (e.g., bacteria, fungi), have been

shown to methylate relatively small amounts (< 0.1 percent) of antimony (Duester et al., 2008;

HSDB, 2012). Information on antimony methylation is limited, but measured concentrations are

usually low, and the extent of antimony methylation in environmental media appears to be

significantly less than that of other elements, such as arsenic (Filella et al., 2009). Certain metal

hydroxides, soil bacteria and marine organisms have also been shown to transform antimony

(Filella et al., 2007; HSDB, 2012); thereby impacting its bioaccessibility and potential toxicity to

ecological organisms.  

There is considerable evidence for the natural attenuation of antimony in soils due to formation

of low‐solubility species during equilibration with the soil constituents. For example, fungal

involvement has been implicated in mineral formation (Fomina et al., 2007); fungal organisms

can acidify their microenvironment and excrete a variety of metabolites (e.g. siderophores,

carboxylic acids) that effectively sequester metals within a matrix of secondary minerals. All

major microbial groups have demonstrated roles in metal immobilization and mineral

formation. Reported mechanisms include microbial oxidation or reduction of a metal species

released via mineral dissolution, and metabolite excretion or other metabolism‐dependent

transformations, with microbial surfaces providing chemically reactive sites for adsorption,

nucleation and subsequent formation of mineral precipitates (Gadd, 2010).   

Studies of sequential extraction techniques have shown that a low percentage of antimony is

extracted from soils using mild or aqueous (e.g., water soluble) solutions; most occurs in the

“residual” or “recalcitrant” fraction. Largely on the basis of such results, it has been established,

and is generally accepted, that antimony is essentially immobile in soils (Filella et al., 2009).  

Similar results were demonstrated in field tests used to confirm the chemical distribution of

antimony in creek sediments contaminated by a local smelter operation. These results showed

that antimony was predominantly associated with the residual sediment fraction, with small

amounts (~ 3 to 7 percent) occurring in the water soluble or exchangeable fractions (Baeza et

al., 2010). As mentioned previously, a major difficulty encountered in applying information on

phase associations is their site‐specific nature; uncertainties regarding the primary factors that

control antimony mobility (and bioaccessibility) in aquatic environments remain.  

Antimony availability to plants is often low, even in contaminated sites, due to limited water

solubility (Gal et al., 2007; Hammel et al., 2000); however, accumulation has been reported in

some plant species (Tschan, 2009). Antimony uptake and bioavailability to plants was

investigated from field sampling of grass near a smelter emitting antimony, and in experiments
Page 30 of 87

in which grass was grown under laboratory conditions using the same antimony‐contaminated

soils (Ainsworth et al., 1990). It was concluded that the high levels of antimony observed in

grass grown near the smelter occurred primarily from atmospheric deposition; antimony

uptake from soil, and translocation to shoots accounted for only a small percentage of the total

antimony concentration, suggesting low bioavailability in soils. Biota obtained from the site also

showed minimal uptake of antimony from soil.  

2.3.2.3 Bioaccumulation

Bioaccumulation is of potential concern due to direct impacts to organisms that accumulate

metals and indirect impacts to those that consume them (i.e., trophic transfer). Previous

studies have indicated that antimony may bioaccumulate in aquatic organisms; however, there

is considerable uncertainty regarding the derivation of these values (EC, 2008).  

Antimony is not expected to exhibit significant bioaccumulation potential in aquatic organisms;

metal bioaccumulation in this environmental compartment is largely governed by trophic

feeding groups, with detritivores accumulating higher levels than herbivores or predatory

groups. Although the potential for secondary impacts as a result of trophic transfer exists, it is

not well characterized. For example, Corbicula fluminea, a filter‐feeding clam that consumes

algae, bacteria, and fine detritus, has been reported to bioaccumulate antimony to a greater

extent than other predators or omnivores (Duran et al., 2007); however, Baeza et al. (2010)

found no significant correlation between antimony bioaccumulation in C. fluminea and elevated

antimony levels in sediments or lower trophic organisms.

Reliable bioaccumulation studies for antimony are limited (Gal et al., 2007; Heier et al., 2009).

In one study, Heier et al. (2009) studied time‐dependent accumulation in brown trout residing

in stream waters that exhibited elevated antimony concentrations (2 to 3 μg/L) due to runoff

from a shooting range. After 23 days of exposure, antimony accumulation was identified on the

gills and in liver tissue taken from exposed fish. Measured antimony concentrations ranged

from 0.01 to 0.08 μg/L in gill, and 0.0 to 0.07 μg/L in liver, suggesting antimony accumulation in

brown trout is neither significant nor tissue specific. In another study, Gal et al. (2007) studied

antimony bioavailability in earthworms near a former antimony mining and smelting site where

total antimony concentrations in sediment ranged from 10 to 1,200 mg/kg. In the two species

of earthworm studied, measured antimony concentrations ranged from 0.72 to 26 mg/kg; the

bioconcentration factors (i.e., the ratio of antimony concentrations in biota to that in soil) were

less than one, suggesting minimal concern for antimony bioconcentration in these organisms.

Additional information on antimony bioaccumulation and bioconcentration can be found in                                   

Table 2‐4 and Table 2‐5.
Page 31 of 87

                                        Table 2‐4. Antimony Bioconcentration in Terrestrial Plants

Terrestrial Plants Speciesa

Maize Zea mays BCF:  0.186

Sunflower Helianthus annuus BCF:  0.266

a

EC/HC (2010).

BCF = bioconcentration factor

        Table 2‐5. Bioconcentration and Bioaccumulation of Antimony in Aquatic Species

Salt Water Speciesa Fresh Water Speciesb

Fish BCF c:  0.15 (Two spot goby); BCF:  

0.40 (Shanny)

BAF c:  0.43 (Labeobarbus kimberleyensis)

Aquatic

invertebrate

No data BAF c:  5.65 (Hyalella azteca)

Algae BCF c: 17.00 (Coralline or red algae) BCF c:  24.00 (Chlorella vulgaris)

a

EPA (2012b). b

EC/HC (2010). c

BAF = bioaccumulation factor; BCF = bioconcentration factor

Considerable uncertainty may be associated with the application of metal bioaccumulation (or

bioconcentration) factors, as these values can be highly variable (EPA, 2007). This may reflect

differences in metal speciation and the related effects on bioavailability and/or complex

toxicokinetic mechanisms that influence body burden (e.g., uptake, metabolism, distribution,

elimination). Saturable uptake and storage mechanisms, and the innate ability to regulate

bioaccumulated metal concentrations within a certain physiological range may partially explain

the inverse relationship that is often reported between bioaccumulation (or bioconcentration)

factors and metal exposure concentrations (McGeer et al., 2003).

2.3.3 Conclusion of Environmental Fate

ATO is an environmentally persistent metalloid that exhibits low bioaccumulation potential.

Generally speaking, pentavalent antimony is the thermodynamically favored under aerobic

conditions in the environment; however, trivalent antimony predominates under anaerobic

conditions. Antimony is primarily removed from the atmosphere via wet and dry deposition.

Antimony generally exhibits limited solubility and slow dissolution kinetics in soil, therefore it is

expected to exhibit limited bioaccessibility.  However, mobility in soil and sediment may be

influenced by site‐specific environmental parameters (e.g., matrix composition, pH; redox

potential) which can significantly impact antimony speciation and bioavailability to ecological

organisms.
Page 32 of 87

3 ENVIRONMENTAL ASSESSMENT

3.1 INTRODUCTION

The environmental assessment evaluated ecological exposures that may be associated with

ATO use as a flame retardant synergist.   

3.1.1 Environmental Exposure Assessment

Exposure characterization focused specifically on industrial sectors reporting ATO production or

use as a flame retardant synergist. The strategy used for the assessment included: 1) a

refinement of the release data reported in the 2010 TRI to identify a subset of ATO production

or use facilities, 2) collection of monitoring data for water, sediment, and soil from the states in

which these facilities are located, and 3) modeling of TRI releases to provide site specific

estimates of environmental exposure concentrations.  

3.1.2 Toxics Release Inventory		

To identify the selected TRI facilities that may use or produce ATO‐containing flame retardants,

EPA/OPPT examined the primary North American Industry Classification System (NAICS) codes

reported in the 2010 TRI for antimony compounds.2 The NAICS codes descriptions include a

general heading and several subcategories that were used to screen TRI facilities for inclusion in

the exposure assessment. ATO is not specifically listed on the TRI, but is reported under the

broader category of antimony compounds, which covers all chemicals that contain antimony.

EPA/OPPT could not distinguish ATO releases from those of other antimony‐containing

chemicals reported on the TRI; therefore, exposure characterizations were based on the

conservative assumption that all releases of antimony compounds reported in the 2010 TRI

reflect ATO releases associated with its use as a synergist in halogenated flame retardants.  

A total of 270 facilities reported releases of antimony compounds to the TRI in 2010 Table 3‐1.

The majority of reported environmental releases were to land and other disposal media (which

included on‐ and off‐site transfers, landfill, underground injection, land treatment, surface

impoundments, and other transfers for off‐site disposal). The on‐ and off‐site land and other

disposal of antimony compounds totaled approximately 970,000 lbs, of which, approximately

one‐third was associated with on‐site land releases from nine facilities (NAICS code 325188 ‐ All

Other Basic Inorganic Chemical Manufacturing), which are subject to Subtitle C regulation of

hazardous materials under the Resource Conservation and Recovery Act.  

                                                      

2 A full description of the 6‐digit NAICS codes is available from the Department of Census

(http://www.census.gov/eos/www/naics/).
Page 33 of 87

The on‐site air releases of antimony compounds totaled approximately 12,000 lbs, of which

approximately two‐thirds were from stack releases, and one‐third were from fugitive air

emissions. A significant portion of these releases likely resulted from the combustion of fossil

fuels; however, the precise quantity cannot be determined from TRI data.  

The 2010 TRI water releases reported for antimony‐containing compounds totaled

approximately 15,700 lbs. Of this total, approximately 3,100 lbs were associated with direct

surface water discharges, and approximately 12,600 lbs were transferred to Waste Water

Treatment Plants (WWTPs) or other off‐site water treatment facilities. Table 3‐1 summarizes

TRI information on environmental releases reported by industry sectors identified under the

scope of this assessment (i.e., those using NAICS reporting codes that indicate production or

use of ATO‐containing flame retardants). No other environmental release data were identified

for this use scenario.
Page 34 of 87

Table 3‐1. Summary of 2010 TRI Releases of Antimony Compounds

Life Cycle Stage

NAICS Total

Number

of

Facilities

Reporting

On‐site and Of

Dispos

Primary

NAICS

Code

Description Total Air Releases Tot

Wat

Releas

Manufacture of antimony oxides 325188 All other basic inorganic chemical

manufacturing

9 2,106 57

Plastics material and resin, synthetic

rubber, and fiber manufacturing

325211 Plastics material and resin manufacturing 27 884 4,93

325212 Synthetic rubber manufacturing 2 85 26 325222 Non‐cellulosic organic fiber manufacturing 5 408 18

Plastics compounding and

conversion

325991 Custom compounding of purchased resins 55 1,231 540 

326113 Unlaminated plastics film and sheet

(except packaging) manufacturing

18 339 110 

326121 Unlaminated plastics profile shape

manufacturing

2 3 0 

326130 Laminated plastics plate, sheet (except

packaging), and shape manufacturing

2 10 1 

326150 Urethane and other foam product (except

polystyrene) manufacturing

3 0 0 

326199 All other plastics product manufacturing 25 330 10 

Rubber compounding and

conversion

326220 Rubber and plastics hoses and belting

manufacturing

8 14 44 

326291 Rubber product manufacturing for

mechanical use

3 30 0 
Page 35 of 87

Life Cycle Stage

NAICS Total

Number

of

Facilities

Reporting

On‐site and Of

Dispos

Primary

NAICS

Code

Description Total Air Releases Tot

Wat

Releas

326299 All other rubber product manufacturing 16 1,163 33 

Fabric Finishing 313210 Broadwoven fabric mills 2 60 0 

313230 Nonwoven fabric mills 1 0 2,16313311 Broadwoven fabric finishing mills 4 53 1,13313312 Textile and fabric finishing (except broad

woven fabric) mills 1 669 669 

313320 Fabric coating mills 15 2,893 4,92

Fiberglass insulation manufacturing 327993 Mineral wool manufacturing 3 0 0 Drawing and insulating of wires and

fiber optic cables

331422 Copper wire (except mechanical) drawing 23 844 90 

335921 Fiber optic cable manufacturing 2 0 0 335929 Other communication and energy wire

manufacturing 38 3 25 

Plastic‐containing

electronic/electrical productmanufacturing

335931 Current‐carrying wiring device

manufacturing 5 255 5 

Textile product mills and apparel

manufacturing

314110 Carpet and rug mills 3 255 207 

314999 All other miscellaneous textile product

mills 1 0 0 

1 See: supplemental data file titled “Supplemental 2010 TRI Data for the TSCA Work Plan Chemical Risk Assessment of ATO (External Review Draa Mass in lbs/yr can be converted to kg/yr by multiplying lbs/yr by 0.4536 kg/lb (numbers are rounded to the nearest integer).

b Includes transfers to WWTPs and other off‐site wastewater treatment facilities.
Page 36 of 87

3.1.3 			Measured Concentrations in the Environment		

EPA/OPPT developed subsets of environmental monitoring data from targeted searches of the

USGS‐NWIS and EPA’s STORET databases. Approximately 8,500 water, sediment and soil

samples were collected from these sources; data collection focused on the ten states in which

selected TRI facilities (i.e., those reporting use or production of ATO‐containing flame

retardants) were located.  

3.1.3.1 Compilation of Environmental Monitoring Data from USGS‐NWIS

Database

Environmental monitoring data were downloaded from the USGS‐NWIS website

(http://nwis.waterdata.usgs.gov/usa/nwis/qwdata, accessed in July 2012). A subset3 of the

entire dataset, which spanned more than 30 years (i.e., 1987 to 2011), was used to evaluate

environmental exposures.   

The subset used for exposure characterization was limited to monitoring samples obtained

from states that included at least one of the TRI facilities identified under the scope of this

assessment (i.e., AL, IN, KY, NC, NJ, NY, SC, TN, TX, and VA). These monitoring sites are shown in

Figure 3‐1. The USGS dataset was further culled to focus specifically on water, sediment, and

soil samples obtained from these regions. The available monitoring data were split into two

groups: (1) measured antimony concentrations from all years, and (2) measured antimony

concentrations from 2009 to 2012.    

3.1.3.2 Compilation of Environmental Monitoring Data from the STORET

Database

Another large source of environmental monitoring data was obtained from EPA’s STORET

database (http://www.epa.gov/storet/dw_home.html; accessed July 2012). The STORET

dataset included measured antimony concentrations in water and sediment only and was

limited to samples taken from states that included at least one of the TRI facilities identified in

Table_Apx C‐1. The STORET data were divided into two categories: (1) concentrations measured

in water (μg/L), and (2) concentrations measured in sediment (mg/kg dw). An attached file has

a supplemental dataset4.

                                                      

3

See attached file titled “Supplemental subset of USGS‐NWIS Dataset for the TSCA Work Plan Chemical Risk

Assessment of ATO (External Review Draft).xlsx”   4 See attached file titled “Supplemental subset of STORET Dataset for the TSCA Work Plan Chemical Risk

Assessment of ATO (External Review Draft).xlsx”
Page 37 of 87

Figure 3‐1 Geographic Locations of TRI Facilities and Environmental Monitoring Data
Page 38 of 87

3.1.3.3 Location of Monitoring Sites Relative to TRI Facilities		

Upon retrieval and analysis of the monitoring data, it became clear that none of the actual

monitoring sites were located within the primary watershed of the identified TRI facilities.

Hence, concentrations of antimony found in water, sediment or soils from the USGS or STORET

monitoring programs could not be clearly associated with ATO production or processing.

Therefore, EPA/OPPT estimated environmental exposure concentrations from TRI facilities

using the modeling approach described in Section 3.2.3.

3.1.4 Estimated Environmental Releases		

Having found that available environmental monitoring data were not associated with industry

sectors identified under the scope of this assessment (i.e., those using NAICS reporting codes

that indicate production or use of ATO‐containing flame retardants), EPA/OPPT used domestic

release data reported in the 2010 TRI to estimate ATO surface water concentrations.  

3.1.4.1 EFAST Modeling

EPA’s Exposure and Fate Assessment Screening Tool, Version 2.0, (E‐FAST2) was developed to

support EPA assessments of potential environmental exposures. For exposure characterization,

the E‐FAST2 model was used to estimate ATO surface water concentrations based on water

releases reported by selected TRI facilities. The exposure scenarios included in the E‐FAST2

model contain default parameter values that allow for exposure estimation with minimal data

entry. Because E‐FAST2 incorporates defaults that encompass either a combination of upper

percentile and mean exposure parametric values, or all upper percentile parametric values, the

resulting model predictions represent high‐end exposures estimates. EPA/OPPT acknowledges

the conservative nature of this approach.  

E‐FAST2 includes a Probabilistic Dilution Model (PDM) which predicts the number of days per

year in which a designated exposure, or effect level (i.e., COC) will be exceeded in ambient

waters as a result of chemical discharges (effluents) released from a given facility. PDM

analyses can be performed on stream reaches with measured flow data or stream reaches that

incorporate estimated streamflow values. The PDM model provides chronic risk estimates that

are derived from a simple mass balance approach of chemical dilution/emulsion into stream

water; however, the input parameters are not single point estimates.  

In reality, streams exhibit highly variable seasonal flow patterns. In addition, manufacturing

processes include various operating procedures that can change intermittently, thereby

affecting effluent flow rates and the total amount of chemical released to the environment over

a given time interval. The PDM incorporates probability distributions from Monte Carlo

simulations as analysis inputs for calculating the resulting probability distribution for the

chemical concentration that may be seen in stream waters. Ultimately it predicts the number of

days per year in which the modelled stream concentrations are expected to exceed the
Page 39 of 87

designated effect levels (i.e., COCs) identified for aquatic organisms based on the total amounts

of chemical released per day. (For more information, see the user guide/background document

for EFAST 2 at this web address: http://www.epa.gov/oppt/exposure/pubs/efast2man.pdf).  

A summary of the input data used for the E‐FAST2 model and a summary table which includes

site specific input parameters is provided in Appendix B.  

3.1.4.2 Criteria for Selection of TRI Facilities

Three selection criteria were used to identify the TRI facilities included in this exposure

characterization:  (1) reported water releases of ‘antimony compounds’ above 1 lb/yr; (2)

reported NAICS codes that indicate production or use of ATO as a flame retardant synergist;

and (3) stream releases for which the lowest streamflow data, or “7Q10” value (i.e., the

average consecutive seven‐day streamflow data calculated over a 10‐year period) was

available. The 7Q10 streamflow data are linked to the National Pollution Discharge Elimination

System (NPDES) codes for each facility, which were found using EPA’s Discharge Monitoring

Report (DMR) Pollutant Loading Tool (available at

http://cfpub.epa.gov/dmr/error_correction.cfm).

Using this approach, EPA/OPPT identified 36 TRI facilities TRI, 2010) that reported water

releases of antimony compounds and listed primary NAICS codes associated with ATO use as a

synergist in halogenated flame retardants; 33 of these indicated releases >1 lbs/yr. The NPDES

numbers matching the reported latitudes and longitudes were found for 14 TRI facilities

providing site‐specific streamflow rates for use in the E‐FAST2 model. The geographic location

of these TRI facilities are shown in Figure 3‐1 (red triangles).

3.1.4.3 Release Scenarios used for Exposure Characterization

Because TRI facilities do not indicate the total number of release days per year (i.e., a required

input for E‐FAST2 modeling), two hypothetical release scenarios were modeled (assuming total

TRI releases of ‘antimony compounds’ reported during the 2010 reporting cycle occurred over a

time period spanning 24 or 250 days per year). The first scenario (i.e., assuming total releases

occurred over a period of 250 days) was developed based on a typical operating schedule which

includes 260 work days and ten vacation days/holidays during the calendar year The more

conservative scenario (i.e., assuming total releases occurred over a period of 24 days), was

developed based on a hypothetical situation wherein a facility might incur intermittent releases

as a result of cleaning practices that span a 24‐day time period. Both hypothetical release

scenarios were used to estimate surface water concentrations based on E‐FAST2 model results

for water releases reported by the 14 TRI facilities evaluated in this assessment. These

hypothetical release scenarios were developed to provide an upper and lower bound for the

range of estimated surface water concentrations predicted by the E‐FAST2 model. The results

of this exposure assessment, which were used to assess risks to aquatic organisms, are

provided in Table 3‐2.    
Page 40 of 87

Table 3‐2 Summary of E‐FAST2 Model Results for TRI Facilities with Assumptions for Daily Release Amounts1

Facility Name Latitude Longitude NPDES Code

Total Surface

Water

Discharge (lbs)

Scaled

Annual

Release (kg)

Results

250 Days

(kg)

Results

24 Days

(kg)

Eastman Kodak Co.,

Kingsport, TN 37662 36.519823  ‐82.540429 TN0002640 1,042 557.22 2.23 23.22

Guilford East Plant,

Kenansville, NC, 28349 35.021587  ‐77.859994 NC0002305 669 357.75 1.43 14.91

Auriga Polymers Inc.,

Spartanburg, SC, 29307 35.030528  ‐81.861366 SC0002798 320 171.12 0.68 7.13

Sabic Innovative Plastics US

LLC, Selkirk, NY, 12158‐0145 42.575416  ‐73.853563 NY0007072 250 133.69 0.53 5.57

Sabic Innovative Plastics

LLC, Mount Vernon, IN,

47620‐9364

37.906389  ‐87.927778 IN0002101 250 133.69 0.53 5.57

Sage Automotive Interiors,

Abbeville, SC, 29260 34.167222  ‐82.371389 SC0000353 248 132.62 0.53 5.53

Performance Fibers

Operations Inc., Salisbury,

NC, 28147

35.720234  ‐80.610867 NC0004944 95 50.80 0.20 2.12

Dupont Teijin Films, Chester,

VA, 23836 37.346259  ‐77.285811 VA0003077 51 27.27 0.11 1.14

Haywood Co Brownsville,

TN, 38012‐1797 35.605911  ‐89.240068 TN0041939 28 14.97 0.06 0.62

Bayshore Industrial,  

La Porte, TX, 77571 29.63038  ‐95.030311 TX0022799 13.505 7.22 0.03 0.30

Covalence Adhesives,  

Franklin, KY, 42134 36.753889  ‐86.575278 KY0074659 5 2.67 0.01 0.11

General Cable Industries

Inc., Marion, IN, 46953‐

2406

40.55425  ‐85.654944 IN0002372 5 2.67 0.01 0.11

Borealis Compounds Inc.,

Port Murray, NJ, 07865 40.813361  ‐74.888111 NJ0028657 5 2.67 0.01 0.11

Cerro Wire LLC, Hartselle,

AL, 35640 34.412913  ‐86.917395 AL0054674 5 2.67 0.01 0.11

1

Screening‐level estimates are based on a range of days of release, which may not represent actual releases or environmental

concentrations at these facilities. 2 Mass in lbs/yr can be converted to kg/yr by multiplying ‘lbs/yr’ by 0.4536 kg/lb (numbers are rounded to the nearest integer).

NPDES = National Pollution Discharge Elimination System

   
Page 41 of 87

3.2 ECOLOGICAL HAZARD ASSESSMENT

The ATO hazard assessment was based on readily available ecotoxicity information, including

published studies and studies evaluated in previous risk assessments. Much of the information

summarized in this section was obtained from the European Union Risk Assessment Report for

Diantimony Trioxide (EC, 2008), and the Environment Canada Screening Assessment for the Challenge

‐ Antimony Trioxide (EC/HC, 2010). This information was supplemented by subsequent searches of

the peer‐reviewed literature (from 1978 to 2012) using EPA’s ECOTOX database to identify additional

information from journal articles which were subsequently retrieved and evaluated for data

adequacy. Data adequacy criteria include, but are not limited to: test substance identification,

number of organism in each dose/concentration group, dose/concentration level, route/type of

exposure, duration of exposure, species, controls (for additional details, see EPA, 1999. Studies

conducted according to established EPA or OECD Guidelines as well as studies using other protocols

were included if they met data adequacy criteria.

Due to the limited experimental data available for ATO [Sb2O3], antimony trichloride [SbCl3] and

antimony sulfate [Sb2(SO4)3] were used as supporting chemicals to assess hazards to aquatic and

sediment‐dwelling organisms. The amount of time necessary to dissolve and liberate antimony ions

capable of exerting adverse effects on ecological organisms is an important consideration for toxicity

testing, as this parameter will vary for different antimony compounds. Antimony trichloride dissolves

more readily (i.e., the liberation of antimony ions will be slower for ATO); however, this is not

expected to significantly impact testing results as the total amounts of antimony issued upon

dissolution are expected to be comparable for both compounds (EC, 2008).  

The majority of the available ecotoxicity studies utilized trivalent antimony compounds for testing,

but did not include specific information on antimony speciation measurements. Since antimony is

oxidized under aerobic conditions, the results of toxicity testing with trivalent antimony likely reflect

exposure to a mixture of trivalent and pentavalent antimony.  Because there is no conclusive

scientific evidence to support a valence‐specific differential for antimony‐related ecological effects,

no further distinction between antimony valence forms was made when evaluating hazard potential.  

Acceptable studies were identified for aquatic and benthic organisms; however, a systematic review

of the published literature revealed a relative paucity of antimony toxicity and exposure data for soil.

Because the available environmental exposure information was inadequate to support a quantitative

assessment of risks to soil‐dwelling organisms, this environmental compartment was not evaluated in

the current assessment. This conclusion is supported by the Ecological Soil Screening Levels (Eco‐SSL)

document for Antimony prepared by EPA’s Office of Solid Waste and Emergency Response (EPA,

2005).  

A summary of the available ecotoxicity data for ATO that were deemed adequate for consideration in

this assessment are provided in Table 3‐3, Table 3‐4, Table 3‐5 and the studies described in more

detail in the appropriate section below.  

Application of uncertainty factors based on established EPA/OPPT methods were used to calculate

lower bound effect levels (referred to as the concentration of concern; COC) that would likely
Page 42 of 87

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., 2006; EPA, 2012d; EPA, 2013).  

3.2.1 Acute Toxicity to Aquatic Organisms

Acute aquatic toxicity studies considered for this assessment are summarized in Table 3‐3.  

EPA/OPPT evaluated the available acute toxicity studies for fish, invertebrates, and plants; antimony

trichloride (SbCl3) was used as the test substance for most laboratory toxicity tests, whereas ATO

(Sb2O3) was used in one plant study. The environmental effect levels provided below are based on

measured antimony concentrations reported in these laboratory studies.

Acute toxicity to fish species was evaluated using SbCl3 with red sea bream (Pagrus major)

(Takayanagi, 2001). Juvenile sea bream (approximately 2.0 g in weight) were exposed to varying

concentrations of either selenium and SbCl3 or selenium and SbCl5 for 96 hours under static exposure

conditions. A measured 96‐hr LC50 of 6.9 mg Sb/L was reported by the study authors. Though this

study is considered to be reliable, uncertainties regarding antimony speciation and the specific test

concentrations used for toxicity evaluations (i.e., only a range of test concentrations was provided in

the report), preclude its use for COC derivation.  

For aquatic invertebrates, a 96‐hr static test with SbCl3 in hydra (Chlorohydra viridissima) was

conducted in duplicate for each test concentration evaluated using five organisms per replicate.

Measured antimony concentrations were 0.0, 1.11, 1.50, 2.54, 4.59, and 5.39 mg Sb/L, resulting in an

LC50 value of 1.77 mg Sb/L (95% C.I. = ± 0.7 mg Sb/L; TAI, 1990).  

Two studies using the algal species (Pseudokirchneriella subcapitata) were conducted using either

SbCl3 or Sb2O3. Heijerick and Vangheluwe (2004) exposed this green algal species to the following

measured concentrations of SbCl3:  < 0.001 (control), 1.22, 2.11, 4.00, 6.73, 12.2, 21.9, 36.6, and 17.3

mg Sb/L for 72 hours under static exposure conditions according to the OECD 201 test guidelines. The

starting cell density was 1 X 106 cells/mL, with three replicates per test concentration and six

replicates per control. The LISEC (2001), which also adhered to OECD 201 test guidelines, exposed P.

subcapitata to Sb2O3 (purity 98 %) at the following measured concentrations: 0, 0.074, 0.156, 0.323,

0.396, 1.32, 2.4 mg Sb/L. The starting cell density was 1 X 104 cells/mL, with three replicates per test

concentration and six replicates per control. In both studies, biomass, growth, and growth rate were

determined for control and treated groups at 24, 48, and 72 hours after test initiation.

Growth rate was determined to be the most relevant endpoint (i.e., ErC50; the effective concentration

wherein 50 percent of tested organisms experienced a decline in growth rate), with the geometric

mean of the maximum allowable toxicant concentration (MATC) determined to be 2.91 and 0.723 mg

Sb/L for SbCl3 and Sb2O3, respectively (Heijerick and Vangheluwe,2004; LISEC (2001). The LISEC (2001)
Page 43 of 87

study is considered unreliable because the ErC50 value could not be determined due to uncertainties

regarding the dose response relationship (e.g., the limit dose induced minimal inhibition of growth

rate, which is the most responsive endpoint); this value only decreased by three percent at the

highest concentration tested (EC, 2008). Though the data reported by Heijerick and Vangheluwe

(2004) are considered to be reliable (i.e., based on the OECD guidelines for the testing of chemicals ‐ 

Freshwater Algae and Cyanobacteria, Growth Inhibition Test), the reported ErC50 value (> 36.6 mg

Sb/L) indicates that this algal species is not particularly sensitive to the effects of SbCl3 exposure.

The acute COC for aquatic organisms was derived from the 96‐hr LC50 of 1.77 mg Sb/L obtained using

SbCl3 in the aquatic invertebrate, Chlorohydra viridissima. This value was divided by an assessment

factor (uncertainty factor) of five for invertebrate species to yield an acute COC of 0.354 mg Sb/L (354

μg Sb/L) in accordance with established EPA/OPPT methods (EPA, 2012d; 2013). Uncertainty factors

are included to account for differences in inter‐ and intraspecies variability, as well as laboratory‐to‐

field variability.
Page 44 of 87

   Table 3‐3. Aquatic Toxicity Data for Sb2O3 and SbCl3 – Acute Toxicity

Test Organism

Fresh/

Salt

Water

Test

Guideline/

Study Type

Duration

(hr‐hour)

End‐

point

Concentration

(mg/L)

Antimony

Compound Effect Reference

Fish ‐ Saltwater

Red Sea Bream

Pagrus major Salt a Static 96‐hr LC50 6.9 SbCl3 Mortality

Takayanagi,

2001

Aquatic Invertebrates ‐ Freshwater

Hydra

Chlorohydra

viridissima

Fresh a Static 96‐hr LC50 1.77 SbCl3 Mortality TAI (1990)

Water flea

Daphnia  

magna

Fresh OECD TG 211

Semi‐static

21‐day

NOEC 1.74

SbCl3

Reproduction

Heijerick and

Vangheluwe

(2003b)

LOEC 3.13

MATC 2.33

Aquatic Plants ‐ Freshwater

Algae, Pseudo‐

kirchneriella

subcapitata

Fresh

OECD 201

Static 72‐hr

EC50 > 36.6

SbCl3 Growth

Heijerick and

Vangheluwe

(2004)

NOEC 2.11

LOEC 4.00

MATC 2.91

Algae, Pseudo‐

kirchneriella

subcapitata

Fresh OECD 201

Static

72‐hr

NOEC 0.396

Sb2O3 Growth

LISEC (2001)

LOEC 1.32

MATC 0.723

Note:  The shaded row indicates the principal study used to assess acute risks to aquatic organisms. aGuideline not reported

3.2.2 Chronic Toxicity to Aquatic Organisms

Chronic aquatic toxicity studies considered for this assessment are summarized in Table 3‐4.  

EPA/OPPT evaluated chronic toxicity studies with Sb2O3 and SbCl3 in fish and aquatic

invertebrates. No reliable chronic toxicity studies were identified for aquatic plants; therefore,

acute toxicity studies were considered for this endpoint.  

In a 30‐day chronic toxicity study with SbCl3 in the fathead minnow (Pimephales promelas),

larvae were exposed under continuous flow‐through conditions from pre‐hatch (i.e., 16 to 40

hours post‐fertilization) until 28 days post‐hatch. Growth was determined to be the most

responsive endpoint (Kimball, 1978). The study authors reported a measured MATC value of

1.62 mg Sb/L, which was used to calculate a chronic COC for aquatic organisms in the current

assessment. In a separate 30‐day chronic toxicity study, P. promelas were exposed to Sb2O3 at

measured concentrations of 0 (control) 0 (positive control), 0.62, 0.74, 1.4, 3.2, or 7.5 μg/L of

Sb203 under continuous flow‐through conditions. The study authors reported a no‐observed

effect concentration (NOEC) > 0.0075 mg Sb/L, based on growth (LeBlanc and Dean, 1984). This

study was not selected for the chronic COC calculation due to uncertainties regarding the dose

response relationship (e.g., the lack of observed effects at the highest concentration tested).  
Page 45 of 87

Heijerick and Vangheluwe (2003b) conducted a 21‐day static‐renewal test using SbCl3 in an

aquatic invertebrate (Daphnia magna; conducted according to OECD 211 test guidelines) with

the following measured concentrations at day 0:  < 2 (control), 55.6, 98.0, 180, 312, 577, 984,

1736, 3130, 5860, and 9964 μg Sb/L and at day 21:  < 2 (control), 46.4, 90.5, 166, 300, 561, 948,

1708, 2990, and 5580 μg Sb/L. Test concentrations were renewed three times per week with

one organism per vessel and ten replicates per test concentration. The 21‐day D. magna NOEC,

LOEC and MATC for SbCl3 are 1.74 mg/L, 3.13 mg/L and 2.33 mg/L, respectively. Reproduction

was the hazard endpoint identified in this study, yielding a MATC value of 2.33 mg Sb/L. Though

this study was considered reliable, the data were not selected for COC derivation because the

MATC for P. promelas indicates that of the species evaluated, fish are more sensitive.  

The chronic COC for aquatic organisms was derived from the 30‐day MATC for growth (1.62 mg

Sb/L) determined using SbCl3 in fish (Pimephales promelas; Kimball, 1978). This value was

divided by an uncertainty factor of ten for fish species in accordance with established EPA/OPPT

methods (EPA, 2012; 2013) to yield a chronic COC of 0.162 mg Sb/L (162 μg Sb/L).

Kimball (1978) was also used to determine a predicted no‐effect concentration (PNEC) of 0.113

mg Sb/L in the EU RAR for ATO (EC, 2008) and the Environment Canada/Health Canada (2010)

risk assessment for ATO; the NOEC value (1.13 μg Sb/L) was used to calculate the PNEC for ATO.

Table 3‐4. Aquatic Toxicity Data for Sb2O3 and SbCl3 – Chronic Toxicity

Test  

Organism

Fresh/

Salt Water

Test

Guideline/

Study Type

Duration End‐

point

Concentration

(mg/L)

Antimony

Compound Effect Reference

Fish‐Freshwater

Pimephales

promelas Fresh Flow‐througha 30‐day

NOEC 1.13

SbCl3

Growth

(length)

Kimball

LOEC 2.31 (1978)

MATC 1.62

Pimephales

promelas Fresh Flow‐througha 30‐day NOEC

> 0.0075 Sb2O3 Growth

LeBlanc and

Dean (1984)

Aquatic Invertebrates ‐ Freshwater

Water flea

Daphnia  

magna

Fresh

OECD TG 211

Semi‐static

21‐day NOEC 1.74 SbCl3

Reproduction

Heijerick and

Vangheluwe

(2003b)

Aquatic Plants ‐ Freshwater

Pseudo‐

kirchneriella

subcapitata

Fresh OECD 201

Static 72‐hr

EC50 > 36.6

SbCl3 Growth

Heijerick and

Vangheluwe

(2004)

NOEC 2.11

LOEC 4.00

MATC 2.91

Pseudo‐

kirchneriella

subcapitata

Fresh OECD 201

Static 72‐hr NOEC 0.396 Sb2O3 Growth

LISEC (2001)

Note: The shaded row indicates the principal study used to assess chronic risks to aquatic organisms. aGuideline not reported

   
Page 46 of 87

3.2.3 Toxicity to Sediment‐Dwelling Organisms		

Toxicity studies for sediment‐dwelling organisms considered for this assessment are

summarized in Table 3‐5.  

Toxicity to sediment‐dwelling organisms in freshwater systems has been investigated in three

different species (EC, 2008). These test species included: (1) the bottom‐dwelling amphipod

(Hyalella azteca), which is a surface deposit and filter feeder, (2) the midge larvae (Chironomus

riparius), which burrows and feeds in sediment and on detritus, and (3) the oligochaete

(Lumbriculus variegatus), which feeds in sediment and the interstitial spaces located below the

water‐sediment interface (Heijerick and Vangheluwe, 2003a, 2005 a, 2005b).  

Heijerick and Vangheluwe (2003a) exposed the amphipod, H. azteca to SbCl3 at five test

concentrations (ranging from 44 to 355 mg Sb/kg dw) in a sediment‐water exposure system for

a total of 28 days. Each test concentration included 13 replicates with ten amphipods each.

Four replicates were used to evaluate the 28‐day growth and survival endpoints, and eight

replicates were used to evaluate treatment‐related effects on survival and reproduction on day

35 and 42 (including growth). The remaining replicate was used to measure the nominal

antimony concentration in sediment. After the initial exposure period, the amphipods were

removed from the treated sediment and placed in sediment‐free chambers for another 14 days

(depuration). Effects on survival (day 28, 35, 42), growth (day 28, 42), and reproduction (i.e.,

the number of offspring per female from day 28 to 42) were subsequently evaluated. Of the

monitored effects, growth (i.e., weight and length) was identified as the most responsive

endpoint after 28 days of SbCl3 exposure, resulting in a NOEC of 124 mg Sb/kg dw.

For the toxicity study with the oligochaete (L. variegatus), Heijerick and Vangheluwe (2005a)

exposed adult organisms to SbCl3 in a sediment‐water exposure system for 28 days. Effects on

survival, reproduction, and growth were subsequently evaluated at the end of this exposure

period. The study was performed using six exposure concentrations (23 to 445 mg Sb/kg ww;

33 to 636 mg Sb/kg dw), a control group (< 1.4 mg Sb/kg ww), and seven replicates per test

concentration. Six replicates were used to evaluate effects on survival, reproduction, and

growth after 28 days of exposure; the remaining replicate was used to measure the nominal

antimony concentration in sediment. Of the endpoints that were evaluated, growth (i.e.,

weight) provided the lowest effect level, resulting in a NOEC of 112 mg Sb/kg dw.

Two‐day‐old midge larvae (C. riparius) were exposed to a range of SbCl3 concentrations in a

sediment‐water exposure system until the midge larvae emerged into the adult phase (Heijerick

and Vangheluwe, 2005b). Effects on larval mortality and growth, and subsequent emergence to

midges were determined after 14 and 28 days of exposure. The study was performed with six

concentrations (ranging from 33 to 636 mg Sb/kg dw) and a control group (< 2.0 mg Sb/kg dw).

Eleven replicates were included for each test concentration, five of which were used to

evaluate effects on survival and growth after 14 days of exposure. Another five replicates were

used to determine the extent of midge emergence after 28 days of exposure, and the remaining

replicate was used to measure the nominal antimony concentration in sediment. Of the
Page 47 of 87

endpoints evaluated, growth (i.e., weight), provided the lowest effect level, resulting in a NOEC

of 112 mg Sb/kg dw.

To determine the chronic COC for toxicity to invertebrates in freshwater sediment, the NOEC

for the growth endpoint (78 mg Sb/kg ww) was used as 112 mg Sb/kg dw (based on a 70/100

calculated ww/dw ratio). The effect levels reported in the Heijerick and Vangheluwe (2005a, b)

studies with SbCl3 exposure in the oligochaete (L. variegatus) and midge larvae (C. Riparius),

were divided by an uncertainty factor of 10 to yield 11.2 mg Sb/kg dw as the chronic COC for

sediment‐dwelling organisms. The dry weight (dw) value was used for comparison with the

environmental monitoring data for suspended sediment obtained from the USGS‐NWIS

database (reported as mg Sb/kg dw).

The Heijerick and Vangheluwe (2005a, b) studies were also used in the EU risk assessment to

derive a PNEC value of 11.2 mg/kg dw; the EU derived a PNEC by utilizing the lowest NOEC

value and then applying an uncertainty factor of 10 (EC, 2008).  

Table 3‐5. Chronic Toxicity Data for Sediment‐Dwelling Invertebrates

Test Organism Test Guideline/

Study Type

Duration End‐

point

Concentration

(mg Sb/kg dw)

Antimony

Compound Effect Reference

Amphipod

Hyalella

azteca

EPA 600/R‐99/064

Section 14:  

Test Method 100.4

Measured

42‐day NOEC 124 SbCl3

Growth

(Length)

Heijerick and

Vangheluwe

(2003a)

Oligochaete

Lumbriculus

variegatus

OECD “Bioaccumulation;

Sediment test using

benthic  oligochaetes”

and EPA 600/R‐99/064

Section 13:

Test Method 100.3

Measured

28‐day NOEC 112 SbCl3

Growth

(Weight)

Heijerick and

Vangheluwe

(2005a)

Midge larvae

Chironomus

riparius

OECD TG 218

Measured

28‐day NOEC 112 SbCl3

Growth

(Weight)

Heijerick and

Vangheluwe

(2005b)

Note:  The shaded row indicates the principal study used for assessing chronic risks to sediment‐dwelling

organisms.

3.2.4 Chronic Toxicity to Terrestrial Organisms

The chronic toxicity studies considered for soil‐dwelling organisms are summarized in Table 3‐6.  

EPA/OPPT identified chronic toxicity studies for soil‐dwelling invertebrates. Several studies

were available from secondary sources; however, only one study, by Moser (2007) was

considered to be reliable. This study was described in the following excerpt taken from the EU

RAR (EC, 2008):
Page 48 of 87

“Moser (2007) studied the toxicity of antimony to the springtail, Folsomia candida

(reproduction and mortality) using aged (approximately 32 w) Sb2O3‐amended soil. Ten (10‐12

day old) synchronized springtails were put onto 30 g moist soil in a glass vessel during a four‐

week exposure period. Five replicates were used for each test concentration and the measured

test concentrations were 90, 322, 999, 2930, and 10119 mg Sb/kg soil dw. Additionally, the

springtails were also tested using uncontaminated field soil and an untreated artificial soil as

controls, based on OECD 207 guidelines. The positive control used on springtail reproduction

was the herbicide Betosip (a.i. 159 g/L phenmedipham). At the end of the exposure period, the

number of juveniles in each test vessel was counted after floating. The mortality of the

springtails was also recorded. At day 28, the pH‐value and the moisture of the artificial soil for

each concentration in additional vessels without springtails were determined for both mortality

and reproduction, the highest effect observed was below 50% in any of the treatments.

Consequently, LC50 and EC50 values could not be calculated, but were estimated as >10,119 mg

Sb/kg soil dw for each one of the endpoints. The NOEC for reproduction was approximately 999

mg Sb/kg dw; the LOEC for reproduction was approximately 2,930 mg Sb/kg dw. All reported

values refer to measured concentrations.”

The remaining studies were reviewed, but had deficiencies (or discrepancies) in study design

and/or uncertainties due to inadequate or missing information.  

Because the available exposure data are insufficient to assess ATO risks to soil‐dwelling

organisms, EPA/OPPT did not calculate a COC from the springtail study.

Table 3‐6. Chronic Toxicity Data for Soil‐dwelling Invertebrates

Test  

Organism

Test

Guideline/

Study Type

Duration End‐

point

Concentration

(mg Sb/kg dw) Antimony

Compound Effect Reference

Springtail

Folsomia

candida

ISO 11267

Measured

28‐day NOEC 999

Sb2O3 Reproduction Moser

LOEC 2930 (2007)

Earthworm

Eisenia fetida

ISO

11268‐2

Nominal

21‐day

NOEC 617

Sb2(SO4)3

Adult survival

Simini et

al. (2002)

LOEC 697

NOEC 60 Juvenile

LOEC 86 production

Enchytraeid

Enchytraeus

crypticus

ISO 16387

Nominal

14‐day

NOEC 384

Sb2(SO4)3

Adult survival Kuperman

et al.

(2002)

LOEC 538

28‐day

NOEC 100 Juvenile

LOEC 140 production

Springtail

Folsomia

candida

ISO 11267

Nominal

28‐day

NOEC 100

Sb2(SO4)3

Adult survival

Phillips et

al. (2002)

LOEC 126

NOEC 100 Juvenile  

LOEC 126 production
Page 49 of 87

3.2.5 Conclusions of Environmental Hazard Assessment

This assessment evaluated environmental risks to aquatic and benthic organisms in association

with ATO use as a flame retardant synergist. The hazard characterization for aquatic organisms

was based on an evaluation of acute and chronic toxicity endpoints. The COCs derived from the

studies for aquatic organisms are presented in Table 3‐7. Many of the reported aquatic toxicity

values for ATO were considered unreliable either because they failed to meet EPA’s adequacy

criteria or exceeded the solubility limits of antimony at environmentally relevant pH levels.  For

this reason, all of the hazard benchmarks identified for aquatic organisms were derived from

toxicity studies with antimony trichloride (SbCl3).

Table 3‐7. Concentrations of Concern for Environmental Toxicity

Toxicity Concentration of Concern (COC)

Acute Toxicity, aquatic organisms 354 μg Sb/L

Chronic Toxicity, aquatic organisms 162 μg Sb/L

Acute Toxicity, sediment‐dwelling organisms 112 mg Sb/kg dw

Chronic Toxicity, sediment‐dwelling organisms 11.2 mg Sb/kg dw

The lowest effect level identified for the acute toxicity (mortality) endpoint, an LC50 value of

1.77 mg Sb/L reported in green hydra, Chlorohydra viridissimus, was used to derive the acute

COC (354 μg Sb/L) for aquatic organisms. Reliable chronic toxicity studies were also available for

antimony in fish, daphnia and algae.  Of the aquatic species evaluated, fish appear to be more

sensitive to chronic antimony exposure; the lowest effect level identified for chronic toxicity

(1.62 mg Sb/L), derived from a study on Pimephales promelas, was used to derive chronic COC

(162 μg Sb/L ) for aquatic organisms. Acceptable chronic toxicity values are also available for

three benthic species which exhibit similar hazard profiles; the lowest effect level (112 mg

Sb/kg dw) was identified for the midge, Chironomus riparius and the oligochaete, Lumbriculus

variegatus.  A slightly higher hazard was identified for the amphipod Hyalella azteca (124 mg

Sb/kg dw); impaired growth was determined to be the most sensitive chronic toxicity endpoint

for the aquatic and benthic organisms included in this hazard characterization. The lowest

effect level (112 mg Sb/kg dw) was used to derive the chronic COC (11.2 mg Sb/kg dw) for

benthic organisms.  

3.3 ENVIRONMENTAL RISK CHARACTERIZATION

The goal of the current assessment was to evaluate the environmental risks that may be

associated with ATO use as a flame retardant synergist. Point estimates used to characterize

environmental exposures were based on reported TRI releases and US monitoring data for

water and sediment. These values were subsequently compared to the COCs identified for

these environmental compartments to determine whether this use scenario may present

unacceptable risks to water‐ and sediment‐dwelling organisms.  
Page 50 of 87

The following assessment questions were addressed:

1. Do ATO concentrations measured in US surface waters pose risks for adverse effects in

aquatic organisms?

2. Do ATO concentrations measured in US sediments pose risks for adverse effects in in

benthic organisms?

3.3.1 Summary of TRI Modeling Results

ATO surface water concentrations were predicted using E‐FAST2 to model water releases

reported by selected TRI facilities in 2010. Because TRI does not indicate the number of days

associated with reported releases, two generic release scenarios were developed (i.e., assuming

either 24 or 250 days of release) to provide a range of predicted water concentrations for

comparison with hazard benchmarks (COCs) identified for aquatic organisms.

The Probabilistic Dilution Model (PDM), a module of E‐FAST2 that predicts the number of days a

stream concentration will exceed the designated COC value, was used for both release

scenarios. The estimated 7Q10 concentrations (μg/L) and the predicted days of exceedance (as

predicted by the PDM model), are shown in Table 3‐8. The COC exceedance days refers to the

total number of release days in a given year that the chronic COC would be exceeded using the

surface water concentrations predicted using the E‐FAST2 model.  
Page 51 of 87

Table 3‐8. Summary of E‐FAST2 Results for Facilities Reporting TRI Water Releases

Facility name

NPDES Code 250 days of release 24 days of release 

7Q10

Concentration

(ppb) RQ

7Q10

Concentration

(ppb) RQ

Ex

Eastman Kodak Co TN Eastman Co

Div, Kingsport, TN, 37662

TN0002640 1.15 0.003 11.99 0.03 

Guilford East Plant, Kenansville,NC, 28349

NC0002305

62.07 0.18 674.14 1.9 

Auriga Polymers Inc. Spartanburg

Facility, Spartanburg, 29307 SC0002798 2.29 0.006 23.97 0.07 

Sabic Innovative Plastics US LLC,

Selkirk, NY, 12158‐0145 NY0007072 0.08 0 0.81 0.002 

Sabic Innovative Plastics Mt.

Vernon, LLC, Mount Vernon, IN,

47620

IN0002101 0.02 0 0.25 0.001 

Sage Automotive InteriorsAbbeville Plant, Abbeville, SC,

29620

SC0000353 29.86 0.08 311.55 0.9 

Performance Fibers OperationsInc. Salisbury Plant, Salisbury, NC,

28147

NC0004944

2.83 0.01 30.04 0.09 
Page 52 of 87

Facility name

NPDES Code 250 days of release 24 days of release 

7Q10

Concentration

(ppb) RQ

7Q10

Concentration

(ppb) RQ

Ex

Dupont Teijin Films, Chester, VA,

23836

VA0003077 0.07 0 0.78 0.002 

Haywood Co, Brownsville, TN,

380121797 TN0041939 0.03 0 0.33 0.001 

Bayshore Industrial La Porte, TX,

77571b TX0022799 2.54b 0.01 25.36b 0.07 

Covalence Adhesives, Franklin, KY,

42134

KY0074659 0.69 0.002 7.61 0.02 

General Cable Industries Inc.,

Marion, IN, 46953 IN0002372 0.19 0 2.14 0.01 

Borealis Compounds Inc., Port

Murray, NJ, 07865 NJ0028657 0.09 0 0.96 0.003 

Cerro Wire LLC, Hartselle, AL

35640

AL0054674 0.08 0 0.92 0.003 

a These screening‐ level estimates are based on a range of days of release, which may not represent actual releases and en

concentrations at these facilities

bDilution factor of 1 used as standard practice for sites that discharge into lakes, bays, estuaries and oceans. This facility 

NPDES = National Pollution Discharge Elimination System; NA = Not Applicable
Page 53 of 87

3.3.2 		Summary of Environmental Monitoring Data		

The environmental exposure assessment included a total of 8,500 environmental monitoring

samples for water, sediment, and soil. Although this data could not be associated directly with

ATO production or processing facilities, EPA/OPPT compared this ambient monitoring data to

the ecotoxicity benchmarks derived in this assessment to gain a broader perspective on the

potential risks of environmental antimony concentrations.  

A total of 5,493 water samples were evaluated in the exposure assessment. Five of the 3,643

water samples (~ 0.4 percent) obtained from the USGS‐NWIS database (collected from 1972 to

2012), exceeded the chronic COC for water (162 μg/L); seven of the 1,850 water samples (~ 0.4

percent) obtained from EPA’s STORET database (collected from 2002 to 2010), exceeded the

chronic COC for water (See Appendix C). The average antimony concentration reported for the

3,643 water samples obtained from the USGS‐NWIS dataset was 4 μg/L, which is roughly 40

times lower than the COC for aquatic organisms.

A total of 2,985 sediment samples were evaluated in the exposure assessment. Fourteen of the

1,753 sediment samples (~ 0.9 percent) obtained from the USGS‐NWIS database (collected

from 1987 to 2011), exceeded the chronic COC for sediment (11.2 mg/kg dw) and six of the

1,222 sediment samples (~ 0.5 percent) obtained from the STORET database (collected from

2002 to 2010), exceeded the chronic COC for sediment (11.2 mg/kg dw; See Appendix C). There

were 1,753 sediment samples in the USGS‐NWIS dataset with an average antimony

concentration of 1.3 mg/kg dw, which is approximately ten times lower than the chronic COC

for sediment‐dwelling organisms.

Only 32 soil samples were retrieved from the USGS‐NWIS database between 2005 and 2010; all

reported antimony concentrations were less than the minimum reporting levels for soil (i.e., 1.2

to 1.7 mg/kg); no soil samples were available from the STORET database.  

3.3.3 Conclusions of Risk Characterization

In summary, modeling results obtained for the TRI facilities evaluated under the scope of this

assessment do not indicate risks to aquatic or benthic organisms. No exceedances were

predicted using the 250 days/yr release scenario; antimony stream concentrations predicted by

the EFAST‐2 model using this release scenario were from two to 100,000 times lower than the

COCs selected to be protective of organisms that dwell in the aquatic compartment. Stream

concentrations predicted using the more conservative (i.e., 24 days/yr) release scenario were

also well below the COCs identified for aquatic organisms for most sites; only one exceedance

of the acute and chronic COC was predicted. Because the number of exceedance days predicted

by the model (i.e., 3 and 6 days, respectively) is less than the time required to trigger a chronic

risk concern (i.e., 20 days), significant risks are not anticipated (EPA, 2012d).  
Page 54 of 87

EPA collected environmental monitoring data from two national datasets (USGS‐NWIS and

STORET) to assess environmental risks that may be associated with ATO use as a synergist in

halogenated flame retardants. Although this exposure data cannot be directly attributed to ATO

production or processing facilities, the analysis showed very few instances where antimony

concentrations measured in the ambient environment exceeded (or came near) the

conservative hazard benchmarks identified to be protective of aquatic and benthic organisms.

Less than 0.01 percent of the environmental monitoring samples collected from the ten states

covered under the scope of this assessment showed antimony concentrations that exceeded

the COC for water or sediment.  No exceedances were identified within the last three years (i.e.,

from 2009 to 2012) for any environmental compartment evaluated in this assessment.  

  

These results must be considered within the context of the conservative assumptions,

uncertainties and data limitations that underpin this assessment, described briefly below.   

3.3.4 Uncertainty and Data Limitations		

The primary sources of uncertainty identified in this assessment are related to limitations in the

available exposure and hazard data, the E‐FAST2 model, and the parameters that support them,

as discussed below.  

3.3.4.1 Limitations of the Exposure and Release Data

The environmental fate and transport of antimony is complex, and can be influenced by site‐

specific environmental conditions. Surface environments may be enriched with antimony as a

result of natural or anthropogenic events and over time, this material may be transformed in

ways that can significantly impact the availability and potential toxicity to living organisms.

Because environmental systems are dynamic, and may be influenced by site‐specific

parameters that exhibit significant spatial and temporal variation, it is difficult to extrapolate

the current findings to other sites or geographic regions within the US.  

The USGS‐NWIS database is one of the largest environmental monitoring databases in the US;

however, comprehensive information needed for data interpretation is not always readily

available. In some instances, proprietary information may be withheld, or specific details

regarding analytical techniques may be unclear, or not reported at all.  As a result of all of these

shortcomings, there are uncertainties in the reported data that are difficult to quantify with

regard to impacts on exposure estimates.

The quality of the data provided in the USGS‐NWIS and STORET datasets varies, and some of

the information provided is non‐quantitative.  While a large number of individual sampling

results were obtained from these datasets, the monitoring studies used to collect the data were

not specifically designed to evaluate ATO distribution across the US. The available data

represent a variety of discrete locations and time periods; therefore it is unclear whether the

data are representative of other locations in the US; however, this limitation does not diminish

the overall findings reported in this assessment, as the exposure data showed very few
Page 55 of 87

instances (i.e., less than 0.01 percent) where measured or predicted antimony levels in the

ambient environment exceeded the identified hazard benchmarks for water or sediment‐

dwelling organisms. It is also important to note that the USGS‐NWIS and STORET data did not

align with the primary watersheds of the TRI facilities identified under the scope of this

assessment, therefore no direct correlation can be made between them.  

The TRI database represents a comprehensive source of environmental release data for the US;

however, there are a few caveats to consider when using this information. TRI data are self‐

reported by facilities subject to minimum reporting thresholds; therefore, it may not capture

releases from smaller facilities (i.e., environmental releases may be underestimated). Some of

the reported information may be inaccurate because it reflects approximations based on mass

balances and emission factors, rather than actual emissions or release data. Use of TRI data to

estimate environmental exposures is constrained by a number of uncertainties  including: the

heterogeneity of processes and releases among facilities grouped within a given sector;

assumptions made regarding sector definitions used to select TRI facilities covered under the

scope; lack of information regarding the distribution of specific antimony compounds reported

in TRI; and fluctuations in the level of production and associated environmental releases

incurred as a result of changes in standard operating procedures. Uncertainty may also arise

from omissions in the TRI reporting data, such as sectors that are not required to report,

facilities that fall below the reporting threshold, or facilities for which forms simply are not

filed. While these limitations are important, their impact on estimating exposure potential may

be less than that associated with the conservative assumptions made regarding environmental

releases. ATO is reported in TRI under the broader category of ‘antimony compounds’, which

covers all chemicals that contain antimony. As such, EPA/OPPT expects the actual releases of

ATO to be significantly lower than those reported in TRI.  

3.3.4.2 E‐FAST2 Model Uncertainties

The major limitations associated with use of the E‐FAST2 model relate to the assumptions made

regarding missing information that was required for model input, including the pattern of TRI

releases and the types of antimony compounds that are reflected in reported TRI releases. It is

important to note however, that these limitations are unlikely to change the stated conclusions

of this assessment, because they are based on a series of conservative assumptions that likely

overestimate exposure potential.  

3.3.4.3 Limitations of the Hazard Data		

The laboratory studies used to identify COCs for ecological organisms varied in terms of the

analytical methods used to measure antimony concentrations and the form of antimony used

to assess toxicity. For example, aquatic COCs were derived from studies using SbCl3, whereas

the data used to assess ecological exposures reflect not only those associated with ATO, but all

antimony‐containing compounds, which likely overestimates exposure potential. Risks to soil‐

dwelling organisms were not evaluated; however, antimony is expected to present minimal

concerns due to limited bioavailability in this environmental compartment.  
Page 56 of 87

3.4 CONCLUSIONS

This assessment used TRI data US monitoring data to characterize environmental exposures.  

There are very few reliable studies of antimony toxicity to terrestrial organisms. EPA (2005)

reviewed the available hazard and exposure information and determined that it was

inadequate for the development of an Ecological Soil Screening Level (Eco‐SSL) for antimony.

Due to the relative paucity of reliable hazard and exposure data for risk estimation, a weight of

evidence approach was used to support the assumption of minimal risks to soil‐dwelling

organisms.

Despite the data limitations and uncertainties described above, significant risks to water‐ or

sediment‐dwelling organisms are not expected based on the following considerations:   

- Use of antimony compounds as a surrogate for ATO and modeled exposure estimates

based on the conservative assumption of total releases (during the 2010 TRI reporting

cycle) occurring over a 24‐day period likely overestimate exposure potential.  

- Measured antimony concentrations from environmental monitoring samples collected

within the last three years showed no exceedance of any of the calculated hazard

benchmarks.

   
Page 57 of 87

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Chemical Structure. Environmental Toxicology and Chemistry, 25(11), 2937‐2945.

Ainsworth, N., J. A. Cooke, and M. S. Johnson. 1990. Distribution of Antimony in Contaminated

Grassland.  I.  Vegetation and Soils. Environmental Pollution, 65, 65‐77.

AkzoNobel (AkzoNobel Polymer Chemicals). 2012. Flame Retardant Synergists.

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1177‐1185.

TRI (US Environmental Protection Agency). 2010. Toxic Release Inventory (TRI) Program.

Washington, DC. http://www.epa.gov/tri/index.htm.

Tschan, M. 2009. Phytoavailability and Plant Uptake of Antimony. (Ph. D dissertation, Diss. ETH

No. 18656), Dipl Biol. Universite de Neuchâtel, Neuchâtel, Switzerland.

Tweats, D. J., D. Blakey, R. H. Heflich, A. Jacobs, S. D. Jacobsen, T. Morita, T. Nohmi, M. R.

O'Donovan, Y. F. Sasaki, T. Sofuni, and R. Tice. 2007a. Report of the IWGT Working

Group on Strategies and Interpretation of Regulatory in Vivo Tests I.  Increases in
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Micronucleated Bone Marrow Cells in Rodents That Do Not Indicate Genotoxic Hazards.

Mutation Research, 627, 78‐91.

Tweats, D. J., D. Blakey, R. H. Heflich, A. Jacobs, S. D. Jacobsen, T. Morita, T. Nohmi, M. R.

O'Donovan, Y. F. Sasaki, T. Sofuni, and R. Tice. 2007b. Report of the IWGT Working

Group on Strategies and Interpretation of Regulatory in Vivo Tests II.  Identification of in

Vivo‐Only Positive Compounds in the Bone of Marrow Micronucleus Test. Mutation

Research, 627, 92‐105.

USAC (U. S. Antimony Corporation). 2012. Antimony Oxide Types.

http://www.usantimony.com/antimony_oxides.htm (November 7, 2012).

USGS (US Geological Survey). 1999. New Reporting Procedures Base on Long‐Term Method

Detection Levels and Some Considerations for Interpretations of Water‐Quality Data

Provided by the US Geological Survey National Water Quality Laboratory. OFR 99‐193.

Reston, VA.

USGS (US Geological Survey). 2004. Mineral Commodity Profiles Antimony. Open‐File Report 03‐

019. US Department of the Interior. http://pubs.usgs.gov/of/2003/of03‐019/of03‐

019.pdf.

USGS (US Geological Survey). 2006. 2005 Minerals Yearbook  Antimony. US Department of the

Interior.

http://minerals.usgs.gov/minerals/pubs/commodity/antimony/antimmyb05.pdf.

USGS (U.S Geological Survey). 2011. 2010 Minerals Yearbook.  Antimony [Advanced Release].

http://minerals.usgs.gov/minerals/pubs/commodity/antimony/myb1‐2010‐antim.pdf.

USGS (US Geological Survey). 2012a. Antimony. Mineral Commodity Summary.

http://minerals.usgs.gov/minerals/pubs/commodity/antimony/mcs‐2012‐antim.pdf.

USGS (US Geological Survey). 2012b. Water Quality Samples for the Nation.  National Water

Information System: Web Interface. US National Geographic ‐ National Water

Information System. http://nwis.waterdata.usgs.gov/usa/nwis/qwdata.

USGS (US Geological Survey). 2014. Geochemical and Mineralogical Maps for Soils of the

Conterminous United States. Open‐file report 2014‐1082. US Department of the Interior.

http://pubs.usgs.gov/of/2014/1082/pdf/ofr2014‐1082.pdf.

Watt, W. D. 1983. Chronic Inhalation Toxicity of Antimony Trioxide: Validation of the Threshold

Limit Value. (Ph.D. dissertation), Wayne State University, Detroit, MI. (as cited in EC,

2008).
Page 69 of 87

Weil, E., and S. Levchik. 2009. Flame Retardants for Plastics and Textiles: Practical Applications.

Munich, Germany: Hanser Publications.

WHO (World Health Organization). 2003. Antimony in Drinking‐Water. Background Document

for Development of WHO Guidelines for Drinking‐Water Quality.

www.who.int/water_sanitation_health/dwq/chemicals/antimony.pdf.

   
Page 70 of 87

APPENDICES

   
Page 71 of 87

Appendix A HUMAN HEALTH HAZARD SUMMARY

A‐1 TOXICOKINETICS

Precise quantitative information regarding antimony absorption, distribution, and retention in

humans is lacking; however, increased levels of antimony have been reported in blood and

urine obtained from workers following ATO inhalation exposure, indicating that trivalent

antimony may be absorbed from the lungs and excreted in the urine (Bailly et al., 1991). A renal

elimination half‐life of four days was estimated in 21 employees of a starter battery

manufacturing plant who were exposed to ATO via inhalation (Kentner et al., 1995).  

Systemic absorption of antimony compounds following inhalation exposure is mainly influenced

by chemical form, particle size, and solubility. In general, large particles (≥ 2.5 μm) tend to

deposit in the upper airway and are cleared more rapidly (via mucociliary transport) than

smaller particles that can deposit in the lower lung, where clearance times depend more on

solubility (i.e., less soluble particles will be cleared more slowly) (Brain et al., 1994).  

Antimony may be retained in the lungs for long periods following inhalation exposure. Lung

retention times measured in seven male workers accidentally exposed to radiolabeled aerosols

of ATO (

125Sb) showed that most of the radiolabeled material was confined to the lungs, with

negligible amounts detected in other parts of the body. Monitoring results based on personal

and stationary air sampling methods from glass‐making factories showed measured antimony

concentrations in air between < 5 and 840 μg/m3 (Lüdersdorf et al., 1987). The estimated

elimination half‐life for antimony varied significantly, ranging from 600 to 1,100 days in non‐

smokers and 1,700 to 3,700 days in smokers, suggesting that ATO lung clearance may be

negatively impacted by smoking (Garg et al., 2003). Deceased smelter and refinery workers

exposed to ATO primarily via inhalation, exhibited elevated lung concentrations when

compared to non‐occupationally exposed individuals (Gerhardsson et al., 1982); however,

antimony concentrations measured in other organs (e.g., liver and kidney) were comparable,  

suggesting either a short biological half‐life or limited systemic distribution for antimony in

these tissues.

Laboratory animal data indicate low systemic absorption of ATO when administered via the oral

route (De Bie and  Salmon‐te Rietstap, 2005; Felicetti et al., 1974). Therapeutic use of

antimonial compounds for the treatment of parasitic diseases has also demonstrated poor

systemic bioavailability following oral exposure in humans (WHO, 2003). Based on results

reported by the International Commission on Radiological Protection, limited systemic

absorption (~1%) may be assumed when estimating ATO uptake from the gastrointestinal tract

(as cited in EC, 2008); dermal absorption (0.26 %) is negligible (Roper and  Stupart, 2006). It has

been suggested that a portion of systemic absorption observed following inhalation exposure is

due to the clearance of large particles (via mucociliary transport) and subsequent uptake into

the gastrointestinal tract (Brain et al., 1994). Using the Multiple Path Particle Deposition model
Page 72 of 87

and predicted deposition patterns for inhaled ATO particles (4 μm), the total systemic

absorption (reflecting combined uptake in the lungs and gastrointestinal tract) was estimated at

approximately 7% (as cited in EC, 2008).  

Once taken up into the systemic circulation, trivalent antimony accumulates in red blood cells

and is distributed primarily to liver, spleen, and bone (Casals, 1972). It has been suggested that

this pattern of distribution may reflect clearance from circulation via phagocytic macrophages

located in these organs which effectively sequester non‐endogenous molecules (Sarin, 2010).  

A‐2 ACUTE TOXICITY

The available laboratory animal data indicate low acute toxicity via oral, dermal, and inhalation

routes (as cited in EC, 2008):

• oral LD50 >20,000 mg/kg‐bw (rats)

• dermal LD50 >8,300 mg/kg‐bw (rabbits)

• inhalation LC50 (4‐hour) >5,200 mg/m3 (rats)  

A‐2‐1 Irritation

No adverse skin reactions have been reported in laboratory animals; however, workers have

reported skin irritation following dermal exposure to ATO under conditions of intense heat and

heavy sweating (NIOSH, 1978; Potkonjak and  Pavlovich, 1983). It has been suggested that ATO

may induce skin rashes (i.e., antimony spots) as a result of physical obstruction of the sweat

ducts. Lung and eye irritation have also been noted with occupational exposure, but no causal

relationship has been established. Although ATO is not classified as an irritant under the

Registration, Evaluation, Authorization and Restriction of Chemicals, workers are encouraged to

avoid excessive contact via use of personal protective equipment.  

A‐3 SUBCHRONIC TOXICITY		

Repeated inhalation exposure has been shown to cause respiratory irritation in rodents (Groth

et al., 1986; Newton et al., 1994; Potkonjak and  Vishnjich, 1983). Although detailed

occupational exposure data are lacking in the US, early case reports of adverse pulmonary

effects (inflammation, pneumoconiosis, emphysema) and decreased lung function (Potkonjak

and  Vishnjich, 1983) indicate a potential for adverse respiratory effects in humans following

repeated inhalation exposure. In addition, gastrointestinal disorders (vomiting, diarrhea),

fatigue, dizziness, and headaches have been reported in workers heavily exposed to smelter

fumes (NIOSH, 1978). It is important to note however, that working conditions have improved

significantly over the past 30 years; a permissible exposure limit of 0.5 mg/m3 has been set to

protect this population.  
Page 73 of 87

A‐4 REPRODUCTIVE TOXICITY

A relative paucity of data exists regarding the potential for ATO‐induced reproductive or

developmental effects; however, a weight‐of‐evidence approach may be used to address these

endpoints. Toxicokinetic considerations (low systemic absorption), together with the absence

of effects on reproductive organs in repeated‐dose studies suggest low concern for

reproductive effects.

Limited information is available from a Russian report on ATO (Belyaeva, 1967), which indicated

adverse reproductive effects in rats (e.g., ovarian cysts, uterine metaplasia, decreased fertility)

following inhalation exposure at 250 mg/m3 for 4 hours/day during a two‐month period. These

study authors also reported adverse reproductive outcomes (spontaneous abortions,

premature births, gynecologic problems) in women following exposure to metallic dusts (of

antimony, ATO, and antimony pentasulfide) at unspecified concentrations while working in a

metallurgical plant. The reported results are largely inconclusive and inadequate for risk

assessment due to inadequate or missing information regarding study design (e.g., lack of

information on experimental conditions, controls, and purity of test substance) and exposure

conditions.

No testicular effects were observed in male rats or mice following repeated gavage exposure to

ATO at doses ≥1,000 mg/kg‐bw/day (Gurnani et al., 1992; Omura et al., 2002). In addition, oral

exposure to ATO during a 90‐day feeding in rats study showed no effects on male reproductive

organs up to the highest dose tested (1,686 mg/kg‐bw/day; Hext et al., 1999). Gerber et al.

(1982) used a readily soluble, radiolabeled compound, antimony trichloride (

125SbCl3), to

determine the extent of uptake in pregnant mice following dietary exposure to a mixture of

antimony chloride and arsenic throughout pregnancy. Minimal absorption (~7 %) was observed

in treated dams, with limited distribution to reproductive organs (~0.1 to 0.2 % in ovaries and

uteri). This finding is supported by another prenatal developmental toxicity study, in pregnant

ewes dosed orally throughout gestation with antimony tartrate (a highly soluble form of

antimony expected to exhibit higher absorption as compared to ATO) at 2 mg/kg‐bw/day  

showed no adverse maternal or developmental effects (James et al., 1966).

A‐5 DEVELOPMENTAL TOXICITY

In a prenatal developmental toxicity study, female Sprague‐Dawley rats (26/dose) were

exposed to ATO via (nose‐only) inhalation at 0, 2.6, 4.4, and 6.3 mg/m3 (mass median

aerodynamic diameter [MMAD] = 1.59 to 1.82 μm) for 6 hours/day from the day of fertilization

through day 19 of gestation. Significant increases in lung weights and an accumulation of ATO

in erythrocytes were observed in all treated dams as compared to controls. In addition,

histological examination revealed an accumulation of pigmented alveolar macrophages, and

discrete areas of acute inflammation and hyperplasia in the lungs. An increase in post‐

implantation loss (3.4, 3.6, 3.5, and 7.1 % in the 0, 2.6, 4.4, and 6.3 mg/m3 groups; p = 0.11) was

observed at the highest concentration tested; however, these changes were within the range of
Page 74 of 87

historical controls and not statistically significant. No developmental effects were reported; the

no‐observed‐adverse‐effect concentration (NOAEC) for developmental toxicity was determined

to be > 6.3 mg/m3. A NOAEC for maternal toxicity could not be established, as increased lung

weight was observed at all levels of exposure (Schroeder, 2003)

A‐6 MUTAGENICITY

The available in vitro data for ATO‐induced mutagenicity is mixed. ATO was not mutagenic

when tested in Salmonella typhimurium strains TA98, TA100, TA1535, and TA1537 or

Escherichia coli WP2P and WP2P uvrA with or without metabolic activation; however, ATO

induced chromosomal aberrations in cultured human leukocytes and lymphocytes (Elliott et al.,

1998). No mutagenic potential was observed in mouse lymphoma cells (Paton and  Allison,

1972), but weak induction levels (~ two‐fold above background) were reported in a sister

chromatid exchange assay using V79 cells in vitro (Kuroda et al., 1991). Conflicting results have

also been obtained for ATO‐induced mutagenicity in vivo. Mice gavaged with aqueous ATO at

400, 668, and 1,000 mg/kg‐bw/day for up to 21 days were positive for chromosomal

aberrations, but not clastogenic effects (Gurnani et al., 1992). A mouse micronucleus test

conducted with ATO showed no clastogenic effects following one high dose (5,000 mg/kg‐bw)

or repeated exposure at 400, 668, or 1,000 mg/kg‐bw/day for 7, 14, or 21 days (Elliott et al.,

1998). Negative results were also reported for micronuclei induction in rats following oral

exposure at 250, 500, or 1,000 mg/kg‐bw/day for 21 days (Kirkland et al., 2007).  

Limited evidence of mutagenic potential has been demonstrated for ATO; however, a number

of technical experts have independently concluded that ATO is unlikely to be mutagenic in vivo

(EC/HC, 2010; European Food Safety Authority, 2004; Tweats et al., 2007a, 2007b). Since a

possible local genotoxic effect of ATO would only be relevant under exposure conditions that

also produce lung overload, minimal concerns are indicated for humans (EC, 2008).  

A‐7 CARCINOGENICITY

The International Agency for Research on Cancer (IARC) concluded that “there is sufficient

evidence for the carcinogenicity of ATO in experimental animals" (IARC, 1989) based on one

published, and one unpublished report of increased lung tumors in female rats following ATO

inhalation exposure (Groth et al., 1986; Watt, 1983). Another chronic inhalation study has since

been published that found no increased cancer incidence following ATO inhalation exposure in

rats (Newton et al., 1994).

The proliferation of alveolar macrophages is generally regarded as a normal physiological

response to the deposition of poorly soluble particulates in the lung; however, in rats

repeatedly exposed at high levels, macrophage clearance is impaired, resulting in “lung

overload”. This condition is characterized by macrophage infiltration, chronic interstitial

inflammation, metaplasia, and the subsequent development of lung tumors. Although the

cascade of effects caused by pulmonary overload are unlikely to be relevant for humans, the rat
Page 75 of 87

is assumed to be an appropriate model for both neoplastic and non‐neoplastic responses to

poorly soluble particulates (Mauderly, 1997). Because the responsiveness of the rat model to

non‐genotoxic agents (under conditions of lung overload) is dependent upon chronic active

inflammation and cell proliferation, no lung cancer hazard is anticipated for humans exposed to

ATO at lower, environmentally relevant concentrations, where chronic active inflammation and

cell proliferation are not present (International Life Sciences Institute Workshop Participants,

2000).

IARC concluded that “there is inadequate evidence for the carcinogenicity of ATO in humans”

(IARC, 1989). Since this evaluation, two epidemiological studies of increased cancer incidence in

smelter workers have been published.

A mortality study conducted in 1,014 male workers employed in an antimony smelter plant

between 1937 and 1971 showed a significant excess in mortality from lung cancer, ischemic

heart disease and non‐malignant respiratory disease; however, the authors acknowledged that

the observed increase in cancer mortality may be influenced by co‐exposure to arsenic (Schnorr

et al., 1995).

A survey of antimony smelter workers employed before 1961 showed a significant increase in

lung cancer mortality (12 expected vs. 5.3 observed, p = 0.016), with a minimum latency period

of approximately 20 years between the first exposure and death from lung cancer (Jones,

1994). No correlation between the length of employment at the smelter and cancer mortality

was found. In addition, no excess cancer mortality was found in the cohort recruited after 1960.

This study is limited by the lack of an appropriate control population, and confounding factors

such as cigarette smoking and co‐exposure to polycyclic aromatic hydrocarbons, arsenic, and

sulfur. As such, the increased cancer mortality observed in the earlier cohort could not be

attributed to a specific agent.  

While it is clear that ATO inhalation exposure can impair pulmonary clearance in rats (Newton

et al., 1994), it is important to note that animal studies tend to employ significantly higher

concentrations and longer durations of exposure than would generally be expected for humans.

Although the available information is limited, there is no evidence of carcinogenicity via the oral

route (Schroeder et al., 1968; Schroeder et al., 1970).

A‐8 SYNOPSIS

Inhalation of airborne dusts is the primary concern for ATO exposure in occupational settings,

whereas the general population is primarily exposed to low levels of ATO in food and water.

Chemical speciation and route of exposure are major determinants of antimony toxicity in

humans. Generally speaking, pentavalent antimony is less toxic than the trivalent species;

however, most antimony compounds are oxidized to ATO upon release into the environment,

and the available evidence suggests that over time, further oxidation to the pentavalent form
Page 76 of 87

may occur under aerobic conditions (EC, 2008). In addition, quantitative studies indicate that

minimal amounts of ATO are released from flame‐retarded consumer products (CPSC, 2006,

2012; Thomas and  Stevens, 2005).  

The general population is exposed mainly via ingestion of food and water, whereas inhalation

and dermal contact are more common in occupational settings. Irritation at the point of contact

(e.g., dermatitis, lung inflammation, pneumoconiosis) has been reported in workers following

chronic inhalation exposure (NIOSH, 1978); however, a permissible exposure limit of 0.5 mg/m3

has been set to protect this population. ATO produces minimal systemic toxicity when

administered via oral, dermal, or inhalation routes due to poor systemic absorption, which

effectively limits bioavailability. Because food and water are the primary sources of general

population exposure, and the pentavalent (i.e., less toxic) form of antimony predominates in

these media, significant health risks are not expected at concentrations to which individuals

may be exposed in the environment (ATSDR, 1992). This conclusion is supported by previous

risk assessments for ATO in Europe and Canada (EC/HC, 2010; EC, 2008).

A‐9 PREVIOUS ASSESSMENTS

A‐9‐1 Environment Canada/Health Canada: Screening Assessment for

the Challenge – ATO (EC/HC, 2010)

(http://www.ec.gc.ca/ese‐ees/9889ABB5‐3396‐435B‐8428‐F270074EA2A7/batch9_1309‐64‐

4_en.pdf)

In Canada, ATO is listed on the Domestic Substances List and is characterized as “inherently

toxic to aquatics,” but did not meet the full categorization criteria and was not recommended

for further assessment. In this assessment, it was concluded:  

“The margins between upper‐bounding estimates of exposure to antimony trioxide from

environmental media (based on antimony) and from use of household items and levels

associated with effects in experimental animals are considered to be adequately protective to

account for uncertainties in the health effects and exposure databases. On the basis of the

adequacy of the margins between conservative estimates of exposure to antimony trioxide and

critical effect levels in experimental animals, it is concluded that antimony trioxide is not

entering the environment in a quantity or concentration or under conditions that constitute or

may constitute a danger in Canada to human life or health.”
Page 77 of 87

A‐9‐2 EU RAR – Diantimony Trioxide (EC, 2008)

(http://esis.jrc.ec.europa.eu/doc/risk_assessment/REPORT/datreport415.pdf)

In Europe, ATO is registered under REACH regulations; the European Commission classifies and

labels ATO as “H351: suspected of causing cancer by inhalation;” therefore, “a need for limiting

the risks” was identified for workers.

The assessment of human health and environmental risks concluded:

“Overall results of the risk assessment for surface water, waste water treatment plants,

terrestrial compartment, atmosphere, secondary poisoning, marine water, marine sediment,

secondary poisoning in the marine environment:  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.”

A‐9‐3 OECD Assessment for Antimony Trioxide (OECD, 2008)

(http://webnet.oecd.org/Hpv/UI/handler.axd?id=13e93c97‐6605‐4eac‐961f‐8af23cc6ad32)

A‐9‐3‐1 Environment

“This chemical is currently of low priority for further work.”

A‐9‐3‐2 Human Health

“Despite the poor quality of the available acute oral and dermal toxicity studies with

diantimony trioxide in laboratory animals, diantimony trioxide is considered to be of low acute

inhalation, oral and dermal toxicity. The chemical possesses properties indicating a hazard for

human health (skin irritation, lung toxicity and lung carcinogenicity after repeated exposure).

Member countries are invited to perform an exposure assessment for workers and if then

indicated, a risk assessment. As all studies on chronic toxicity/carcinogenicity deviate from the

OECD guideline and because of other critical shortcomings in the data set on chronic

toxicity/carcinogenicity, the US NTP (National Toxicology Program) has initiated 2‐year

inhalation toxicology and carcinogenicity studies on diantimony trioxide in rats and mice.”

(http://ntp.niehs.nih.gov/?objectid=BCD52E6A‐123F‐7908‐7BD5298A5D912AF9)

   
Page 78 of 87

Appendix B EFAST2 MODEL INPUTS USED WITH TRI DATATRI does not require facilities to report the number of days associated with reported releases, therefore two re(assuming 24 or 250 release days per year) were developed to provide an upper and lower bound for the range 

concentrations predicted by the EFAST2 model. Typically one would enter a removal rate associated with waste 

but in regard to the TRI data, the amount listed is the self‐reported amount released into water and a removal 

therefore 0% is used. One of the cases below was a facility that released into a bay (Galveston). Additional inpu

types of cases, therefore the table for Galveston Bay is different from the others.

FACILITY NAME: TN EASTMAN CO KINGSPORT

FACILITY LOCATION: KINGSPORT TN 37660RECEIVING STREAM NAME: HOLSTON R S FK

REACH NUMBER: 06010102004 FACILITY ON REACH: Yes DISCHARGE TYPE: Direct

NPDES PERMIT NUMBER: TN0002640 GAGING STATION ID: N/A

GAGING STATIONPERIOD OF RECORD:N/A

GAGING STATIONNUMBER OF OBSERVATIONS: N/A

NUMBER OF STATIONS ON REACH: N/

MEAN FLOW (MLD): 6343.16 7q10 FLOW (MLD): 1937.37 EFFLUENT FLOW (MLD):1937.37

FACILITY NAME: GUILFORD MILLS INC/GUIL. EAST

FACILITY LOCATION: KENANSVILLE NC 28349RECEIVING STREAM NAME: NE CAPE FEAR R

REACH NUMBER: 03030007009 FACILITY ON REACH: Yes DISCHARGE TYPE: Direct

NPDES PERMIT NUMBER: NC0002305 GAGING STATION ID: N/A
Page 79 of 87

FACILITY NAME: GUILFORD MILLS INC/GUIL. EAST

GAGING STATIONPERIOD OF RECORD:N/A

GAGING STATIONNUMBER OF OBSERVATIONS: N/A

NUMBER OF STATIONS ON REACH: N/A

MEAN FLOW (MLD): 1041.00 7q10 FLOW (MLD): 23.04 EFFLUENT FLOW (MLD):2.70

FACILITY NAME: ARTEVA SPECIALTIES D/B/A KOSA

FACILITY LOCATION: SPARTANBURG SC 29307RECEIVING STREAM NAME: PACOLET R

REACH NUMBER: 03050105066 FACILITY ON REACH: No DISCHARGE TYPE: Direct

NPDES PERMIT NUMBER: SC0002798 GAGING STATION ID: N/A

GAGING STATIONPERIOD OF RECORD:N/A

GAGING STATIONNUMBER OF OBSERVATIONS: N/A

NUMBER OF STATIONS ON REACH: N/A

MEAN FLOW (MLD): 1797.77 7q10 FLOW (MLD): 297.48 EFFLUENT FLOW (MLD):3.50

FACILITY NAME: G E PLASTICS ‐ SELKIRK OPER

FACILITY LOCATION: SELKIRK NY 12158RECEIVING STREAM NAME: HUDSON R

REACH NUMBER: 02020006041 FACILITY ON REACH: Yes DISCHARGE TYPE: Direct

NPDES PERMIT NUMBER: NY0007072 GAGING STATION ID: N/A

GAGING STATIONPERIOD OF RECORD:N/A

GAGING STATIONNUMBER OF OBSERVATIONS: N/A

NUMBER OF STATIONS ON REACH: N/A

MEAN FLOW (MLD): 3.65E+04 7q10 FLOW (MLD): 6866.81 EFFLUENT FLOW (MLD):7.38
Page 80 of 87

FACILITY NAME: MILLIKEN/ABBEVILLE PLANT

FACILITY LOCATION: ABBEVILLE SC 29620RECEIVING STREAM NAME: LONG CANE CR

REACH NUMBER: 03060103021 FACILITY ON REACH: Yes DISCHARGE TYPE: Direct

NPDES PERMIT NUMBER: SC0000353 GAGING STATION ID: N/A

GAGING STATIONPERIOD OF RECORD:N/A

GAGING STATIONNUMBER OF OBSERVATIONS: N/A

NUMBER OF STATIONS ON REACH: N/A

MEAN FLOW (MLD): 88.42 7q10 FLOW (MLD): 17.75 EFFLUENT FLOW (MLD):2.79

FACILITY NAME: ARTEVA SPECIALTIES – KOSA

FACILITY LOCATION: SALISBURY NC 28144RECEIVING STREAM NAME: SECOND CR

REACH NUMBER: 03040102028 FACILITY ON REACH: Yes DISCHARGE TYPE: Direct

NPDES PERMIT NUMBER: NC0004944 GAGING STATION ID: N/A

GAGING STATIONPERIOD OF RECORD:N/A

GAGING STATIONNUMBER OF OBSERVATIONS: N/A

NUMBER OF STATIONS ON REACH: N/A

MEAN FLOW (MLD): 727.70 7q10 FLOW (MLD): 70.58 EFFLUENT FLOW (MLD):6.29

FACILITY NAME: DU PONT TEIJIN FILMS

FACILITY LOCATION: HOPEWELL VA 23860RECEIVING STREAM NAME: JAMES R

REACH NUMBER: 02080206045 FACILITY ON REACH: Yes DISCHARGE TYPE: Direct

NPDES PERMIT NUMBER: VA0003077 GAGING STATION ID: N/A
Page 81 of 87

FACILITY NAME: MILLIKEN/ABBEVILLE PLANT

GAGING STATIONPERIOD OF RECORD:N/A

GAGING STATIONNUMBER OF OBSERVATIONS: N/A

NUMBER OF STATIONS ON REACH: N/A

MEAN FLOW (MLD): 1.85E+04 7q10 FLOW (MLD): 1468.34 EFFLUENT FLOW (MLD):0.68

FACILITY NAME: HAYWOOD CO‐BROWNSVILLE

FACILITY LOCATION: BROWNSVILLE TN 38012RECEIVING STREAM NAME: HATCHIE R

REACH NUMBER: 08010208012 FACILITY ON REACH: No DISCHARGE TYPE: Direct

NPDES PERMIT NUMBER: TN0041939 GAGING STATION ID: N/A

GAGING STATIONPERIOD OF RECORD:N/A

GAGING STATIONNUMBER OF OBSERVATIONS: N/A

NUMBER OF STATIONS ON REACH: N/A

MEAN FLOW (MLD): 5709.76 7q10 FLOW (MLD): 1902.97 EFFLUENT FLOW (MLD):15.07
Page 82 of 87

FACILITY NAME: CITY OF LA PORTE

FACILITY LOCATION: LA PORTE TX 775721115RECEIVING WATER NAME: GALVESTON BAY

REACH NUMBER: 12040204001 FACILITY ON REACH: No DISCHARGE TYPE: Direct

NPDES PERMIT #: TX0022799 EXPOSED POPULATION: Adult

WWT REMOVAL

(%)

RELEASE DAYS PLANT FLOW

(MLD)

PRETREATMENT

RELEASE

(kg/site/day)

POST‐TREATMENT

RELEASE (kg/site/day) BCF

(L/kg)

0. 250. 11.83 0.03 3.00E‐02 0.00

AQUATIC EXPOSURE ESTIMATESMIXING ZONE DILUTION FACTOR WATER CONC (μg/l)ACUTE SCENARIO 1.00 2.54CHRONIC SCENARIO 1.00 2.54

FACILITY NAME: HONEYWELL INTERNATIONAL INC

FACILITY LOCATION: MONCURE NC 27559RECEIVING STREAM NAME: HAW R

REACH NUMBER: 03030002001 FACILITY ON REACH: Yes DISCHARGE TYPE: Direct

NPDES PERMIT NUMBER: NC0001899 GAGING STATION ID: N/A

GAGING STATION GAGING STATION NUMBER OF STATIONS ON REACH: N/A
Page 83 of 87

AQUATIC EXPOSURE ESTIMATES

PERIOD OF RECORD:N/A NUMBER OF OBSERVATIONS: N/A

MEAN FLOW (MLD): 3876.49 7q10 FLOW (MLD): 144.25 EFFLUENT FLOW (MLD):0.87

FACILITY NAME: HARTSELLE CITY OF UTIL BOARD

FACILITY LOCATION: HARTSELLE AL 35640RECEIVING STREAM NAME: FLINT CREEK

REACH NUMBER: 06030002032 FACILITY ON REACH: No DISCHARGE TYPE: Direct

NPDES PERMIT NUMBER: AL0054674 GAGING STATION ID: N/A

GAGING STATIONPERIOD OF RECORD:N/A

GAGING STATIONNUMBER OF OBSERVATIONS: N/A

NUMBER OF STATIONS ON REACH: N/A

MEAN FLOW (MLD): 570.89 7q10 FLOW (MLD): 119.88 EFFLUENT FLOW (MLD):5.16

FACILITY NAME: GENERAL CABLE

FACILITY LOCATION: MARION IN 46953RECEIVING STREAM NAME: MISSISSINEWA R

REACH NUMBER: 05120103010 FACILITY ON REACH: Yes DISCHARGE TYPE: Direct

NPDES PERMIT NUMBER: IN0002372 GAGING STATION ID: N/A

GAGING STATIONPERIOD OF RECORD:N/A

GAGING STATIONNUMBER OF OBSERVATIONS: N/A

NUMBER OF STATIONS ON REACH: N/A

MEAN FLOW (MLD): 1562.64 7q10 FLOW (MLD): 51.52 EFFLUENT FLOW (MLD):0.28
Page 84 of 87

FACILITY NAME: TYCO ADHESIVES

FACILITY LOCATION: FRANKLIN KY 42134RECEIVING STREAM NAME: DRAKES CR W FK

REACH NUMBER: 05110002008 FACILITY ON REACH: No DISCHARGE TYPE: Direct

NPDES PERMIT NUMBER: KY0074659 GAGING STATION ID: 03313700

GAGING STATIONPERIOD OF RECORD:10/01/68 ‐ 09/29/89

GAGING STATIONNUMBER OF OBSERVATIONS: 7670.00

NUMBER OF STATIONS ON REACH: 1

50 %ile FLOW (MLD): 173.66 10 %ile FLOW (MLD): 20.55 EFFLUENT FLOW (MLD):0.82

FACILITY NAME: BOREALIS COMPOUNDS LLC

FACILITY LOCATION: PORT MURRAY NJ 07865RECEIVING STREAM NAME: MUSCONETCONG R

REACH NUMBER: 02040105008 FACILITY ON REACH: Yes DISCHARGE TYPE: DirectNPDES PERMIT NUMBER: NJ0028657 GAGING STATION ID: 01457000

GAGING STATIONPERIOD OF RECORD:01/01/66 ‐ 12/11/90

GAGING STATIONNUMBER OF OBSERVATIONS: 9112.00

NUMBER OF STATIONS ON REACH: 3

50 %ile FLOW (MLD): 503.87 10 %ile FLOW (MLD): 215.25 EFFLUENT FLOW (MLD):0.47
Page 85 of 87

Appendix C SUMMARY OF USGS AND STORET DATATable_Apx C‐1. Summary of All USGS‐NWIS and STORET Data for Exceedances of the Concentrations of Concern per Me

Summary of All USGS‐

NWIS and STORET Data

Sample Type

Parameter Code Date of Sample Value Above COC Station Name Latitude Longitud

Water – USGS‐NWIS 01097:

Antimony,water, filtered,

μg/L

26‐Aug‐87 350c μg/L Red River near Quanah,

TX

34.4131291  ‐99.73453

17‐Feb‐88 350c μg/L Red River near Quanah,

TX

34.4131291  ‐99.73453

25‐Aug‐87 350c μg/L Buck Creek at Loco, TX 34.70283589 ‐

100.1881

25‐Aug‐87 350c μg/L Buck Creek near

Childress, TX

34.61172738 ‐

100.1109

26‐Aug‐87 350c μg/L Salt Fork Red River near

Wellington, TX

4.95755294  ‐100.2209

Water – STORET 10‐Dec‐08 7,600 μg/L Grays Creek, TN 35.200600  ‐89.68150

26‐Oct‐04 2,000 μg/L Beech River, TN 36.605000  ‐88.118803‐Nov‐04 2,000 μg/L Muddy Creek‐Unnamed

Tributary, TN

35.095833  ‐88.49944

18‐Dec‐02 1,000 μg/L Alexander Creek, TN 36.520600  ‐82.7178026‐Oct‐04 1,000 μg/L Big Sandy River, TN 36.232500  ‐88.1003023‐Sep‐04 1,000 μg/L Buffalo River, TN 35.457800  ‐87.7758012‐Oct‐04 1,000 μg/L Clinch River, TN 36.569400  ‐83.04170
Page 86 of 87

Summary of All USGS‐

NWIS and STORET Data

Sample Type

Parameter Code Date of Sample Value Above COC Station Name Latitude Longitud

Sediment – USGS 29816:

Antimony,suspendedsediment, total

digestion, dw

12‐Jan‐88 90 mg/kg Kentucky River at Lock 4

at Frankfort, KY

38.2017388  ‐84.88161

20‐Dec‐89 42 mg/kg Kentucky River at Lock 2

at Lockport, KY

38.43895654 ‐84.96328

01098 26‐Oct‐05 12 mg/kg Lake Mohawk 1.3 miles

upstream of dam near

Sparta, NJ

41.0176111  ‐74.65780

34795 26‐Oct‐05 12 mg/kg Lake Mohawk 1.3 miles

upstream of dam near

Sparta, NJ

41.0176111  ‐74.65780

29816 14‐Apr‐95 26 mg/kg Mohawk River at Cohoes,

NY

42.78535559 ‐73.70761

29816 14‐Apr‐95 23 mg/kg Mohawk River at Cohoes,

NY

42.78535559 ‐73.70761

29816 17‐Sep‐96 13 mg/kg St. Lawrence River at

Cornwall Ontario near

Massena, NY

45.0061599  ‐74.79490

29816 26‐Sep‐96 20 mg/kg Rio Grande bl Falcon

Dam, TX

26.5572768  ‐99.16836

29816 21‐Mar‐96 14 mg/kg Rio Grande bl Falcon

Dam, TX

26.5572768  ‐99.16836

29816 01‐May‐96 18 mg/kg Rio Grande at Foster Rh

near Langtry, TX

29.78075694 ‐

101.7559

34795 19‐Dec‐05 20 mg/kg ZK‐58‐27‐7 at Avery Deer

Spring, TX

30.5075  ‐97.74916

29816 29‐May‐96 18 mg/kg Rio Grande at El Paso, TX 31.80288488 ‐06.5408229816 23‐Oct‐96 16 mg/kg Rio Grande at El Paso, TX 31.80288488 ‐06.54082
Page 87 of 87

Summary of All USGS‐

NWIS and STORET Data

Sample Type

Parameter Code Date of Sample Value Above COC Station Name Latitude Longitud

34795 30‐Aug‐03 18 mg/kg AF Plant 4 Ofl No. 4 near

Bomber Road, Fort

Worth, TX

32.77619444 ‐97.45358

Sediment – STORET 21‐Sept‐

0404

88.1mg/kg

West/East Creeks, NJ 39.176000  ‐74.90000

24‐Sept‐0404

57.2mg/kg

Great Sound, NJ 39.009000  ‐74.81500

28‐Sept‐0404

35.8mg/kg

Great Egg Harbor, NJ 39.366000  ‐74.71600

8‐Aug‐005 20.5

mg/kg

Upper NY/NJ Harbor, NY 40.699000  ‐74.02000

3‐Oct‐006 13.7

mg/kg

Raritan Bay, NJ 40.494000  ‐74.28800

3‐Oct‐0606 11.8

mg/kg

Raritan River, NJ 40.503000  ‐74.29200

aUSGS‐NWIS dataset collected was filtered using the following selection criteria:

1st Data selection factor = Site type name (Site_tp_In). All types listed; those in blue bold are the ones selected: Atmosphere; Coastal ‐ 18 water sam

type well; Ditch; Estuary ‐ only sediment samples; Hyporheic‐zone well; Lake, Reservoir, Impoundment ‐ 279 water samples; Outfall ‐ 27 water sam

Spring ‐ 174 water samples; Storm sewer; Stream ‐ 3,151 water samples; Test hole not completed as well; Tunnel, shaft, or mine; Unsaturated zone

application; Water‐distribution system; Water‐use establishment; and Well. 2nd Data selection factor = Media Name (Medium_nm). All types listed; 

ones selected: Animal tissue; Artificial; Bottom material; Core material; Effluent‐ 1 water sample; Groundwater; Interstitial water; Soil; Surface wat

Suspended sediment; and Wet deposition. Data selection factor for Water samples: medium_nm = Surface water and Effluent. Data selection factor 

Medium Name ‐ Bottom Material, Effluent, Surface water (as reported in mg/kg), Suspended sediment, and wet deposition. Data selection factor for 

Soil.bThe USGS‐NWIS dataset contains a field named ‘dqui_nm,’ which is defined as ‘Data Quality Descriptions.’ STORET does not have this field; the

denoted in this column.

c1⁄2 detection/reported value.