EPA/635/R-15/034a 

Public Comment Draft

www.epa.gov/iris 

Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX)

(CASRN 121-82-4)

In Support of Summary Information on the

Integrated Risk Information System (IRIS)

March 2016

NOTICE

This document is a Public Comment draft. This information is distributed solely for the purpose 

of pre-dissemination peer review under applicable information quality guidelines. It has not been 

formally disseminated by EPA. It does not represent and should not be construed to represent any 

Agency determination or policy. It is being circulated for review of its technical accuracy and 

science policy implications.

National Center for Environmental Assessment

Office of Research and Development

U.S. Environmental Protection Agency

Washington, DC
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

DISCLAIMER

This document is a preliminary draft for review purposes only. This information is 

distributed solely for the purpose of pre-dissemination peer review under applicable information 

quality guidelines. It has not been formally disseminated by EPA. It does not represent and should 

not be construed to represent any Agency determination or policy. Mention of trade names or 

commercial products does not constitute endorsement or recommendation for use.

This document is a draft for review purposes only and does not constitute Agency policy.

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CONTENTS

AUTHORS | CONTRIBUTORS | REVIEWERS..................................................................................... ix

PREFACE.......................................................................................................................................... xi

PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS...............................................................................xiv

EXECUTIVE SUMMARY ................................................................................................................ ES-1

LITERATURE SEARCH STRATEGY | STUDY SELECTION AND EVALUATION .................................. LS-1

1. HAZARD IDENTIFICATION....................................................................................................... 1-1

1.1.Overview of Chemical Properties and Toxicokinetics...................................................... 1-1

1.1.1. Chemical Properties ............................................................................................. 1-1

1.1.2. Toxicokinetics....................................................................................................... 1-2

1.1.3. Description of Toxicokinetic Models.................................................................... 1-3

1.2.PRESENTATION AND SYNTHESIS OF EVIDENCE BY ORGAN/SYSTEM............................... 1-3

1.2.1. Nervous System Effects........................................................................................ 1-4

1.2.2. Kidney and Other Urogenital System Effects..................................................... 1-21

1.2.3. Reproductive and Developmental Effects.......................................................... 1-38

1.2.4. Liver Effects........................................................................................................ 1-49

1.2.5. Carcinogenicity ................................................................................................... 1-60

1.2.6. Other Noncancer Effects.................................................................................... 1-69

1.3.INTEGRATION AND EVALUATION .................................................................................. 1-69

1.3.1. Effects Other Than Cancer.................................................................................. 1-69

1.3.2. Carcinogenicity ................................................................................................... 1-72

1.3.3. Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes . 1-73

2. DOSE-RESPONSE ANALYSIS .................................................................................................... 2-1

2.1.ORAL REFERENCE DOSE FOR EFFECTS OTHER THAN CANCER......................................... 2-1

2.1.1. Identification of Studies for Dose-Response Analysis of Selected Effects........... 2-1

2.1.2. Methods of Analysis............................................................................................. 2-5

2.1.3. Derivation of Candidate Values.......................................................................... 2-10

2.1.4. Derivation of Organ/System-Specific Reference Doses..................................... 2-16

2.1.5. Selection of the Overall Reference Dose............................................................ 2-17

2.1.6. Comparison with Mortality LD01s ....................................................................... 2-18

This document is a draft for review purposes only and does not constitute Agency policy.

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2.1.7. Uncertainties in the Derivation of the Reference Dose ..................................... 2-22

2.1.8. Confidence Statement........................................................................................ 2-23

2.1.9. Previous IRIS Assessment................................................................................... 2-24

2.2.INHALATION REFERENCE CONCENTRATION FOR EFFECTS OTHER THAN CANCER........ 2-24

2.2.1. Previous IRIS Assessment................................................................................... 2-25

2.3.ORAL SLOPE FACTOR FOR CANCER................................................................................ 2-25

2.3.1. Analysis of Carcinogenicity Data ........................................................................ 2-25

2.3.2. Dose-Response Analysis—Adjustments and Extrapolation Methods................ 2-26

2.3.3. Derivation of the Oral Slope Factor.................................................................... 2-27

2.3.4. Uncertainties in the Derivation of the Oral Slope Factor................................... 2-29

2.3.5. Previous IRIS Assessment: Oral Slope Factor ..................................................... 2-31

2.4.INHALATION UNIT RISK FOR CANCER ............................................................................ 2-31

2.5.APPLICATION OF AGE-DEPENDENT ADJUSTMENT FACTORS......................................... 2-31

REFERENCES.................................................................................................................................. R-1

This document is a draft for review purposes only and does not constitute Agency policy.

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TABLES

Table ES-1. Organ/system-specific RfDs and overall RfD for RDX.............................................. ES-2

Table ES-2. Summary of reference dose (RfD) derivation ......................................................... ES-3

Table LS-1. Inclusion-exclusion criteria for health effect studies .............................................. LS-5

Table LS-2. Studies determined not to be informative because of significant issues with design, 

conduct, or reporting .............................................................................................. LS-7

Table LS-3. Considerations and relevant experimental information for evaluation of 

experimental animal studies................................................................................... LS-8

Table LS-4. Summary of experimental animal database.......................................................... LS-10

Table LS-5. Experimental animal studies considered less informative because of certain study 

design, conduct, or reporting limitations.............................................................. LS-14

Table 1-1. Chemical identity and physicochemical properties of RDX ........................................ 1-1

Table 1-2. Evidence pertaining to nervous system effects in humans ........................................ 1-9

Table 1-3. Evidence pertaining to nervous system effects in animals....................................... 1-10

Table 1-4. Evidence pertaining to kidney effects in humans..................................................... 1-24

Table 1-5. Evidence pertaining to kidney and other urogenital system effects in animals....... 1-25

Table 1-6. Six-, 12-, and 24-month incidence of kidney endpoints in male F344 rats reported for 

statistical evaluation in (Levine et al., 1983)........................................................... 1-31

Table 1-7. Six-, 12-, and 24-month incidence of urinary bladder endpoints in male F344 rats 

reported for statistical evaluation in (Levine et al., 1983) ...................................... 1-32

Table 1-8. Six-, 12-, and 24-month incidence of prostate endpoints in male F344 rats reported

for statistical evaluation in (Levine et al., 1983) ..................................................... 1-33

Table 1-9. Evidence pertaining to male reproductive effects in animals .................................. 1-41

Table 1-10. Evidence pertaining to reproductive and developmental effects in animals......... 1-45

Table 1-11. Evidence pertaining to liver effects in humans....................................................... 1-52

Table 1-12. Evidence pertaining to liver effects in animals....................................................... 1-53

Table 1-13. Liver tumors observed in chronic animal bioassays ............................................... 1-63

Table 1-14. Lung tumors observed in chronic animal bioassays................................................ 1-66

Table 2-1. Information considered for evaluation of studies that examined convulsions.......... 2-2

Table 2-2. Summary of derivation of PODs following oral exposure to RDX............................... 2-7

Table 2-3. Effects and corresponding derivation of candidate values....................................... 2-14

Table 2-4. Organ/system-specific RfDs and overall RfD for RDX ............................................... 2-16

Table 2-5. Comparison of dose levels associated with mortality and convulsions in selected 

studies ..................................................................................................................... 2-19

Table 2-6. Summary of dose-response evaluation for mortality following oral exposure to RDX 2-

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Table 2-7. Model predictions and OSFs for hepatocellular and alveolar/bronchiolar adenomas or 

carcinomas in female B6C3F1 mice administered RDX in the diet for 2 years (Lish et 

al., 1984) .................................................................................................................. 2-28

Table 2-8. Summary of uncertainty in the derivation of the cancer risk value for RDX ............ 2-29

This document is a draft for review purposes only and does not constitute Agency policy.

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FIGURES 

Figure LS-1. Summary of literature search and screening process for RDX............................... LS-4

Figure 1-1. Exposure response array of nervous system effects following oral exposure ........ 1-16

Figure 1-2. Exposure-response array of kidney and other urogenital system effects............... 1-35

Figure 1-3. Exposure response array of male reproductive effects following oral exposure.... 1-44

Figure 1-5. Exposure response array of liver effects following oral exposure........................... 1-59

Figure 2-1. Conceptual approach to dose-response modeling for oral exposure. ...................... 2-5

Figure 2-2. Candidate values with corresponding POD and composite UF. .............................. 2-15

This document is a draft for review purposes only and does not constitute Agency policy.

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ABBREVIATIONS

AAP Army ammunition plant

ACGIH American Conference of Governmental 

Industrial Hygienists 

AChE acetylcholinesterase

ADAF age-dependent adjustment factor

AIC Akaike’s information criterion 

ALP alkaline phosphatase

ALT alanine aminotransferase 

AST aspartate aminotransferase 

atm atmosphere

ATSDR Agency for Toxic Substances and 

Disease Registry 

AUC area under the curve

BDNF brain-derived neurotrophic factor

BHC beta-hexachlorocyclohexane

BMC benchmark concentration

BMCL benchmark concentration lower 

confidence limit

BMD benchmark dose 

BMDL benchmark dose lower confidence limit 

BMDS Benchmark Dose Software 

BMDU benchmark dose upper bound 

BMR benchmark response 

BUN blood urea nitrogen 

BW body weight 

CASRN Chemical Abstracts Service Registry 

Number 

CCL Contaminant Candidate List

CI confidence interval

CICAD Concise International Chemical 

Assessment Document

CNS central nervous system 

CSF cerebrospinal fluid

CYP450 cytochrome P450 

DAF dosimetric adjustment factor 

DDT dichlorodiphenyltrichloroethane

d.f. degrees of freedom

DMSO dimethylsulfoxide 

DNA deoxyribonucleic acid 

DNX 1-nitro-3,5-dinitroso1,3,5-triazacyclohexane 

DTIC Defense Technical Information Center

EEG electroencephalogram

EHC Environmental Health Criteria

EPA Environmental Protection Agency 

ER extra risk

FDA Food and Drug Administration 

FOB functional observational battery

FUDS Formerly Used Defense Sites

GABA gamma-amino butyric acid

GD gestational day

GI gastrointestinal

GLP good laboratory practices

HED human equivalent dose 

HERO Health and Environmental Research 

Online

HGPRT hypoxanthine-guanine 

phosphoribosyltransferase

HMX octahydro-1,3,5,7-tetranitro1,3,5,7-tetrazocine

IARC International Agency for Research on 

Cancer

i.p. intraperitoneal

IPCS International Programme on Chemical 

Safety

IRIS Integrated Risk Information System 

IUR inhalation unit risk

i.v. intravenous

LDH lactate dehydrogenase

LOAEL lowest-observed-adverse-effect level 

LOD limit of detection

miRNA microRNA

MNX hexahydro-1-nitroso-3,5-dinitro1,3,5-triazine 

MOA mode of action

MRL Minimal Risk Level

NAPDH nicotinamide adenine dinucleotide 

phosphate 

NAS National Academy of Science

NCE normochromatic erythrocyte

NCEA National Center for Environmental 

Assessment 

NCI National Cancer Institute

NCTR National Center for Toxicological 

Research

NHANES National Health and Nutrition 

Examination Survey

NICNAS National Industrial Chemicals 

Notification and Assessment Scheme

NIEHS National Institute of Environmental 

Health Sciences

NIOSH National Institute for Occupational 

Safety and Health

NOAEL no-observed-adverse-effect level 

NOEL no-observed-effect level 

NPL National Priorities List

NRC Nuclear Regulatory Commission

NSCEP National Service Center for 

Environmental Publications

NTP National Toxicology Program 

NZW New Zealand White

OR odds ratio

This document is a draft for review purposes only and does not constitute Agency policy.

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ORD Office of Research and Development 

OSF oral slope factor

OSHA Occupational Safety and Health 

Administration

PBPK physiologically based pharmacokinetic 

PCB polychlorinated biphenyl

PCE polychromatic erythrocyte

PEL Permissible Exposure Limit

PND postnatal day

POD point of departure 

PWG Pathology Working Group

RBC red blood cell

RDX Royal Demolition eXplosive 

(hexahydro-1,3,5-trinitro1,3,5-triazine) 

REL Recommended Exposure Limit

RfC inhalation reference concentration 

RfD oral reference dose 

SDMS spontaneous death or moribund 

sacrifice

SDWA Safe Drinking Water Act

SGOT glutamic oxaloacetic transaminase, also 

known as AST 

SGPT glutamic pyruvic transaminase, also 

known as ALT 

SLE systemic lupus erythematosus

SS scheduled sacrifice

TLV Threshold Limit Value

TNT trinitrotoluene

TNX hexahydro-1,3,5-trinitroso1,3,5-triazine 

TSCATS Toxic Substances Control Act Test 

Submissions

TWA time-weighted average 

U.S. United States of America

UCM Unregulated Contaminant Monitoring

UF uncertainty factor 

UFA animal-to-human uncertainty factor 

UFD database deficiencies uncertainty factor 

UFH human variation uncertainty factor 

UFL LOAEL-to-NOAEL uncertain factor

UFS subchronic-to-chronic uncertainty 

factor 

WBC white blood cell

WHO World Health Organization

This document is a draft for review purposes only and does not constitute Agency policy.

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AUTHORS | CONTRIBUTORS | REVIEWERS

Assessment Team

Louis D’Amico, Ph.D. (Assessment Manager) U.S. EPA/ORD/NCEA

Todd Blessinger, Ph.D. Washington, DC

Ravi Subramaniam, Ph.D.

Christopher Brinkerhoff, Ph.D. ORISE Postdoctoral Fellow at U.S. EPA/ORD/NCEA 

Currently at U.S. EPA/Office of Chemical Safety and 

Pollution Prevention

Washington, DC

Contributors

Rob DeWoskin, Ph.D. U.S. EPA/ORD/NCEA

Karen Hogan, MS Washington, DC

Jordan Trecki, Ph.D.

Belinda Hawkins, Ph.D. U.S. EPA/ORD/NCEA

Scott Wesselkamper, Ph.D. Cincinnati, OH

Tammy Stoker, Ph.D. U.S. EPA/ORD/National Health and Environmental 

Charles Wood, DVM, PhD, DACVP Research Laboratory (NHEERL)

Research Triangle Park, NC

Anne Loccisano, Ph.D. ORISE Postdoctoral Fellow at U.S. EPA/ORD/NCEA 

Washington, DC

Production Team

Maureen Johnson U.S. EPA/ORD/NCEA

Vicki Soto Washington, DC

Contractor Support

Heather Carlson-Lynch, S.M., DABT SRC, Inc., North Syracuse, NY

Julie Melia, Ph.D., DABT

Megan Riccardi, M.S. 

Pam Ross, M.S. ICF International, Fairfax, VA

Robin Blain, Ph.D.

Executive Direction

Kenneth Olden, Ph.D., Sc.D., L.H.D. (Center Director) U.S. EPA/ORD/NCEA

John Vandenberg, Ph.D, (National Program Director, HHRA) Washington, DC 

Lynn Flowers, Ph.D., DABT (Associate Director for Health)

Vincent Cogliano, Ph.D. (IRIS Program Director) 

Samantha Jones, Ph.D. (IRIS Associate Director for Science)

Susan Rieth, MPH (Quantitative Modeling Branch Chief)

This document is a draft for review purposes only and does not constitute Agency policy.

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Internal Review Team

General Toxicity/Immunotoxicity/Cancer Workgroup U.S. EPA/ORD/NCEA

Epidemiology Workgroup

Neurotoxicity Workgroup

Pharmacokinetics Workgroup

Reproductive and Developmental Toxicity Workgroup

Scoping and Problem Formulation Workgroup

Statistics Workgroup

Toxicity Pathways Workgroup

Reviewers

This assessment was provided for review to scientists in EPA’s program and regional offices. 

Comments were submitted by:

Office of the Administrator/Office of Children’s Health Protection

Office of Chemical Safety and Pollution Prevention/Office of Pesticide Programs 

Office of Land and Emergency Management

Office of Land and Emergency Management/Federal Facilities Forum

Office of Water

Region 2, New York, NY

Region 8, Denver, CO

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This assessment was provided for review to other federal agencies and the Executive Office of the 

President. Comments were submitted by:

Department of Defense 

Department of Energy

Department of Health and Human Services/Agency for Toxic Substances and Disease Registry

Department of Health and Human Services/National Institute of Environmental Health Sciences/National 

Toxicology Program

Department of Health and Human Services/National Institute for Occupational Safety and Health

National Aeronautics and Space Administration

Executive Office of the President/Council on Environmental Quality

Executive Office of the President/Office of Management and Budget

This document is a draft for review purposes only and does not constitute Agency policy.

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PREFACE 

1 This Toxicological Review critically reviews the publicly available studies on hexahydro2 

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1,3,5-trinitro-1,3,5-triazine (RDX, Royal Demolition eXplosive, or cyclonite) in order to identify its 

adverse health effects and to characterize exposure-response relationships. It was prepared under 

the auspices of the U.S. Environmental Protection Agency (EPA) Integrated Risk Information System 

(IRIS) Program. This assessment updates a previous IRIS assessment of RDX that included an oral 

reference dose (RfD) for effects other than cancer (posted in 1988), a determination on the 

carcinogenicity of RDX, and a derivation of an oral slope factor (OSF) to quantify the cancer risk 

associated with RDX exposure (posted in 1990). New information has become available, and this 

assessment reviews information on all health effects by all exposure routes. 

A public meeting was held in December 2013 to obtain input on preliminary materials for 

RDX, including draft literature searches and associated search strategies, evidence tables, and 

exposure-response arrays prior to the development of the IRIS assessment. All public comments 

provided were taken into consideration in developing the draft assessment. The complete set of 

public comments are available on the docket at http://www.regulations.gov (Docket ID No. EPAHQ-ORD-2013-0430).

Organ/system-specific reference values are calculated based on nervous system, kidney/

urogenital system, and male reproductive toxicity data. These reference values may be useful for 

cumulative risk assessments that consider the combined effect of multiple agents acting on the 

same biological system.

This assessment was conducted in accordance with EPA guidance, which is summarized in 

the Preamble to IRIS Toxicological Reviews and cited at appropriate places in this assessment. The 

findings of this assessment and related documents produced during its development are available 

on the IRIS website (http://www.epa.gov/iris). Appendices containing information on assessments 

by other health agencies, details of the literature search strategy, toxicokinetic information, 

summaries of supplementary toxicity information, and dose-response modeling are provided as 

Supplemental Information to this assessment (see Appendices A to D).

For additional information about this assessment or for general questions regarding IRIS, 

please contact EPA’s IRIS Hotline at 202-566-1676 (phone), or hotline.iris@epa.gov. 

Uses and Environmental Occurrence

RDX is a military munitions explosive with limited civilian uses (Gadagbui et al., 2012). In 

the United States, RDX is produced at Army ammunition plants (AAPs) and is not produced 

commercially. RDX production peaked in the 1960s; 180 million pounds per year were produced 

from 1969 to 1971. Yearly total production dropped to 16 million pounds in 1984 (ATSDR, 2012). 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 According to the U.S. EPA Chemical Data Access Tool (http://java.epa.gov/oppt_chemical_search/), 

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the aggregate national production volume in 2012 was approximately 6.3 million pounds per year. 

RDX can be released into environmental media (air, water, soil) as a result of waste 

generated during manufacture, packing, or disposal of the pure product, or use and disposal of RDXcontaining munitions (ATSDR, 2012; Gadagbui et al., 2012; ATSDR, 1999, 1993, 1992). RDX is 

mobile in soil; leaching into groundwater has been reported in samples from military facilities (Best 

et al., 1999a; Godejohann et al., 1998; Bart et al., 1997; Steuckart et al., 1994; Spanggord et al., 

1980a). RDX transport in soil is generally through dissolution by precipitation and subsequent 

downward movement, including migration to groundwater aquifers, and not much via surface 

runoff (U.S. EPA, 2012c). An extensive discussion of RDX properties and fate and transport is 

available in U.S. EPA (2012c). Detectable levels of RDX have been observed in plants irrigated or 

grown with RDX-contaminated water (Best et al., 1999b; Simini and Checkai, 1996; Harvey et al., 

1991). RDX has also been detected in indoor air samples from military facilities where RDX is 

produced (Bishop et al., 1988). 

Exposures to RDX among the general population are likely to be confined to individuals in 

or around active or formerly-used military facilities where RDX is or was produced, stored, or used. 

Oral, inhalation, and dermal routes of exposure may be relevant. 

As of 2015, RDX was detected in surface water, groundwater, sediment, or soil at 34 current 

U.S. EPA National Priorities List (NPL) sites. The NPL serves as a list of sites with known or 

threatened releases of hazardous substances, pollutants, or contaminants throughout the United 

States and its territories. The NPL aids the Agency in identifying the most serious sites that may 

warrant cleanup. The majority of the NPL sites where RDX was listed are associated with military 

facilities. Based on Department of Defense records, Gadagbui et al. (2012) reported that RDX 

contamination is present on 76 active military sites, 9 closed sites, and 15 sites under the Formerly 

Used Defense Sites (FUDS) program. Not all sites under the FUDS program have been sampled, and 

additional sites with RDX contamination in this program could be identified. 

As of 2015, RDX was not regulated under the Safe Drinking Water Act (SDWA), although it 

was included as a contaminant to be monitored under the Unregulated Contaminant Monitoring 

(UCM) Rule by EPA’s Office of Water from 2007 to 2011. Contaminants included in the UCM 

program are suspected of being present in drinking water, but do not have existing health-based 

standards set under the SDWA. RDX has also been included on the Office of Water’s Drinking Water 

Contaminant Candidate List (CCL) since the initial listing was published in 1998. The presence of a 

chemical on the list suggests that it is known or anticipated to occur in public water systems. 

Assessments by Other National and International Health Agencies

Toxicity information on RDX has been evaluated by the Agency for Toxic Substances and 

Disease Registry (ATSDR), National Institute for Occupational Safety and Health (NIOSH), 

Occupational Safety and Health Administration (OSHA), and Australian National Industrial 

Chemicals Notification and Assessment Scheme (NICNAS). The results of these assessments (as of 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 2015) are presented in Appendix A of the Supplemental Information. It is important to recognize 

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that the assessments performed by other health agencies may have been prepared for different 

purposes and may utilize different methods. In addition, newer studies may be included in the IRIS 

assessment.

This document is a draft for review purposes only and does not constitute Agency policy.

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1 

2 PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS

Note: The Preamble to IRIS assessments is being revised based on comments received from 

external peer reviewers and the public, and based on IRIS Program experience with the 

implementation of systematic review methods. Subsequent drafts of the RDX assessment will 

include the revised Preamble.

3 1. Scope of the IRIS Program

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Soon after the EPA was established in 

1970, it was at the forefront of developing 

risk assessment as a science and applying it in 

decisions to protect human health and the 

environment. The Clean Air Act, for example, 

mandates that the EPA provide “an ample 

margin of safety to protect public health”; the 

Safe Drinking Water Act, that “no adverse 

effects on the health of persons may 

reasonably be anticipated to occur, allowing 

an adequate margin of safety.” Accordingly, 

the EPA uses information on the adverse 

effects of chemicals and on exposure levels 

below which these effects are not anticipated 

to occur.

IRIS assessments critically review the 

publicly available studies to identify adverse 

health effects from exposure to chemicals and 

to characterize exposure-response 

relationships. In terms set forth by the 

National Research Council (NRC, 1983), IRIS 

assessments cover the hazard identification 

and dose-response assessment steps of risk 

assessment, not the exposure assessment or 

risk characterization steps that are 

conducted by the EPA’s program and regional 

offices and by other federal, state, and local 

health agencies that evaluate risk in specific 

populations and exposure scenarios. IRIS

assessments are distinct from and do not 

address political, economic, and technical 

considerations that influence the design and 

selection of risk management alternatives. 

An IRIS assessment may cover a single 

chemical, a group of structurally or 

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complex mixture. These agents may be found 

in air, water, soil, or sediment. Exceptions are 

chemicals currently used exclusively as 

pesticides, ionizing and non-ionizing 

radiation, and criteria air pollutants listed 

under Section 108 of the Clean Air Act 

(carbon monoxide, lead, nitrogen oxides, 

ozone, particulate matter, and sulfur oxides).

Periodically, the IRIS Program asks other 

EPA programs and regions, other federal 

agencies, state health agencies, and the 

general public to nominate chemicals and 

mixtures for future assessment or 

reassessment. Agents may be considered for 

reassessment as significant new studies are 

published. Selection is based on program and 

regional office priorities and on availability of 

adequate information to evaluate the 

potential for adverse effects. Other agents 

may also be assessed in response to an urgent 

public health need. 

2. Process for developing and peerreviewing IRIS assessments

The process for developing IRIS 

assessments (revised in May 2009 and

enhanced in July 2013) involves critical 

analysis of the pertinent studies, 

opportunities for public input, and multiple 

levels of scientific review. The EPA revises 

draft assessments after each review, and 

external drafts and comments become part of 

the public record (U.S. EPA, 2009). 

Before beginning an assessment, the IRIS 

Program discusses the scope with other EPA 

programs and regions to ensure that the 

This document is a draft for review purposes only and does not constitute Agency policy.

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ssessment will meet their needs. Then a 

ublic meeting on problem formulation 

vites discussion of the key issues and the 

udies and analytical approaches that might 

ontribute to their resolution.

ep 4. Public review and comment, 

followed by external peer review. The 

EPA releases the draft assessment for 

public review and comment. A public 

meeting provides an opportunity to 

discuss the assessment prior to peer 

review. Then the EPA releases a draft for 

external peer review. The peer review 

meeting is open to the public and includes 

time for oral public comments. The peer 

reviewers assess whether the evidence 

has been assembled and evaluated 

according to guidelines and whether the 

conclusions are justified by the evidence. 

The peer review draft, written public 

comments, and peer review report 

ep 1. Development of a draft 

Toxicological Review. The draft

assessment considers all pertinent 

publicly available studies and applies 

consistent criteria to evaluate study 

quality, identify health effects, identify 

mechanistic events and pathways, 

integrate the evidence of causation for 

each effect, and derive toxicity values. A

public meeting prior to the integration of 

evidence and derivation of toxicity values 

promotes public discussion of the 

literature search, evidence, and key 

issues.

ep 2. Internal review by scientists in 

EPA programs and regions. The draft 

assessment is revised to address the 

comments from within the EPA.

ep 3. Interagency science consultation 

with other federal agencies and the 

Executive Offices of the President. The 

draft assessment is revised to address the 

interagency comments. The science 

consultation draft, interagency 

comments, and the EPA’s response to 

major comments become part of the 

public record.

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become part of the public record.

Step 5. Revision of draft Toxicological 

Review and development of draft IRIS 

summary. The draft assessment is 

revised to reflect the peer review 

comments, public comments, and newly 

published studies that are critical to the 

conclusions of the assessment. The 

disposition of peer review comments and 

public comments becomes part of the 

public record.

Step 6. Final EPA review and interagency 

science discussion with other federal 

agencies and the Executive Offices of 

the President The draft assessment and 

summary are revised to address the EPA 

and interagency comments. The science 

discussion draft, written interagency 

comments, and EPA’s response to major 

comments become part of the public 

record. 

Step 7. Completion and posting. The 

Toxicological Review and IRIS summary 

are posted on the IRIS website 

(http://www.epa.gov/iris/).

The remainder of this Preamble 

addresses step 1, the development of a draft 

Toxicological Review. IRIS assessments 

follow standard practices of evidence 

evaluation and peer review, many of which 

are discussed in EPA guidelines (U.S. EPA, 

2005a, b, 2000b, 1998, 1996, 1991, 1986a, b) 

and other methods (U.S. EPA, 2012a, b, 2011, 

2006a, b, 2002, 1994). Transparent 

application of scientific judgment is of 

paramount importance. To provide a 

harmonized approach across IRIS 

assessments, this Preamble summarizes 

concepts from these guidelines and 

emphasizes principles of general 

applicability.

3. Identifying and selecting 

pertinent studies

3.1. Identifying studies

This document is a draft for review purposes only and does not constitute Agency policy.

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conducts a comprehensive search of the 

primary scientific literature. The literature 

search follows standard practices and 

includes the PubMed and ToxNet databases 

of the National Library of Medicine, Web of 

Science, and other databases listed in the 

EPA’s HERO system (Health and 

Environmental Research Online, 

http://hero.epa.gov/). Searches for 

information on mechanisms of toxicity are 

inherently specialized and may include 

studies on other agents that act through 

related mechanisms.

Each assessment specifies the search 

strategies, keywords, and cut-off dates of its 

literature searches. The EPA posts the results 

of the literature search on the IRIS website 

and requests information from the public on 

additional studies and ongoing research.

The EPA also considers studies received 

through the IRIS Submission Desk and 

studies (typically unpublished) submitted 

under the Toxic Substances Control Act or the 

Federal Insecticide, Fungicide, and 

Rodenticide Act. Material submitted as 

Confidential Business Information is 

considered only if it includes health and 

safety data that can be publicly released. If a 

study that may be critical to the conclusions 

of the assessment has not been peerreviewed, the EPA will have it peer-reviewed.

The EPA also examines the toxicokinetics 

of the agent to identify other chemicals (for 

example, major metabolites of the agent) to 

include in the assessment if adequate 

information is available, in order to more 

fully explain the toxicity of the agent and to 

suggest dose metrics for subsequent 

modeling. 

In assessments of chemical mixtures, 

mixture studies are preferred for their ability 

to reflect interactions among components. 

The literature search seeks, in 

decreasing order of preference (U.S. EPA, 

2000b, §2.2; 1986b, §2.1)]: 

– Studies of the mixture being assessed.

– Studies of a sufficiently similar 

mixture. In evaluating similarity, the 

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of mixtures in the environment 

through partitioning and 

transformation.

– Studies of individual chemical 

components of the mixture, if there 

are not adequate studies of 

sufficiently similar mixtures.

3.2. Selecting pertinent epidemiologic 

studies

Study design is the key consideration for 

selecting pertinent epidemiologic studies 

from the results of the literature search. 

– Cohort studies, case-control studies, 

and some population-based surveys 

(for example, NHANES) provide the 

strongest epidemiologic evidence, 

especially if they collect information 

about individual exposures and 

effects.

– Ecological studies (geographic 

correlation studies) relate exposures 

and effects by geographic area. They 

can provide strong evidence if there 

are large exposure contrasts between 

geographic areas, relatively little 

exposure variation within study 

areas, and population migration is 

limited.

– Case reports of high or accidental 

exposure lack definition of the 

population at risk and the expected 

number of cases. They can provide 

information about a rare effect or 

about the relevance of analogous 

results in animals.

The assessment briefly reviews 

ecological studies and case reports but 

reports details only if they suggest effects not 

identified by other studies.

3.3. Selecting pertinent experimental 

studies

Exposure route is a key design 

consideration for selecting pertinent 

This document is a draft for review purposes only and does not constitute Agency policy.

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clinical studies.

– Studies of oral, inhalation, or dermal 

exposure involve passage through an 

absorption barrier and are 

considered most pertinent to human 

environmental exposure.

– Injection or implantation studies are 

often considered less pertinent but

may provide valuable toxicokinetic or 

mechanistic information. They also 

may be useful for identifying effects in 

animals if deposition or absorption is 

problematic (for example, for 

particles and fibers).

Exposure duration is also a key design 

consideration for selecting pertinent 

experimental animal studies.

– Studies of effects from chronic 

exposure are most pertinent to 

lifetime human exposure.

– Studies of effects from less-thanchronic exposure are pertinent but 

less preferred for identifying effects 

from lifetime human exposure. Such 

studies may be indicative of effects 

from less-than-lifetime human 

exposure.

Short-duration studies involving animals 

or humans may provide toxicokinetic or 

mechanistic information.

For developmental toxicity and 

reproductive toxicity, irreversible effects 

may result from a brief exposure during a 

critical period of development. Accordingly, 

specialized study designs are used for these 

effects (U.S. EPA, 2006b, 1998, 1996, 1991). 

4. Evaluating the quality of 

individual studies

After the subsets of pertinent 

epidemiologic and experimental studies have 

been selected from the literature searches, 

the assessment evaluates the quality of each 

individual study. This evaluation considers 

the design, methods, conduct, and 

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ocumentation of each study, but not 

whether the results are positive, negative, or 

ull. The objective is to identify the stronger, 

more informative studies based on a uniform 

valuation of quality characteristics across 

tudies of similar design.

.1. Evaluating the quality of 

epidemiologic studies

The assessment evaluates design and 

methodological aspects that can increase or 

ecrease the weight given to each 

pidemiologic study in the overall evaluation 

U.S. EPA, 2005a, 1998, 1996, 1994, 1991): 

– Documentation of study design, 

methods, population characteristics, 

and results.

– Definition and selection of the study 

group and comparison group.

– Ascertainment of exposure to the 

chemical or mixture.

– Ascertainment of disease or health 

effect.

– Duration of exposure and follow-up 

and adequacy for assessing the 

occurrence of effects.

– Characterization of exposure during 

critical periods.

– Sample size and statistical power to 

detect anticipated effects.

– Participation rates and potential for 

selection bias as a result of the 

achieved participation rates.

– Measurement error (can lead to 

misclassification of exposure, health 

outcomes, and other factors) and 

other types of information bias.

– Potential confounding and other 

sources of bias addressed in the study 

design or in the analysis of results. 

The basis for consideration of 

confounding is a reasonable 

expectation that the confounder is 

related to both exposure and 

This document is a draft for review purposes only and does not constitute Agency policy.

xvii DRAFT—DO NOT CITE OR QUOTE
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to result in bias.

For developmental toxicity, reproductive 

toxicity, neurotoxicity, and cancer there is 

further guidance on the nuances of evaluating 

epidemiologic studies of these effects (U.S. 

EPA, 2005a, 1998, 1996, 1991). 

4.2. Evaluating the quality of 

experimental studies

The assessment evaluates design and 

methodological aspects that can increase or 

decrease the weight given to each 

experimental animal study, in-vitro study, or 

human clinical study (U.S. EPA, 2005a, 1998, 

1996, 1991). Research involving human 

subjects is considered only if conducted 

according to ethical principles.

– Documentation of study design, 

animals or study population, 

methods, basic data, and results.

– Nature of the assay and validity for its 

intended purpose.

– Characterization of the nature and 

extent of impurities and 

contaminants of the administered 

chemical or mixture.

– Characterization of dose and dosing 

regimen (including age at exposure) 

and their adequacy to elicit adverse 

effects, including latent effects.

– Sample sizes and statistical power to 

detect dose-related differences or 

trends. 

– Ascertainment of survival, vital signs, 

disease or effects, and cause of death.

– Control of other variables that could 

influence the occurrence of effects.

The assessment uses statistical tests to 

evaluate whether the observations may be 

due to chance. The standard for determining 

statistical significance of a response is a trend 

test or comparison of outcomes in the 

exposed groups against those of concurrent 

controls. In some situations, examination of 

43 historical control data from the same 

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laboratory within a few years of the study 

may improve the analysis. For an uncommon 

effect that is not statistically significant 

compared with concurrent controls, 

historical controls may show that the effect is 

unlikely to be due to chance. For a response 

that appears significant against a concurrent 

control response that is unusual, historical 

controls may offer a different interpretation 

(U.S. EPA, 2005a, §2.2.2.1.3). 

For developmental toxicity, reproductive 

toxicity, neurotoxicity, and cancer there is 

further guidance on the nuances of evaluating 

experimental studies of these effects (U.S. 

EPA, 2005a, 1998, 1996, 1991). In multigeneration studies, agents that produce 

developmental effects at doses that are not 

toxic to the maternal animal are of special 

concern. Effects that occur at doses 

associated with mild maternal toxicity are not 

assumed to result only from maternal 

toxicity. Moreover, maternal effects may be 

reversible, while effects on the offspring may 

be permanent (U.S. EPA, 1998, §3.1.2.4.5.4; 

1991, §3.1.1.4),.

4.3. Reporting study results

The assessment uses evidence tables to 

present the design and key results of 

pertinent studies. There may be separate 

tables for each site of toxicity or type of study.

If a large number of studies observe the 

same effect, the assessment considers the 

study quality characteristics in this section to 

identify the strongest studies or types of 

study. The tables present details from these 

studies, and the assessment explains the 

reasons for not reporting details of other 

studies or groups of studies that do not add 

new information. Supplemental information 

provides references to all studies considered, 

including those not summarized in the tables.

The assessment discusses strengths and 

limitations that affect the interpretation of 

each study. If the interpretation of a study in 

the assessment differs from that of the study 

authors, the assessment discusses the basis 

for the difference.

This document is a draft for review purposes only and does not constitute Agency policy.

xviii DRAFT—DO NOT CITE OR QUOTE
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As a check on the selection and evaluation 

 pertinent studies, the EPA asks peer 

eviewers to identify studies that were not 

dequately considered.

. Evaluating the overall evidence 

of each effect

1. Concepts of causal inference

For each health effect, the assessment 

valuates the evidence as a whole to 

etermine whether it is reasonable to infer a 

ausal association between exposure to the 

gent and the occurrence of the effect. This 

ference is based on information from 

ertinent human studies, animal studies, and 

echanistic studies of adequate quality. 

ositive, negative, and null results are given 

eight according to study quality.

Causal inference involves scientific 

dgment, and the considerations are 

uanced and complex. Several health 

gencies have developed frameworks for 

ausal inference, among them the U.S. 

urgeon General (CDC, 2004; HEW, 1964), 

e International Agency for Research on 

ancer (IARC, 2006), the Institute of Medicine 

OM, 2008), and the (U.S. EPA (2010), §1.6; 

005a), §2.5). Although developed for 

fferent purposes, the frameworks are 

milar in nature and provide an established 

ructure and language for causal inference. 

ach considers aspects of an association that 

uggest causation, discussed by Hill (1965)

nd elaborated by Rothman and Greenland 

1998), and (U.S. EPA (2005a), §2.2.1.7; 

994), Appendix C). 

rength of association: The finding of a 

large relative risk with narrow 

confidence intervals strongly suggests 

that an association is not due to chance, 

bias, or other factors. Modest relative 

risks, however, may reflect a small range 

of exposures, an agent of low potency, an 

increase in an effect that is common, 

exposure misclassification, or other 

sources of bias.

43 Consistency of association: An inference of 

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causation is strengthened if elevated 

risks are observed in independent studies 

of different populations and exposure 

scenarios. Reproducibility of findings 

constitutes one of the strongest 

arguments for causation. Discordant 

results sometimes reflect differences in 

study design, exposure, or confounding

factors.

Specificity of association: As originally 

intended, this refers to one cause 

associated with one effect. Current 

understanding that many agents cause 

multiple effects and many effects have 

multiple causes make this a less 

informative aspect of causation, unless 

the effect is rare or unlikely to have 

multiple causes.

Temporal relationship: A causal 

interpretation requires that exposure 

precede development of the effect.

Biologic gradient (exposure-response 

relationship): Exposure-response 

relationships strongly suggest causation. 

A monotonic increase is not the only 

pattern consistent with causation. The

presence of an exposure-response 

gradient also weighs against bias and 

confounding as the source of an 

association.

Biologic plausibility: An inference of 

causation is strengthened by data 

demonstrating plausible biologic 

mechanisms, if available. Plausibility 

may reflect subjective prior beliefs if 

there is insufficient understanding of the 

biologic process involved.

Coherence: An inference of causation is 

strengthened by supportive results from 

animal experiments, toxicokinetic 

studies, and short-term tests. Coherence 

may also be found in other lines of 

evidence, such as changing disease 

patterns in the population.

“Natural experiments”: A change in 

exposure that brings about a change in 

This document is a draft for review purposes only and does not constitute Agency policy.

xix DRAFT—DO NOT CITE OR QUOTE
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evidence, as it tests the hypothesis of 

causation. An example would be an 

intervention to reduce exposure in the 

workplace or environment that is 

followed by a reduction of an adverse 

effect.

Analogy: Information on structural 

analogues or on chemicals that induce 

similar mechanistic events can provide 

insight into causation.

These considerations are consistent with 

guidelines for systematic reviews that 

evaluate the quality and weight of evidence. 

Confidence is increased if the magnitude of 

effect is large, if there is evidence of an 

exposure-response relationship, or if an 

association was observed and the plausible 

biases would tend to decrease the magnitude 

of the reported effect. Confidence is 

decreased for study limitations, 

inconsistency of results, indirectness of 

evidence, imprecision, or reporting bias 

(Guyatt et al., 2008b; Guyatt et al., 2008a). 

5.2. Evaluating evidence in humans

For each effect, the assessment evaluates 

the evidence from the epidemiologic studies 

as a whole. The objective is to determine 

whether a credible association has been 

observed and, if so, whether that association 

is consistent with causation. In doing this, the 

assessment explores alternative explanations 

(such as chance, bias, and confounding) and 

draws a conclusion about whether these 

alternatives can satisfactorily explain any 

observed association. 

To make clear how much the 

epidemiologic evidence contributes to the 

overall weight of the evidence, the 

assessment may select a standard descriptor 

to characterize the epidemiologic evidence of 

association between exposure to the agent 

and occurrence of a health effect.

Sufficient epidemiologic evidence of an 

association consistent with causation:

The evidence establishes a causal 

association for which alternative 

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confounding can be ruled out with 

reasonable confidence.

uggestive epidemiologic evidence of an 

association consistent with causation:

The evidence suggests a causal 

association but chance, bias, or 

confounding cannot be ruled out as 

explaining the association.

nadequate epidemiologic evidence to infer 

a causal association: The available 

studies do not permit a conclusion 

regarding the presence or absence of an 

association.

pidemiologic evidence consistent with no 

causal association: Several adequate 

studies covering the full range of human 

exposures and considering susceptible 

populations, and for which alternative 

explanations such as bias and 

confounding can be ruled out, are 

mutually consistent in not finding an 

association.

.3. Evaluating evidence in animals

For each effect, the assessment evaluates 

he evidence from the animal experiments as 

whole to determine the extent to which they 

ndicate a potential for effects in humans. 

onsistent results across various species and 

rains increase confidence that similar 

esults would occur in humans. Several 

oncepts discussed by Hill (1965) are 

ertinent to the weight of experimental 

esults: consistency of response, doseesponse relationships, strength of response, 

iologic plausibility, and coherence (U.S. EPA, 

005a, §2.2.1.7; 1994, Appendix C). 

In weighing evidence from multiple 

xperiments, U.S. EPA (2005a), §2.5

istinguishes: 

onflicting evidence (that is, mixed positive 

and negative results in the same sex and 

strain using a similar study protocol) 

from 

This document is a draft for review purposes only and does not constitute Agency policy.

xx DRAFT—DO NOT CITE OR QUOTE
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negative results are in different sexes or 

strains or use different study protocols). 

Negative or null results do not invalidate 

positive results in a different experimental 

system. The EPA regards all as valid 

observations and looks to explain differing 

results using mechanistic information (for 

example, physiologic or metabolic 

differences across test systems) or 

methodological differences (for example, 

relative sensitivity of the tests, differences in 

dose levels, insufficient sample size, or timing 

of dosing or data collection).

It is well established that there are critical 

periods for some developmental and 

reproductive effects (U.S. EPA, 2006b, 2005a, 

b, 1998, 1996, 1991). Accordingly, the 

assessment determines whether critical 

periods have been adequately investigated. 

Similarly, the assessment determines 

whether the database is adequate to evaluate 

other critical sites and effects.

In evaluating evidence of genetic toxicity:

– Demonstration of gene mutations, 

chromosome aberrations, or 

aneuploidy in humans or 

experimental mammals (in vivo) 

provides the strongest evidence.

– This is followed by positive results in 

lower organisms or in cultured cells 

(in vitro) or for other genetic events.

– Negative results carry less weight, 

partly because they cannot exclude 

the possibility of effects in other 

tissues (IARC, 2006). 

For germ-cell mutagenicity, the EPA has 

defined categories of evidence, ranging from 

positive results of human germ-cell 

mutagenicity to negative results for all effects 

of concern (U.S. EPA, 1986a, §2.3). 

5.4. Evaluating mechanistic data 

Mechanistic data can be useful in 

answering several questions.

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interpretation of human studies.

– The generalizability of animal studies 

to humans.

– The susceptibility of particular 

populations or lifestages.

The focus of the analysis is to describe, if 

ossible, mechanistic pathways that lead to a 

ealth effect. These pathways encompass:

– Toxicokinetic processes of absorption, 

distribution, metabolism, and 

elimination that lead to the formation 

of an active agent and its presence at 

the site of initial biologic interaction.

– Toxicodynamic processesthat lead to a 

health effect at this or another site 

(also known as a mode of action).

For each effect, the assessment discusses 

he available information on its modes of 

ction and associated key events (key events

eing empirically observable, necessary 

recursor steps or biologic markers of such 

eps; mode of action being a series of key 

vents involving interaction with cells, 

perational and anatomic changes, and 

esulting in disease). Pertinent information 

may also come from studies of metabolites or 

f compounds that are structurally similar or 

hat act through similar mechanisms. 

nformation on mode of action is not required 

or a conclusion that the agent is causally 

elated to an effect (U.S. EPA, 2005a, §2.5). 

The assessment addresses several 

uestions about each hypothesized mode of 

ction (U.S. EPA, 2005a, §2.4.3.4). 

) Is the hypothesized mode of action 

sufficiently supported in test animals?

Strong support for a key event being 

necessary to a mode of action can come 

from experimental challenge to the 

hypothesized mode of action, in which 

studies that suppress a key event observe 

suppression of the effect. Support for a 

mode of action is meaningfully 

strengthened by consistent results in 

different experimental models, much 

This document is a draft for review purposes only and does not constitute Agency policy.

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the same model. The assessment may 

consider various aspects of causation in 

addressing this question.

2) Is the hypothesized mode of action 

relevant to humans? The assessment 

reviews the key events to identify critical 

similarities and differences between the 

test animals and humans. Site 

concordance is not assumed between 

animals and humans, though it may hold 

for certain effects or modes of action. 

Information suggesting quantitative 

differences in doses where effects would 

occur in animals or humans is considered 

in the dose-response analysis. Current 

levels of human exposure are not used to 

rule out human relevance, as IRIS 

assessments may be used in evaluating 

new or unforeseen circumstances that 

may entail higher exposures.

3) Which populations or lifestages can be 

particularly susceptible to the 

hypothesized mode of action? The 

assessment reviews the key events to 

identify populations and lifestages that 

might be susceptible to their occurrence. 

Quantitative differences may result in 

separate toxicity values for susceptible 

populations or lifestages.

The assessment discusses the likelihood 

that an agent operates through multiple 

modes of action. An uneven level of support 

for different modes of action can reflect 

disproportionate resources spent 

investigating them (U.S. EPA, 

2005a, §2.4.3.3). It should be noted that in 

clinical reviews, the credibility of a series of 

studies is reduced if evidence is limited to 

studies funded by one interested sector 

(Guyatt et al., 2008a). 

For cancer, the assessment evaluates 

evidence of a mutagenic mode of action to 

guide extrapolation to lower doses and 

consideration of susceptible lifestages. Key 

data include the ability of the agent or a 

metabolite to react with or bind to DNA, 

positive results in multiple test systems, or 

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milar properties and structure-activity 

elationships to mutagenic carcinogens (U.S. 

PA, 2005a ,§2.3.5). 

.5. Characterizing the overall weight 

of the evidence

After evaluating the human, animal, and 

mechanistic evidence pertinent to an effect, 

he assessment answers the question: Does 

he agent cause the adverse effect? (NRC, 

009, 1983). In doing this, the assessment 

evelops a narrative that integrates the 

vidence pertinent to causation. To provide 

larity and consistency, the narrative 

ncludes a standard hazard descriptor. For 

xample, the following standard descriptors 

ombine epidemiologic, experimental, and 

mechanistic evidence of carcinogenicity (U.S. 

PA, 2005a, §2.5). 

arcinogenic to humans: There is 

convincing epidemiologic evidence of a 

causal association (that is, there is 

reasonable confidence that the 

association cannot be fully explained by 

chance, bias, or confounding); or there is 

strong human evidence of cancer or its 

precursors, extensive animal evidence, 

identification of key precursor events in 

animals, and strong evidence that they 

are anticipated to occur in humans.

ikely to be carcinogenic to humans: The 

evidence demonstrates a potential 

hazard to humans but does not meet the 

criteria for carcinogenic. There may be a 

plausible association in humans, multiple 

positive results in animals, or a 

combination of human, animal, or other 

experimental evidence.

uggestive evidence of carcinogenic 

potential: The evidence raises concern 

for effects in humans but is not sufficient 

for a stronger conclusion. This descriptor 

covers a range of evidence, from a 

positive result in the only available study 

to a single positive result in an extensive 

database that includes negative results in 

other species.

This document is a draft for review purposes only and does not constitute Agency policy.

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nadequate information to assess 

carcinogenic potential: No other 

descriptors apply. Conflicting evidence

can be classified as inadequate 

information if all positive results are 

opposed by negative studies of equal 

quality in the same sex and strain. 

Differing results, however, can be 

classified as suggestive evidence or as 

likely to be carcinogenic. 

Not likely to be carcinogenic to humans:

There is robust evidence for concluding 

that there is no basis for concern. There 

may be no effects in both sexes of at least 

two appropriate animal species; positive 

animal results and strong, consistent 

evidence that each mode of action in 

animals does not operate in humans; or 

convincing evidence that effects are not 

likely by a particular exposure route or 

below a defined dose.

Multiple descriptors may be used if there 

s evidence that carcinogenic effects differ by 

dose range or exposure route (U.S. EPA, 

2005a, §2.5). 

Another example of standard descriptors 

comes from the EPA’s Integrated Science 

Assessments, which evaluate causation for 

the effects of the criteria pollutants in 

ambient air (U.S. EPA, 2010). 

Causal relationship: Sufficient evidence to 

conclude that there is a causal 

relationship. Observational studies 

cannot be explained by plausible 

alternatives, or they are supported by 

other lines of evidence, for example, 

animal studies or mechanistic 

information.

Likely to be a causal relationship: Sufficient 

evidence that a causal relationship is 

likely, but important uncertainties 

remain. For example, observational 

studies show an association but coexposures are difficult to address or other 

lines of evidence are limited or 

inconsistent; or multiple animal studies 

from different laboratories demonstrate 

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data.

Suggestive of a causal relationship: At least 

one high-quality epidemiologic study 

shows an association but other studies 

are inconsistent.

Inadequate to infer a causal relationship:

The studies do not permit a conclusion 

regarding the presence or absence of an 

association.

Not likely to be a causal relationship:

Several adequate studies, covering the 

full range of human exposure and 

considering susceptible populations, are 

mutually consistent in not showing an 

effect at any level of exposure.

The EPA is investigating and may on a 

trial basis use these or other standard 

descriptors to characterize the overall weight 

of the evidence for effects other than cancer.

6. Selecting studies for derivation 

of toxicity values

For each effect where there is credible 

evidence of an association with the agent, the 

assessment derives toxicity values if there 

are suitable epidemiologic or experimental 

data. The decision to derive toxicity values 

may be linked to the hazard descriptor.

Dose-response analysis requires 

quantitative measures of dose and response. 

Then, other factors being equal:

– Epidemiologic studies are preferred 

over animal studies, if quantitative 

measures of exposure are available 

and effects can be attributed to the 

agent.

– Among experimental animal models, 

those that respond most like humans 

are preferred, if the comparability of 

response can be determined.

– Studies by a route of human 

environmental exposure are 

preferred, although a validated 

This document is a draft for review purposes only and does not constitute Agency policy.

xxiii DRAFT—DO NOT CITE OR QUOTE
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extrapolate across exposure routes.

– Studies of longer exposure duration

and follow-up are preferred, to 

minimize uncertainty about whether 

effects are representative of lifetime 

exposure.

– Studies with multiple exposure levels 

are preferred for their ability to 

provide information about the shape 

of the exposure-response curve. 

– Studies with adequate power to 

detect effects at lower exposure 

levels are preferred, to minimize the 

extent of extrapolation to levels found 

in the environment.

Studies with non-monotonic exposureresponse relationships are not necessarily 

excluded from the analysis. A diminished 

effect at higher exposure levels may be 

satisfactorily explained by factors such as 

competing toxicity, saturation of absorption 

or metabolism, exposure misclassification, or 

selection bias.

If a large number of studies are suitable 

for dose-response analysis, the assessment 

considers the study characteristics in this 

section to focus on the most informative data. 

The assessment explains the reasons for not 

analyzing other groups of studies. As a check 

on the selection of studies for dose-response 

analysis, the EPA asks peer reviewers to 

identify studies that were not adequately 

considered. 

7. Deriving toxicity values

7.1. General framework for doseresponse analysis

The EPA uses a two-step approach that 

distinguishes analysis of the observed doseresponse data from inferences about lower 

doses (U.S. EPA, 2005a, §3). 

Within the observed range, the preferred 

approach is to use modeling to incorporate a 

wide range of data into the analysis. The 

modeling yields a point of departure (an 

43 exposure level near the lower end of the 

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observed range, without significant 

extrapolation to lower doses) (Sections 7.2-

7.3).

Extrapolation to lower doses considers 

what is known about the modes of action for 

each effect (Sections 7.4-7.5). If response 

estimates at lower doses are not required, an 

alternative is to derive reference values, 

which are calculated by applying factors to 

the point of departure in order to account for 

sources of uncertainty and variability 

(Section 7.6).

For a group of agents that induce an effect 

through a common mode of action, the doseresponse analysis may derive a relative 

potency factor for each agent. A full doseresponse analysis is conducted for one wellstudied index chemical in the group, then the 

potencies of other members are expressed in 

relative terms based on relative toxic effects, 

relative absorption or metabolic rates, 

quantitative structure-activity relationships, 

or receptor binding characteristics (U.S. EPA, 

2005a, §3.2.6; 2000b, §4.4). 

Increasingly, the EPA is basing toxicity 

values on combined analyses of multiple data 

sets or multiple responses. The EPA also 

considers multiple dose-response 

approaches if they can be supported by 

robust data.

7.2. Modeling dose to sites of biologic 

effects

The preferred approach for analysis of 

dose is toxicokinetic modeling because of its 

ability to incorporate a wide range of data. 

The preferred dose metric would refer to the

active agent at the site of its biologic effect or 

to a close, reliable surrogate measure. The

active agent may be the administered 

chemical or a metabolite. Confidence in the 

use of a toxicokinetic model depends on the 

robustness of its validation process and on 

the results of sensitivity analyses (U.S. EPA, 

2006a; 2005a, §3.1; 1994, §4.3). 

Because toxicokinetic modeling can 

require many parameters and more data than 

are typically available, the EPA has developed 

This document is a draft for review purposes only and does not constitute Agency policy.

xxiv DRAFT—DO NOT CITE OR QUOTE
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typical data sets. These standard approaches 

also facilitate comparison across exposure 

patterns and species.

– Intermittent study exposures are 

standardized to a daily average over 

the duration of exposure. For chronic 

effects, daily exposures are averaged 

over the lifespan. Exposures during a 

critical period, however, are not 

averaged over a longer duration (U.S. 

EPA, 2005a, §3.1.1; 1991, §3.2). 

– Doses are standardized to equivalent 

human terms to facilitate comparison 

of results from different species.

– Oral doses are scaled allometrically 

using mg/kg3/4-day as the equivalent 

dose metric across species. 

Allometric scaling pertains to 

equivalence across species, not 

across lifestages, and is not used to 

scale doses from adult humans or 

mature animals to infants or children 

(U.S. EPA, 2011; 2005a, §3.1.3). 

– Inhalation exposures are scaled using 

dosimetry models that apply speciesspecific physiologic and anatomic 

factors and consider whether the 

effect occurs at the site of first contact 

or after systemic circulation (U.S. 

EPA, 2012a; 1994, §3). 

It can be informative to convert doses 

across exposure routes. If this is done, the 

assessment describes the underlying data, 

algorithms, and assumptions (U.S. EPA, 

2005a, §3.1.4). 

In the absence of study-specific data on, 

for example, intake rates or body weight, the 

EPA has developed recommended values for 

use in dose-response analysis (U.S. EPA, 

1988). 

7.3. Modeling response in the range 

of observation

Toxicodynamic (“biologically based”) 

modeling can incorporate data on biologic 

processes leading to an effect. Such models 

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action and to quantitatively support model 

parameters associated with its key events. 

Because different models may provide 

equivalent fits to the observed data but 

diverge substantially at lower doses, critical 

biologic parameters should be measured 

from laboratory studies, not by model fitting. 

Confidence in the use of a toxicodynamic 

model depends on the robustness of its 

validation process and on the results of 

sensitivity analyses. Peer review of the 

scientific basis and performance of a model is 

essential (U.S. EPA, 2005a, §3.2.2). 

Because toxicodynamic modeling can 

require many parameters and more 

knowledge and data than are typically 

available, the EPA has developed a standard 

set of empirical (“curve-fitting”) models 

(http://www.epa.gov/ncea/bmds/) that can 

be applied to typical data sets, including those 

that are nonlinear. The EPA has also 

developed guidance on modeling doseresponse data, assessing model fit, selecting 

suitable models, and reporting modeling 

results (U.S. EPA, 2012b). Additional 

judgment or alternative analyses are used if 

the procedure fails to yield reliable results, 

for example, if the fit is poor, modeling may 

be restricted to the lower doses, especially if 

there is competing toxicity at higher doses 

(U.S. EPA, 2005a, §3.2.3). 

Modeling is used to derive a point of 

departure (U.S. EPA, 2012b; 2005a, §3.2.4). 

(See Section 7.6 for alternatives if a point of 

departure cannot be derived by modeling.):

– If linear extrapolation is used, 

selection of a response level 

corresponding to the point of 

departure is not highly influential, so 

standard values near the low end of 

the observable range are generally 

used (for example, 10% extra risk for 

cancer bioassay data, 1% for 

epidemiologic data, lower for rare 

cancers).

– For nonlinear approaches, both 

statistical and biologic considerations 

are taken into account.

This document is a draft for review purposes only and does not constitute Agency policy.

xxv DRAFT—DO NOT CITE OR QUOTE
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– For dichotomous data, a response 

level of 10% extra risk is generally 

used for minimally adverse effects, 

5% or lower for more severe effects.

– For continuous data, a response level 

is ideally based on an established 

definition of biologic significance. In 

the absence of such definition, one 

control standard deviation from the 

control mean is often used for 

minimally adverse effects, one-half 

standard deviation for more severe 

effects.

The point of departure is the 95% lower 

ound on the dose associated with the 

elected response level.

.4. Extrapolating to lower doses and 

response levels

The purpose of extrapolating to lower 

oses is to estimate responses at exposures 

elow the observed data. Low-dose 

xtrapolation, typically used for cancer data, 

onsiders what is known about modes of 

ction (U.S. EPA, 2005a, §3.3.1 and §3.3.2). 

) If a biologically based model has been 

developed and validated for the agent, 

extrapolation may use the fitted model 

below the observed range if significant 

model uncertainty can be ruled out with 

reasonable confidence.

2) Linear extrapolation is used if the doseresponse curve is expected to have a 

linear component below the point of 

departure. This includes:

– Agents or their metabolites that are 

DNA-reactive and have direct 

mutagenic activity.

– Agents or their metabolites for which 

human exposures or body burdens 

are near doses associated with key 

events leading to an effect.

Linear extrapolation is also used when 

data are insufficient to establish mode of 

action and when scientifically plausible. 

43 The result of linear extrapolation is 

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described by an oral slope factor or an 

inhalation unit risk, which is the slope of 

the dose-response curve at lower doses 

or concentrations, respectively.

3) Nonlinear models are used for 

extrapolation if there are sufficient data 

to ascertain the mode of action and to 

conclude that it is not linear at lower 

doses, and the agent does not 

demonstrate mutagenic or other activity 

consistent with linearity at lower doses. 

Nonlinear approaches generally should 

not be used in cases where mode of action 

has not ascertained. If nonlinear 

extrapolation is appropriate but no 

model is developed, an alternative is to 

calculate reference values.

4) Both linear and nonlinear approaches 

may be used if there a multiple modes of 

action. For example, modeling to a low 

response level can be useful for 

estimating the response at doses where a 

high-dose mode of action would be less 

important.

If linear extrapolation is used, the 

assessment develops a candidate slope factor 

or unit risk for each suitable data set. These 

results are arrayed, using common dose 

metrics, to show the distribution of relative 

potency across various effects and 

experimental systems. The assessment then 

derives or selects an overall slope factor and 

an overall unit risk for the agent, considering 

the various dose-response analyses, the 

study preferences discussed in Section 6, and 

the possibility of basing a more robust result 

on multiple data sets.

7.5. Considering susceptible 

populations and lifestages

The assessment analyzes the available 

information on populations and lifestages 

that may be particularly susceptible to each 

effect. A tiered approach is used (U.S. EPA, 

2005a, §3.5). 

This document is a draft for review purposes only and does not constitute Agency policy.

xxvi DRAFT—DO NOT CITE OR QUOTE
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reports quantitative results for a 

susceptible population or lifestage, these 

data are analyzed to derive separate 

toxicity values for susceptible 

individuals.

2) If data on risk-related parameters allow 

comparison of the general population and 

susceptible individuals, these data are 

used to adjust the general-population 

toxicity values for application to 

susceptible individuals.

3) In the absence of chemical-specific data, 

the EPA has developed age-dependent 

adjustment factors for early-life exposure 

to potential carcinogens that have a 

mutagenic mode of action. There is 

evidence of early-life susceptibility to 

various carcinogenic agents, but most 

epidemiologic studies and cancer 

bioassays do not include early-life 

exposure. To address the potential for 

early-life susceptibility, the EPA 

recommends (U.S. EPA, 2005b, §5): 

– 10-fold adjustment for exposures 

before age 2 years.

– 3-fold adjustment for exposures 

between ages 2 and 16 years.

7.6. Reference values and uncertainty 

factors

An oral reference dose or an inhalation 

reference concentration is an estimate of an 

exposure (including in susceptible 

subgroups) that is likely to be without an 

appreciable risk of adverse health effects 

over a lifetime (U.S. EPA, 2002, §4.2). 

Reference values are typically calculated for 

effects other than cancer and for suspected 

carcinogens if a well characterized mode of 

action indicates that a necessary key event 

does not occur below a specific dose. 

Reference values provide no information 

about risks at higher exposure levels.

The assessment characterizes effects that 

form the basis for reference values as 

adverse, considered to be adverse, or a 

precursor to an adverse effect. For

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oses only a

evelopmental toxicity, reproductive toxicity, 

nd neurotoxicity there is guidance on 

dverse effects and their biologic markers 

U.S. EPA, 1998, 1996, 1991). 

To account for uncertainty and variability 

 the derivation of a lifetime human 

xposure where adverse effects are not 

nticipated to occur, reference values are 

lculated by applying a series of uncertainty 

ctors to the point of departure. If a point of 

eparture cannot be derived by modeling, a 

o-observed-adverse-effect level or a lowestbserved-adverse-effect level is used instead. 

he assessment discusses scientific 

onsiderations involving several areas of 

ariability or uncertainty.

uman variation. The assessment accounts 

for variation in susceptibility across the 

human population and the possibility 

that the available data may not be 

representative of individuals who are 

most susceptible to the effect. A factor of 

10 is generally used to account for this 

variation. This factor is reduced only if 

the point of departure is derived or 

adjusted specifically for susceptible 

individuals (not for a general population 

that includes both susceptible and nonsusceptible individuals) (U.S. EPA, 

2002, §4.4.5; 1998, §4.2; 1996, §4; 

1994, §4.3.9.1; 1991, §3.4). 

nimal-to-human extrapolation. If animal 

results are used to make inferences about 

humans, the assessment adjusts for 

cross-species differences. These may 

arise from differences in toxicokinetics or 

toxicodynamics. Accordingly, if the point 

of departure is standardized to 

equivalent human terms or is based on 

toxicokinetic or dosimetry modeling, a 

factor of 101/2 (rounded to 3) is applied to 

account for the remaining uncertainty 

involving toxicokinetic and 

toxicodynamic differences. If a 

biologically based model adjusts fully for 

toxicokinetic and toxicodynamic 

differences across species, this factor is 

not used. In most other cases, a factor of 

10 is applied (U.S. EPA, 2011; 

This document is a draft for review purp nd does not constitute Agency policy.

xxvii DRAFT—DO NOT CITE OR QUOTE
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2002, §4.4.5; 1998, §4.2; 1996, §4; 

1994, §4.3.9.1; 1991, §3.4). 

dverse-effect level to no-observedadverse-effect level. If a point of 

departure is based on a lowest-observedadverse-effect level, the assessment must 

infer a dose where such effects are not 

expected. This can be a matter of great 

uncertainty, especially if there is no 

evidence available at lower doses. A

factor of 10 is applied to account for the 

uncertainty in making this inference. A

factor other than 10 may be used, 

depending on the magnitude and nature 

of the response and the shape of the doseresponse curve (U.S. EPA, 2002, §4.4.5; 

1998, §4.2; 1996, §4; 1994, §4.3.9.1; 

1991, §3.4). 

ubchronic-to-chronic exposure. If a point 

of departure is based on subchronic 

studies, the assessment considers 

whether lifetime exposure could have 

effects at lower levels of exposure. A

factor of 10 is applied to account for the 

uncertainty in using subchronic studies 

to make inferences about lifetime 

exposure. This factor may also be applied 

for developmental or reproductive effects 

if exposure covered less than the full 

critical period. A factor other than 10 

may be used, depending on the duration 

of the studies and the nature of the 

response (U.S. EPA, 2002, §4.4.5; 1998, 

§4.2; 1994, §4.3.9.1). 

ncomplete database. If an incomplete 

database raises concern that further 

studies might identify a more sensitive 

effect, organ system, or lifestage, the 

assessment may apply a database 

uncertainty factor (U.S. EPA, 2002, §4.4.5; 

1998, §4.2; 1996, §4; 1994, §4.3.9.1; 

1991, §3.4). The size of the factor 

depends on the nature of the database 

deficiency. For example, the EPA 

typically follows the suggestion that a 

factor of 10 be applied if both a prenatal 

toxicity study and a two-generation 

reproduction study are missing and a 

factor of 101/2 49 if either is missing (U.S. 

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EPA, 2002, §4.4.5). 

In this way, the assessment derives 

candidate values for each suitable data set 

and effect that is credibly associated with the 

agent. These results are arrayed, using 

common dose metrics, to show where effects 

occur across a range of exposures (U.S. EPA, 

1994, §4.3.9). 

The assessment derives or selects an 

organ- or system-specific reference value for 

each organ or system affected by the agent. 

The assessment explains the rationale for 

each organ/system-specific reference value 

(based on, for example, the highest quality 

studies, the most sensitive outcome, or a 

clustering of values). By providing these 

organ/system-specific reference values, IRIS 

assessments facilitate subsequent cumulative 

risk assessments that consider the combined 

effect of multiple agents acting at a common 

site or through common mechanisms (NRC, 

2009). 

The assessment then selects an overall 

reference dose and an overall reference 

concentration for the agent to represent 

lifetime human exposure levels where effects 

are not anticipated to occur. This is generally 

the most sensitive organ/system-specific 

reference value, though consideration of 

study quality and confidence in each value 

may lead to a different selection.

7.7. Confidence and uncertainty in the 

reference values

The assessment selects a standard 

descriptor to characterize the level of 

confidence in each reference value, based on 

the likelihood that the value would change 

with further testing. Confidence in reference 

values is based on quality of the studies used 

and completeness of the database, with more 

weight given to the latter. The level of 

confidence is increased for reference values 

based on human data supported by animal 

data (U.S. EPA, 1994, §4.3.9.2). 

High confidence: The reference value is not 

likely to change with further testing, 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 except for mechanistic studies that might 

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affect the interpretation of prior test 

results. 

Medium confidence: This is a matter of 

judgment, between high and low 

confidence.

Low confidence: The reference value is 

especially vulnerable to change with 

further testing.

These criteria are consistent with 

guidelines for systematic reviews that 

evaluate the quality of evidence. These also 

focus on whether further research would be 

likely to change confidence in the estimate of 

effect (Guyatt et al., 2008b). 

All assessments discuss the significant 

uncertainties encountered in the analysis. 

The EPA provides guidance on 

characterization of uncertainty (U.S. EPA, 

2005a, §3.6). For example, the discussion 

distinguishes model uncertainty (lack of 

knowledge about the most appropriate 

experimental or analytic model) and 

parameter uncertainty (lack of knowledge 

about the parameters of a model). 

Assessments also discuss human variation 

(interpersonal differences in biologic 

susceptibility or in exposures that modify the 

effects of the agent).

August 2013

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EXECUTIVE SUMMARY

Summary of Occurrence and Health Effects

Hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) is a synthetic chemical used 

primarily as a military explosive. RDX releases have been reported in air, water, and 

soil. Exposure to RDX is likely limited to individuals in or around military facilities 

where RDX is or was produced, used, or stored. Oral exposure may occur from 

drinking contaminated groundwater or ingesting crops irrigated with contaminated 

water. Inhalation or dermal exposures are more likely in occupational settings. 

Epidemiological studies provide only limited information on worker 

populations exposed to RDX; several case reports describe effects primarily in the 

nervous system following acute exposure to RDX. Animal studies of ingested RDX 

demonstrate toxicity, including nervous system effects, kidney and other urogenital 

effects, and male reproductive effects. 

Results from animal studies provide suggestive evidence of carcinogenic 

potential for RDX based on evidence of positive trends in liver and lung tumor 

incidence in experimental animals. There are no data on the carcinogenicity of RDX 

in humans. 

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Effects Other Than Cancer Observed Following Oral Exposure

Nervous system effects are a human hazard of RDX exposure. Several human case reports 

and animal studies provide consistent evidence of an association between RDX exposure and effects 

on the nervous system, including seizures or convulsions, tremors, hyperirritability, hyperreactivity, and behavioral changes. Mechanistic data support the hypothesis that RDX-induced 

hyperactivity and seizures likely result from inhibition of GABAergic signaling in the limbic system. 

Kidney and other urogenital effects are a potential human hazard of RDX exposure based on 

observations in 2-year oral toxicity studies of increased relative kidney weights in male and female 

mice and histopathological changes in the urogenital system of male rats exposed to RDX. An 

increased incidence of suppurative prostatitis was identified, and is considered a marker for RDXrelated urogenital effects. There is no established mode of action (MOA) for RDX-related effects on 

the urogenital system.

There is suggestive evidence of male reproductive effects associated with RDX exposure 

based on the finding of testicular degeneration in male mice exposed to RDX in the diet for 2 years, 

in the only mouse study conducted of that duration. There is no known MOA for male reproductive 

effects of RDX exposure. Evidence for effects on other organs/systems, including the liver and 

developmental effects, was more limited than for the endpoints summarized above. 

This document is a draft for review purposes only and does not constitute Agency policy.

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Oral Reference Dose (RfD) for Effects Other Than Cancer 

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Table ES-1). These organ- or system-specific reference values may be useful for subsequent 

cumulative risk assessments that consider the combined effect of multiple agents acting at a 

common site.

Table ES-1. Organ/system-specific RfDs and overall RfD for RDX

Effect Basis RfD (mg/kg-d)

Study exposure 

description Confidence

Nervous system Convulsions 3 × 10−3 Subchronic Medium

Kidney/urogenital Suppurative prostatitis 8 × 10−3 Chronic Low

Male reproductive Testicular degeneration 8 × 10−2 Chronic Low

Overall RfD Nervous system effects 3 × 10−3 Subchronic Medium

The overall RfD (see Table ES-2) is derived to be protective of all types of hazards 

associated with RDX exposure. The effect of RDX on the nervous system was chosen as the basis for 

the overall RfD because nervous system effects were observed most consistently across studies, 

species, and exposure durations, and because they represent the most sensitive human hazard of 

RDX exposure. Incidence of seizures or convulsions as reported in a subchronic gavage study

(Crouse et al., 2006) was selected for derivation of the overall RfD as this endpoint was measured in 

a study that was well-conducted, utilized a test material of higher purity than other studies, and had 

five closely-spaced dose groups that allowed characterization of the dose-response curve. 

Benchmark dose (BMD) modeling was utilized to derive the point of departure (POD) for RfD 

derivation (expressed as the BMDL01). A 1% response level was chosen because of the severity of 

the endpoint. Further, the doses associated with nervous system effects in the Crouse et al. (2006)

study also caused increased mortality in the animals. Experimental animal studies provide some 

evidence of an association between RDX-induced convulsions and mortality. In three studies in rats 

(Crouse et al., 2006; Levine et al., 1983b; Cholakis et al., 1980), investigators noted that early deaths 

were frequently preceded by neurotoxic signs such as tremors and convulsions; however, the 

recorded data from Crouse et al. (2006) do not show as clear a correspondence between 

convulsions (and other neurotoxic signs) and mortality (Section 1.2.1). 

 A physiologically-based pharmacokinetic (PBPK) model was used to extrapolate the 

BMDL01 derived from a rat study to a human equivalent dose (HED) based on RDX arterial blood 

concentration, which was then used for RfD derivation.

The overall RfD was calculated by dividing the BMDL01-HED for nervous system effects by a 

composite uncertainty factor (UF) of 100 to account for extrapolation from animals to humans (3), 

interindividual differences in human susceptibility (10), and uncertainty in the database (3). 

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1 Table ES-2. Summary of reference dose (RfD) derivation

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Critical effect Point of departurea UF Chronic RfD

Nervous system effects (convulsions)

90-d F344 rat study

Crouse et al. (2006)

BMDL01-HED: 0.28 mg/kg-d 100 3 × 10−3 mg/kg-d

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A benchmark response (BMR) of 1% was used to derive the BMD and BMDL given the severity of the endpoint. 

The resulting POD was converted to a BMDL01-HED using a PBPK model based on modeled arterial blood 

concentration. The concentration was derived from the area under the curve (AUC) of modeled RDX 

concentration in arterial blood, which reflects the average blood RDX concentration for the exposure duration 

normalized to 24 hours.

Effects Other Than Cancer Observed Following Inhalation Exposure

No studies were identified that provided useful information on the effects observed 

following inhalation exposure to RDX. Of the available human epidemiological studies of RDX, none 

provided data that could be used for dose-response analysis of inhalation exposures. The single 

experimental animal study involving inhalation exposure is not publicly available, and was 

excluded from consideration due to significant study limitations, including small numbers of 

animals tested, lack of controls, and incomplete reporting of exposure levels. Therefore, the 

available health effects literature does not support the identification of hazards following inhalation 

exposure to RDX.

Inhalation Reference Concentration (RfC) for Effects Other Than Cancer

An RfC for RDX could not be derived based on the available health effects data. While 

inhalation absorption of RDX particulates is a plausible route of exposure, there are no toxicokinetic 

studies of RDX inhalation absorption to support an inhalation model. Therefore, a PBPK model for 

inhaled RDX was not developed to support route-to-route extrapolation of an RfC from the RfD.

Evidence for Human Carcinogenicity

Under EPA’s cancer guidelines (U.S. EPA, 2005a), there is suggestive evidence of carcinogenic

potential for RDX. RDX induced benign and malignant tumors in the liver and lungs of mice (Parker 

et al., 2006; Lish et al., 1984) or rats (Levine et al., 1983b) following long-term administration in the 

diet. The potential for carcinogenicity applies to all routes of human exposure.

Quantitative Estimate of Carcinogenic Risk from Oral Exposure

A quantitative estimate of carcinogenic risk from oral exposure to RDX was based on the

increased incidence of hepatocellular adenomas or carcinomas and alveolar/bronchiolar adenomas 

or carcinomas in female B6C3F1 mice observed in the carcinogenicity bioassay in mice (Lish et al., 

1984). This 2-year dietary study included four dose groups and a control group, adequate numbers 

of animals per dose group (85/sex/group, with interim sacrifices of 10/sex/group at 6 and 

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1 12 months), and detailed reporting of methods and results (including individual animal data). The 

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initial high dose (175 mg/kg-day) was reduced to 100 mg/kg-day at week 11 due to high mortality. 

Considering these data along with the uncertainty associated with the suggestive nature of 

the weight of the evidence for RDX carcinogenicity, quantitative analysis of the mouse tumor data

may be useful for providing a sense of the magnitude of potential carcinogenic risk.

An oral slope factor (OSF) that considered the combination of female mouse liver and lung 

tumors was derived from BMD and BMDL estimates that correspond to a 10% extra risk (ER) of 

either tumor. The BMDL10 so derived was extrapolated to the HED using BW3/4 scaling, and an OSF 

was derived by linear extrapolation from the BMDL10-HED. The OSF is 0.04 per mg/kg-day, based on 

the liver and lung tumor response in female mice (Lish et al., 1984). 

Quantitative Estimate of Carcinogenic Risk from Inhalation Exposure

An inhalation unit risk (IUR) value was not calculated because inhalation carcinogenicity 

data for RDX are not available. While inhalation absorption of RDX particulates is a plausible route 

of exposure, there are no toxicokinetic studies of RDX inhalation absorption to support an 

inhalation model. Therefore, a PBPK model for inhaled RDX was not developed to support route-toroute extrapolation of an IUR from the OSF. Thus, a quantitative cancer assessment was not 

conducted. 

Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes 

Little information is available on populations that may be especially vulnerable to the toxic 

effects of RDX. Lifestage, and in particular childhood, susceptibility has not been observed in 

human or animal studies of RDX toxicity. In rats, transfer of RDX from the dam to the fetus during 

gestation and to pups via maternal milk has been reported; however, reproductive and 

developmental toxicity studies did not identify effects in offspring at doses below those that also 

caused maternal toxicity. Data to suggest that males may be more susceptible than females to 

noncancer toxicity associated with RDX exposure are limited. Specifically, urogenital effects have 

been noted at lower doses in males than in females. Data on the incidence of convulsions and 

mortality provide some indication that pregnant animals may be a susceptible population, although 

the evidence is unclear. Some evidence suggests that cytochrome P450 (CYP450) enzymes may be 

involved in the metabolism of RDX, indicating a potential for genetic polymorphisms in these 

metabolic enzymes to affect susceptibility to RDX. Similarly, individuals with epilepsy or other 

seizure syndromes that have their basis in genetic mutation to GABAA receptors may represent 

another group that may be susceptible to RDX exposure; however, there is no information to 

indicate how genetic polymorphisms may affect susceptibility to RDX. 

Key Issues Addressed in Assessment

Selection of a 1% benchmark response (BMR) for convulsions. In most instances, the 

spectrum of effects associated with chemical exposure will range in severity, with relatively less 

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1 severe effects generally occurring at doses lower than those associated with more severe or “frank” 

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toxicity. Convulsions in rats were selected as the basis for derivation of the RDX RfD; less severe 

nervous system effects were generally not observed at lower doses. U.S. EPA (2012b) recommends 

considering the statistical and biological characteristics of the dataset when selecting a BMR, 

including the severity of the effect. For convulsions, a BMR level of 1% ER was selected for 

modeling, balancing the quantitative limitations of the available animal bioassays and the severity 

of this effect. Modeling convulsion incidence from Crouse et al. (2006) using this BMR resulted in a 

moderate extrapolation of the BMD (3.0 mg/kg-day) below the range of experimental data (dose 

range from Crouse et al. (2006): 4−15 mg/kg-day). 

Influence of the method of oral dosing (diet and gavage). Some uncertainty in the RfD is 

also associated with the influence of the method of oral dosing on the magnitude of dose required 

to induce nervous system effects. Findings from animal studies suggest that gavage administration 

generally induced convulsions in experimental animals at lower doses than did dietary 

administration, possibly due to the bolus dose received from gavage administration resulting in a

comparatively faster absorption and higher peak blood concentrations of RDX (see Section 1.2.1). 

The difference in neurotoxic response associated with gavage versus dietary administration is in 

part reflected in the 14-fold difference in the candidate PODHED values derived from the Crouse et al. 

(2006) (gavage administration) and Levine et al. (1983b) (dietary administration) studies (see 

Table 2-2). A more rigorous examination of the effect of oral dosing method cannot be performed 

because of the differences across studies in test materials and experimental designs (e.g., test article 

purity and particle size, number and spacing of dose groups, exposure duration, frequency of 

clinical observations, and thoroughness of the reporting of observations) that could also have 

contributed to differences in response. As dietary administration is more representative of 

potential human exposures to RDX, the use of toxicity data from a gavage (bolus dosing) study may 

introduce uncertainty in the RfD. 

Suppurative prostatitis as a marker for kidney and other urogenital effects. The 

candidate RfD for kidney and other urogenital effects is based on a dose-related increase in the 

incidence of suppurative prostatitis from a 2-year feeding study in male F344 rats (Levine et al., 

1983b). This study is the only 2-year study in rats that examined the prostate. Some reports have 

hypothesized that the observed suppurative prostatitis is a secondary effect from a bacterial 

infection unrelated to RDX toxicity (ATSDR, 2012; Sweeney et al., 2012a; Crouse et al., 2006). While 

an opportunistic bacterial infection may have been the proximal cause of the suppurative 

prostatitis, the infection was considered secondary to urogenital effects associated with RDX 

exposure. Histopathological findings for the bladder are not definitive because the design of the 

principal study called for histopathological examination of the bladder only if gross abnormalities 

were observed. Although the pathogenesis of kidney and urogenital effects is unclear, suppurative 

prostatitis was considered to be a marker for the broader array of kidney and other urogenital 

effects observed by Levine et al. (1983b). 

This document is a draft for review purposes only and does not constitute Agency policy.

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LITERATURE SEARCH STRATEGY | 

STUDY SELECTION AND EVALUATION

Literature Search and Screening Strategy

1 A literature search and screening strategy was applied to identify literature related to 

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characterizing the health effects of hexadydro-1,3,5-trinitro-1,3,5-triazine (RDX). This strategy 

consisted of a search of online scientific databases and other sources, casting a wide net in order to 

identify all potentially pertinent studies. In subsequent steps, references were screened to exclude 

papers not pertinent to an assessment of the chronic health effects of RDX, and the remaining 

references were sorted into categories for further evaluation. 

The literature search for RDX was conducted in four online scientific databases—PubMed, 

Toxline, Toxcenter, and Toxic Substances Control Act Test Submissions (TSCATS). The initial 

search was performed in April 2012, and literature search updates were conducted in February 

2013, January 2014, and January 2015. Searches of TSCATS were performed in February 2013 and 

January 2015 only. The detailed search approach for these databases, including the query strings, 

and the numbers of citations identified per database are provided in Appendix B, Table B-1. The 

Department of Defense has conducted several unpublished toxicological studies on RDX; to ensure 

that all such studies were located, the Defense Technical Information Center (DTIC) database, a 

central online repository of defense-related scientific and technical information within the 

Department of Defense, was also searched. A separate strategy was applied in searching DTIC 

because of limitations in the classification and distribution of materials in DTIC; the detailed search 

strategy is described in Appendix B, Table B-2. Searches of the five online databases identified 

1,143 citations (after electronically eliminating duplicates). The computerized database searches 

were supplemented by reviewing online regulatory sources, performing “forward” and “backward” 

searches of Web of Science (see Appendix B, Table B-3), and adding additional references that were 

identified during the development of the Toxicological Review (including submissions from the 

Department of Defense); 34 citations were obtained using these additional search strategies. In 

total, 1,177 citations were identified using online scientific databases and additional search 

strategies.

The U.S. Environmental Protection Agency (EPA) requested public submissions of 

additional information in 2010 (75 FR 76982; December 10, 2010). EPA also issued a request to 

the public for additional information in a Federal Register Notice in 2013 (78 FR 48674; August 9, 

2013), and established a docket for public comment (EPA-HQ-ORD-2013-0430; available at 

www.regulations.gov) maintained through the development of the assessment. No submissions 

were received in response to these calls for data. 

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1 The citations identified using the search strategy described above were screened using the 

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title, abstract, and in limited instances, full text for pertinence to examining the health effects of 

chronic RDX exposure. The process for screening the literature is described below and is shown 

graphically in Figure LS-1.1 The objective of this manual screen was to identify sources of primary 

human health effects data and sources of primary data that inform the assessment of RDX health 

effects (i.e., the bottom three boxes in Figure LS-1). Inclusion and exclusion criteria used to 

manually screen the references in order to identify health effect studies (i.e., the green boxes in 

Figure LS-1) are provided in Table LS-1. Specific inclusion criteria were not applied in identifying 

sources of mechanistic and toxicokinetic data. The number of such studies for RDX is not large, and 

therefore, all studies that provided data on adsorption, distribution, metabolism, or elimination, 

physiologically-based pharmacokinetic (PBPK) models, or relevant RDX mode of action (MOA) 

were considered. Studies that met one or more of the exclusion criteria in Table LS-1 were binned 

as “Excluded/Not on Topic” and were not further considered in this assessment. A final group of 

studies consisted of reviews and other sources of RDX information (e.g., exposure, ecosystem 

effects) that did not meet the inclusion criteria in Table LS-1. These studies were binned into a 

category called “Secondary Literature and Sources of Other RDX Information,” and were considered 

as appropriate during development of this assessment. 

The results of this literature screening are described below and graphically in Figure LS-1:

• 25 references (including both human and animal studies) were identified as sources of 

health effects data and were considered for data extraction to evidence tables and 

exposure-response arrays.2

• 25 references were identified as sources of supplementary health effects data, including 

experimental animal studies involving acute or short-term exposures or dermal 

exposure, and human case reports. Studies investigating the effects of acute/short-term 

and dermal exposures and case reports are generally less pertinent for characterizing 

health hazards associated with chronic oral and inhalation exposure. Therefore, 

information from these studies was not extracted into evidence tables. Nevertheless, 

these studies were still considered as possible sources of supplementary health effects 

information.

1Studies were assigned (or “tagged”) to a given category in Health and Environmental Research Online 

(HERO) that best reflected the primary content of the study. In general, studies were not assigned multiple 

tags in order to simplify the tracking of references. Nevertheless, the inclusion of a citation in a given 

category (or tag) did not preclude its use in one or more other categories. For example, Woody et al. (1986), a 

case report of accidental ingestion of RDX by a child, was tagged to the human case reports under

Supplementary Studies in Figure LS-1. This case report also provides pharmacokinetic data and was a 

pertinent source of information on RDX toxicokinetics, but was not assigned a second tag for toxicokinetics. 2 On the HERO project page, 27 records are associated with “Sources of Health Effects Data,” rather than the 

25 identified in Figure LS-1. Two of the records in HERO are not unique studies or reports. Rather, these two 

records provide links to the multi-volume laboratory reports for the 2-year studies in rats by Levine et al. 

(1983b) and mice by Lish et al. (1984)

This document is a draft for review purposes only and does not constitute Agency policy.

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1 

• 47 references were identified as sources of mechanistic and toxicokinetic data; these 

included 19 studies describing PBPK models and other toxicokinetic information, 

11 studies providing genotoxicity information, and 17 studies pertaining to other 

mechanistic information. Information from these studies was not extracted into 

evidence tables; however, these studies supplemented the assessment of RDX health 

effects (e.g., evaluation of MOA and extrapolation of experimental animal findings to 

humans). 

• 280 references were identified as secondary literature (e.g., reviews and other agency 

assessments) or as studies providing potentially useful information on RDX (e.g., studies 

providing information on exposure levels or effects on nonmammalian species); these 

references were kept as additional resources for development of the Toxicological 

Review. 

• 801 references were identified as not being pertinent (or not on topic) to an evaluation 

of the chronic health effects of RDX and were excluded from further consideration (see 

Figure LS-1 and Table LS-1 for exclusion criteria). Retrieving a large number of 

references that are not on topic is a consequence of applying an initial search strategy 

designed to cast a wide net and to minimize the possibility of missing potentially 

relevant health effects data. 

This document is a draft for review purposes only and does not constitute Agency policy.

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The numbers on this figure match the HERO project page as of 1/12/2016; subsequent changes may not be 

reflected. A limited number of references were assigned more than one tag; therefore, the sum of the references 

in boxes below “Manual Screening for Pertinence” does not match exactly the total number of references in the 

“Combined Dataset.”

Figure LS-1. Summary of literature search and screening process for RDX.

This document is a draft for review purposes only and does not constitute Agency policy.

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1 Table LS-1. Inclusion-exclusion criteria for health effect studies

Inclusion criteria Exclusion criteria

Population • Humans

• Standard mammalian animal models, 

including rat, mouse, rabbit, guinea pig, 

monkey, dog

• Ecological speciesa

• Nonmammalian speciesa

Exposure • Exposure is to RDX

• Exposure is measured in an 

environmental medium (e.g., air, water, 

diet)

• Exposure via oral or inhalation routes

• Study population is not exposed to RDX

• Exposure to a mixture only 

• Exposure via injection (e.g., intravenous [i.v.]) 

Outcome • Study includes a measure of one or 

more health effect endpoints, including 

effects on the nervous, 

kidney/urogenital, musculoskeletal, 

cardiovascular, immune, and 

gastrointestinal systems, reproduction, 

development, liver, eyes, and cancer

Other Not on topic, including:

• Abstract only, inadequately reported abstract, or no 

abstract, and not considered further because study 

was not potentially relevant

• Bioremediation, biodegradation, or chemical or 

physical treatment of RDX and other munitions, 

including evaluation of wastewater treatment 

technologies and methods for remediation of 

contaminated water and soil

• Chemical, physical, or explosive properties, including 

studies of RDX crystal quality, energetics 

characteristics, sublimation kinetics, isotope ratios, 

and thermal decomposition and other explosive 

properties

• Analytical methods for measuring/detecting/ 

remotely sensing RDX in environmental media, and 

use in sample preparations and assays

• Not chemical specific (studies that do not involve 

testing of RDX) 

• Other studies not informative for evaluating RDX 

health effects and not captured by other exclusion 

criteria, including: 

-- Superfund site records of decision that describe 

remedial action plans for waste sites

-- characterization of waste sites contaminated by 

explosives

-- foreign language studies where translation was not 

warranted because, based on title or abstract, the 

added value to the evaluation of RDX health effects 

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Inclusion criteria Exclusion criteria

was considered small (e.g., Chinese paper of case 

reports of RDX poisonings)

-- duplicate studies not previously identified

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a

Studies that met this exclusion criterion were not considered a source of health effects or supplementary health 

effects data, but were considered as other sources of information potentially useful in assessing the health effects 

of RDX.

The documentation and results for the literature search and screen can be found on the 

Health and Environmental Research Online (HERO) website on the RDX project page at: 

(http://hero.epa.gov/index.cfm/project/page/project_id/2216).

Selection of Critical Studies for Presentation in Evidence Tables

Selection of Critical Studies 

In order to systematically summarize the important information from the primary health 

effects studies in the RDX database, evidence tables were constructed in a standardized tabular 

format as recommended by the NRC (2011). Of the studies that were retained after the literature 

search and screen, 25 were categorized as “Sources of Health Effects Data” (Figure LS-1, Table LS-1) 

and were considered for extraction into evidence tables for hazard identification in Chapter 1.

A study was not presented in the evidence tables if flaws in its design, conduct, or reporting 

were so great that the results would not be considered credible (e.g., studies where concurrent 

control information is lacking). Such study design flaws are discussed in a number of EPA’s 

guidelines (see http://www.epa.gov/iris/backgrd.html) or summarized in the Preamble. For RDX, 

four studies were considered uninformative and were removed from further consideration in the 

assessment because of fundamental issues with study design, conduct, or reporting. The specific 

studies and basis for considering the studies to be uninformative are summarized in Table LS-2. 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 Table LS-2. Studies determined not to be informative because of significant 

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issues with design, conduct, or reporting

Reference Rationale for exclusion

Haskell Laboratories (1942);

14-wk study in dogs

Incomplete information on exposure levels; breed of dog was not 

reported; inadequate reporting of results; sections of document 

were illegible.

von Oettingen et al. (1949); 

10-wk oral study in rats

No control group; strain of rat was not reported.

ATSDR (1996); 

Disease prevalence study in residential 

population

Study of a population residing in two neighborhoods where RDX 

had been detected in well water. The study was conducted 7 yrs 

after residents were provided the opportunity to connect to a 

municipal water supply. Only one target-area household reported 

using private well water for bathing and cooking at the time of the 

health study. The study was not considered informative because 

the design was not able to adequately define the exposed 

population. 

Unpublished report (dated 1944) from the 

DTIC database; 

Human and animal data

One section of the report describes a human case series with no 

referent group. Issues with the inhalation experimental animal 

studies included lack of control groups, incomplete information on 

exposure levels, and inadequate reporting of results. [Because this 

report is classified as a limited distribution document in the DTIC 

database, it was not added to the HERO project page for RDX.]

The health effects literature for RDX is not extensive. With the exception of the studies 

listed in Table LS-2 (i.e., those determined to be uninformative), all human and experimental animal 

studies of RDX involving repeated exposure were considered in assessing the evidence for health 

effects associated with chronic exposure to RDX. 

Studies that contain pertinent information for the toxicological review and augment hazard 

identification conclusions, such as genotoxicity and other mechanistic studies, studies describing 

the toxicokinetics of RDX, human case reports, and experimental animal studies involving 

exposures of acute/short-term duration or routes of exposure other than oral and inhalation, were 

not included in evidence tables. Nevertheless, these studies were considered, where relevant, in the 

evaluation of RDX health hazards.

This document is a draft for review purposes only and does not constitute Agency policy.

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Study Evaluation

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studies3 and 21 reports4 presenting results of experimental animal studies. These studies were 

evaluated using the study quality considerations outlined in the Preamble, considering aspects of 

design, conduct, or reporting that could affect the interpretation of results, overall contribution to 

the synthesis of evidence, and determination of hazard potential as noted in various EPA guidance 

documents (U.S. EPA, 2005a, 2002, 1994). The objective was to identify the stronger, more 

informative studies based on a uniform evaluation of quality characteristics across studies of 

similar design. 

Additionally, a number of general questions, presented in Table LS-3, were considered in 

evaluating the animal studies. Much of the key information for conducting this evaluation can be 

determined based on study methods and how the study results were reported. Importantly, the 

evaluation at this stage does not consider the direction or magnitude of any reported effects. 

Table LS-3. Considerations and relevant experimental information for 

evaluation of experimental animal studies 

Methodological 

feature

Considerations 

(relevant information extracted into evidence tables)

Test animal Suitability of the species, strain, sex, and source of the test animals

Experimental design Suitability of animal age/lifestage at exposure and endpoint testing; periodicity and 

duration of exposure (e.g., hrs/d, d/wk); timing of endpoint evaluations; and sample size 

and experimental unit (e.g., animals, dams, litters)

Exposure Characterization of test article source, composition, purity, and stability; suitability of the 

control (e.g., vehicle control); documentation of exposure techniques (e.g., route, 

chamber type, gavage volume); verification of exposure levels (e.g., consideration of 

homogeneity, stability, analytical methods)

Endpoint evaluation Suitability of specific methods for assessing the endpoint(s) of interest

Results presentation Data presentation for endpoint(s) of interest (including measures of variability) and for 

other relevant endpoints needed for results interpretation (e.g., maternal toxicity, 

decrements in body weight in relation to organ weight)

3Two reports with human data were determined not to be informative; see Table LS-2. The study by ATSDR 

(1996) was included in HERO and in Figure LS-1. The unpublished report from the DTIC database was not 

included in either HERO or Figure LS-1 because this report is classified as a limited distribution document in 

DTIC. This accounts for the three human studies being reviewed for study evaluation rather than the four 

identified in the literature search (see Figure LS-1). 4The number of reports of experimental animal studies does not equal the number of studies. The results of 

some studies were documented in multiple reports (e.g., a 2-year study in F344 rats by Levine et al. (1983b)

was published in three volumes). The Cholakis et al. (1980) study included, in a single report, subchronic 

studies in rats and mice, a 2-generation reproductive toxicity study in rats, and developmental toxicity 

studies in rats and rabbits. 

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1 Information on study features related to this evaluation is reported in evidence tables and 

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was considered in the synthesis of evidence. Discussion of study strengths and limitations (that 

ultimately supported preferences for the studies and data relied upon) were included in the text 

where relevant. If EPA’s interpretation of a study differed from that of the study authors, the 

assessment discusses the basis for the difference.

The general findings of this evaluation are presented in the remainder of this section. Study 

evaluation considerations that are outcome specific are discussed in the relevant health effect 

sections in Section 1.2. 

Human Studies

The body of literature on RDX includes three studies of populations occupationally exposed 

to RDX (one case-control and two cross-sectional studies) (West and Stafford, 1997; Hathaway and 

Buck, 1977). To varying degrees, these epidemiology studies are limited in their ability to assess 

the relationship between RDX exposure and the incidence of human health effects. Some studies 

lacked information related to study design, such as a precise definition of the study population, 

while others did not include a comprehensive exposure assessment or details regarding potential 

confounders. All three studies had small sample sizes (60−69 exposed workers in the crosssectional studies and 32 cases in the case-control study), which limits their statistical power when 

comparing exposed workers or cases and unexposed or control participants. 

The study by Ma and Li (1993) of Chinese industrial workers provided limited information 

on participant recruitment, selection, and participation rate; the available information was not 

adequate to evaluate the potential for selection bias. Also, no information on adjustment for coexposure to trinitrotoluene (TNT) or other neurological risk factors (e.g., alcohol consumption) was 

provided. The study by Hathaway and Buck (1977) included details on exposure assessment, but 

did not provide information on length of employment or other metrics that could be used to 

ascertain duration of exposure. In the case-control study by West and Stafford (1997), RDX was 

identified as one of the many chemicals that workers may have been exposed to in the ordnance 

factory. Thus, there is a potential for co-exposure to other chemicals that may elicit the observed 

effects. The methodological limitations in these three studies were considered in the synthesis of 

evidence for each of the health effects and in reaching determinations of hazard (see Section 1.2). 

In addition to the aforementioned studies, the human health effects literature includes 

16 case reports that describe effects following acute exposure to RDX. Case reports can suggest 

organ systems and health outcomes that might be related to RDX exposure but are often anecdotal, 

and typically describe unusual or extreme exposure situations; thus, they provide little information 

that would be useful for characterizing chronic health effects. Therefore, RDX case reports were 

only briefly reviewed; a critical evaluation was not undertaken. A summary of these case reports is 

provided in Appendix C, Section C.2. 

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Experimental Animal Studies

1 The oral toxicity database for RDX includes three chronic studies in rats and mice, eight 

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subchronic studies in rats, mice, dogs, and monkeys, two shorter-term studies in dogs and rats, one 

two-generation reproductive toxicity study in the rat, four developmental toxicity studies in rats 

and rabbits, and a single-exposure study of audiogenic seizures in rats (Table LS-4). 

Table LS-4. Summary of experimental animal database

Study category Study duration, species/strain, and oral administration method

Chronic 2-Yr study in B6C3F1 mice (diet) (Lish et al., 1984)

2-Yr study in Sprague-Dawley rats (diet) (Hart, 1976) 

2-Yr study in F344 rats (diet) (Levine et al., 1983b)

Subchronic 13-Wk study in B6C3F1 mice, experiment 1 (diet) (Cholakis et al., 1980)

13-Wk study in B6C3F1 mice, experiment 2 (diet) (Cholakis et al., 1980) 

13-Wk study in F344 rats (diet) (Cholakis et al., 1980) 

13-Wk study in F344 rats (diet) (Levine et al., 1990; Levine et al., 1981a, b) 

13-Wk study in F344 rats (gavage) (Crouse et al., 2006) 

13-Wk study in rats, strain not specified (diet) (von Oettingen et al., 1949) 

13-Wk study in beagle dogs (diet) (Hart, 1974) 

13-Wk study in monkeys (gavage) (Martin and Hart, 1974) 

6-Wk study in dogs, breed not specified (diet) (von Oettingen et al., 1949) 

30-D study in Sprague-Dawley rats (gavage) (MacPhail et al., 1985)

Reproductive 2-Generation reproductive toxicity study in CD rats (diet) (Cholakis et al., 1980)

Developmental Developmental study (gestational days [GDs] 6−19) in F344 rats (gavage) (Cholakis et al., 1980)

Developmental study (GDs 6−15) in Sprague-Dawley rats, range-finding (gavage) (Angerhofer et 

al., 1986) 

Developmental study (GDs 6−15) in Sprague-Dawley rats (gavage) (Angerhofer et al., 1986) 

Developmental study (GDs 7−29) in New Zealand White (NZW) rabbits (gavage) (Cholakis et al., 

1980)

Nervous system 8-Hr study of audiogenic seizures in Long Evans rats (gavage) (Burdette et al., 1988)

 With the exception of two studies (Levine et al., 1990; von Oettingen et al., 1949), these 

toxicity studies are available only as unpublished contract laboratory reports. Peer reviews of 

three unpublished studies identified as most informative to the assessment of the health effects of 

RDX—the 2-year bioassays by Levine et al. (1983b) and Lish et al. (1984) and the subchronic 

toxicity study by Crouse et al. (2006)—were conducted by Versar, Inc. for EPA in 2012. The report 

of the peer reviews (U.S. EPA, 2012d) is available at https://epa.gov/hero. The peer reviewers 

generally concluded that the 2-year bioassay reports provided useful information on the toxicity of 

RDX, noting that there were limitations in interpretation due to aspects of the histopathological 

analysis and the statistical approaches employed. The peer reviewers similarly determined that the 

report by Crouse et al. (2006) provided useful information on RDX toxicity, including an array of 

endpoints for neurotoxicity and immunotoxicity (U.S. EPA, 2012d). 

Only one unpublished inhalation study of RDX (dated 1944) was identified. As discussed in 

Appendix B and Table LS-2, this inhalation study was considered uninformative and was excluded 

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1 from consideration in the development of the Toxicological Review because of study design issues 

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(including lack of a control group, incomplete information on exposure levels, and inadequate 

reporting). Therefore, evaluation of the experimental animal database for RDX is limited to studies 

of oral toxicity. An evaluation of the oral toxicity literature, organized by general methodological 

features, is provided in the remainder of this section. 

Test animal

The RDX database consists of health effect studies conducted in multiple strains of rats 

(F344, Sprague-Dawley, CD), mice (B6C3F1), dogs (beagle), and monkeys. The species and strains 

of animals used are consistent with those typically used in laboratory studies. All of these species 

or strains were considered relevant to assessing the potential human health effects of RDX. Several 

studies in the RDX database provided inadequate information on test animals. The strain of 

monkey (Rhesus or Cynomolgus) used in the study by Martin and Hart (1974) was not clearly 

specified. In one study, the breed of dog and strain of rat were unreported (von Oettingen et al., 

1949). The species, strain, and sex of the animals used are recorded in the evidence tables. 

Other studies of RDX were identified that used nonstandard species, including deer mice 

(Peromyscus maniculatus), western fence lizards (Sceloporus occidentalis), prairie voles (Microtus 

ochrogaster), and northern bobwhite quail (Colinus virginianus). These studies provide information

relevant to RDX toxicokinetics and mechanism of action on the nervous system, but not health 

effects data. Therefore, these studies are not included in evidence tables, but are discussed where 

relevant in the assessment.

Experimental setup

General aspects of study design and experimental setup were evaluated for all studies that 

included health effects data to determine if they were appropriate for evaluation of specific 

endpoints. Key features of the experimental setup, including the periodicity and duration of 

exposure, timing of exposure (e.g., gestational days for developmental studies), experimental group 

sample sizes, and interim sacrifices are summarized in the evidence tables. Note that sample size 

was not a basis for excluding a study from consideration, as studies with a small number of animals

can still inform the consistency of effects observed for a specific endpoint. Nevertheless, the 

informativeness of studies with small sample sizes, e.g., three animals/sex/group in the case of Hart 

(1974) and Martin and Hart (1974), was reduced. Elements of the experimental setup that could 

influence interpretation of study findings are discussed in the relevant hazard identification 

sections of the assessment.

Exposure

Properties of the test material were also considered in determining whether the exposures 

were sufficiently specific to the compound of interest. Two properties of the RDX test materials 

that varied across experimental animal studies and that were taken into consideration in evaluating 

the evidence for RDX hazards are the particle size and purity of the test material. The purity of RDX 

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1 used in health effects studies varied from 84 to 99.99%. The major contaminants were octahydro2 

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1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX) and water, which are the primary contaminants of RDX 

produced during manufacturing. The majority of studies used RDX with ~10% impurities; only

Crouse et al. (2006) used 99.99% pure RDX as a test material in their study. The toxicity of HMX 

was assessed by the Integrated Risk Information System (IRIS) Program in 1988

(http://cfpub.epa.gov/ncea/iris2/chemicalLanding.cfm?substance_nmbr=311); histopathological 

changes in the liver in male F344 rats and in the kidney in female rats were reported in a 13-week 

feeding study. No chronic studies were available to evaluate the carcinogenicity of HMX. The 

presence of the impurities introduces some uncertainty in attribution of toxicity to RDX. However, 

consistency in the doses at which some toxic effects were seen across studies suggests that the 

uncertainty associated with the use of less pure test materials may be relatively small. Evidence of 

neurotoxic effects in the study with 99.99% pure RDX occurred at doses of 8−15 mg/kg-day; 

studies with less pure RDX reported similar symptoms at doses ≥20 mg/kg-day. It should be noted 

that the test materials employed in these studies (i.e., with ~10% impurities) are consistent with

the purity of RDX that would be released into the environment. 

Differences in milling procedures used to generate the test material resulted in the use of 

RDX of varying particle sizes across studies. Some studies utilized a test material with a relatively 

fine particle size (majority of particles <66 μm in size), while others used a test material with 

comparatively coarse particle size (~200 μm particle size). Differences in particle size across 

studies could result in different rates of absorption of RDX into the blood stream, which could 

account for differences in response observed across studies, including neurotoxicity. Information 

on test material purity and particle size, as provided by study authors, is reported in the evidence 

tables, and was considered in evaluating the toxicity of RDX. The lack of characterization of the test 

material in the studies by Hart (1974), Hart (1976), and Martin and Hart (1974) was considered a 

deficiency. 

Endpoint evaluation procedures

Some methodological considerations used to evaluate studies of RDX toxicity are outcome 

specific—in particular, effects on the nervous system and development. Outcome-specific 

methodological considerations are discussed in the relevant health effect sections in Section 1.2. 

For example, many of the studies that noted neurotoxicity in the form of seizures or convulsions 

were not designed to assess that specific endpoint and reported the number of animals with 

seizures as part of clinical observations that, in general, were recorded only once daily. This 

frequency of observations could have missed neurobehavioral events. While these studies can 

provide qualitative evidence of neurotoxicity, they may have underestimated the true incidence of 

seizures or convulsions because they were not designed to systematically evaluate neurotoxic 

outcomes. 

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Outcomes and data reporting

1 In evaluating studies, consideration was given to whether data were reported for all 

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endpoints specified in the methods section and for all study groups, and whether any data were 

excluded from presentation or analysis. For example, it was noted where histopathological analysis 

was limited to control and high-dose groups, a study reporting feature that limited the ability to 

identify dose-related trends. In limited cases, EPA performed additional statistical analysis to 

identify trends or refine analyses consistent with EPA guidance (e.g., analyzing developmental data 

sets on a per litter basis rather than by individual fetus). Study results have been extracted and 

presented in evidence tables.

Notable features of the RDX database

Three 2-year toxicity bioassays of RDX are available as unpublished laboratory studies (Lish 

et al., 1984; Levine et al., 1983b; Hart, 1976). The bioassays by Levine et al. (1983b) in the rat and 

by Lish et al. (1984) in the mouse were conducted in accordance with Food and Drug 

Administration (FDA) Good Laboratory Practices (GLPs) in place at the time of the studies. Both 

studies included interim sacrifices (at 6 and 12 months). Complete histopathological examinations 

were performed on all animals in the control and high-dose groups; however, only a subset of

tissues was examined in the mid-dose groups, limiting the ability to identify dose-related trends for 

tissues with incomplete histopathology. Additionally, in the mouse bioassay by Lish et al. (1984), 

the initial high dose (175 mg/kg-day) was reduced to 100 mg/kg-day at week 11 because of high 

mortality, thereby reducing the number of high-dose animals on study for the full 2 years of dosing 

(see Table LS-5). 

An earlier unpublished 2-year study in rats by Hart (1976) used a dose range that was 

lower than the Levine et al. (1983b) and Lish et al. (1984) bioassays. Histopathology findings were 

limited by the lack of pathology examinations in the mid-dose groups and lack of individual time of 

death, which impacts the ability to interpret the histopathology data. In addition, a heating system 

malfunction on days 75-76 of the study resulted in the death of 59 rats from the control and 

treatment groups, thereby reducing the number of animals in the study (see Table LS-5).

Experimental animal toxicity studies of RDX involving less-than-lifetime exposure (Crouse 

et al., 2006; Angerhofer et al., 1986; MacPhail et al., 1985; Levine et al., 1981a; Cholakis et al., 1980; 

Hart, 1974; Martin and Hart, 1974; von Oettingen et al., 1949) were published or reported between 

the years 1949 and 2006, and differences in robustness of study design, conduct, and reporting 

reflect that time span. All but two of the eight short-term and subchronic toxicity studies of RDX 

are available as unpublished laboratory studies; published studies include von Oettingen et al. 

(1949) and Levine et al. (1981a), a laboratory report of a 13-week study of RDX in F344 rats with 

subsets of the data subsequently published as Levine et al. (1981b) and Levine et al. (1990). The 

majority of studies conducted histopathological examinations on only some of the experimental 

groups (e.g., control and high dose). One subchronic study Crouse et al. (2006) was peer-reviewed 

by Versar, Inc. for EPA in 2012. The peer reviewers determined that the report provided useful 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 information on the toxicity of RDX, including an array of endpoints for neurotoxicity and 

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immunotoxicity (U.S. EPA, 2012d). The assessment of neurotoxicity in the study could have been 

improved with more histological evaluation as well as additional behavioral assessment. 

Some of the more important limitations in study design, conduct, and reporting of 

experimental animal toxicity studies of RDX are summarized in Table LS-5. Limitations of these 

studies as well as the study evaluation consideration described in this section were taken into 

consideration in evaluating and synthesizing the evidence for each of the health effects in 

Section 1.2.

Table LS-5. Experimental animal studies considered less informative because 

of certain study design, conduct, or reporting limitations

References Study design, conduct, and reporting limitations

Lish et al. (1984)

2-yr mouse study

The initial high dose (175 mg/kg-d) was reduced to 100 mg/kg-d at wk 11 

due to high mortality. Mortality of surviving mice was similar to controls 

after dose reduction.

Hart (1976)

2-yr rat study

A heating system malfunction on d 75−76 of the study resulted in the 

deaths of 59 rats from the control and treatment groups. Dead animals 

were subsequently eliminated from the analysis. Interpretation of the 

histopathology findings was limited by the lack of pathology examinations 

in the mid-dose groups and lack of individual time of death. Test material 

poorly characterized; purity was not reported.

Cholakis et al. (1980)

13-wk mouse study (Experiment 1)

The dose range was too low to produce effects in mice. Histopathological 

examinations were not performed.

Cholakis et al. (1980)

13-wk mouse study (Experiment 2)

Nonstandard dosing regimen followed: 0, 40, 60, or 80 mg/kg-d for 2 wks. 

For the next 11 wks, the dosing was inverted, so that the 40 mg/kg-d 

group received 320 mg/kg-d, the 60 mg/kg-d group received 160 mg/kg-d, 

and the 80 mg/kg-d group continued to receive the same dose. The 

rationale for this dosing regimen was not provided in the study report.

Levine et al. (1981a)

13-wk rat study

Analysis of one lot of rodent feed showed measurable levels of 

contaminants, including chlorinated pesticides (dieldrin, heptachlor 

epoxide, beta-hexachlorocyclohexane [BHC], and 

dichlorodiphenyltrichloroethane [DDT]), polychlorinated biphenyls (PCBs), 

and organophosphates (methyl parathion, carbophenothion, and 

disulfeton).

Martin and Hart (1974)

13-wk monkey study

The species of monkey is unclear (either Cynomolgus or Rhesus). Some 

test subjects may have had variable dosing due to emesis. Small sample 

size per dose group (n = 3). Test material poorly characterized; purity was 

not reported.

von Oettingen et al. (1949)

12-wk rat study

The strain of rat was not reported. Only gross observations were made at 

autopsy.

von Oettingen et al. (1949)

6-wk dog study

The breed of dog was not reported. Only gross observations were made 

at autopsy.

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1.HAZARD IDENTIFICATION

1.1. Overview of Chemical Properties and Toxicokinetics

1.1.1. Chemical Properties

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nitrate explosives (Boileau et al., 2003; Bingham et al., 2001) and is not found naturally in the 

environment. It has low solubility in water (Yalkowsky and He, 2003) and slowly volatilizes from 

water or moist soil (ATSDR, 2012). The normalized soil organic carbon/water partition coefficient 

(KOC) values for RDX indicate a potential for RDX to be mobile in soil (Spanggord et al., 1980a). The 

vapor pressure suggests that RDX will exist as particulate matter in air and be removed by both wet 

and dry deposition (Spanggord et al., 1980a). RDX degrades in the environment and can be subject 

to both photolysis (Sikka et al., 1980; Spanggord et al., 1980a) and biodegradation (Funk et al., 

1993; McCormick et al., 1981) (Table 1-1). 

Table 1-1. Chemical identity and physicochemical properties of RDX

Characteristic or property Value Reference

Chemical structure NLM (2011)

CASRN 121-82-4

Synonyms Hexahydro-1,3,5-trinitro-s-triazine; 

1,3,5-trinitro-1,3,5-triazacyclohexane; 

1,3,5-trinitrohexahydro-s-triazine; cyclonite; 

cyclotrimethylenetrinitramine; hexogen; 

cyclotrimethylenenitramine; Research 

Department Explosive; Royal Demolition 

eXplosive; RDX

Color/form White, crystalline solid Bingham et al. (2001)

Molecular formula C3H6N6O6 NLM (2011)

Molecular weight 222.12 Lide (2005)

Density (g/cm3 at 20°C) 1.82 Lide (2005)

Boiling point (°C) 276−280 Bingham et al. (2001)

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Characteristic or property Value Reference

Melting point (°C) 205.5 Lide (2005)

Heat of formation (kJ/g) −0.277 Ryon et al. (1984)

Log KOW 0.87−0.90 Hansch et al. (1995)

KOC 42−167 Spanggord et al. (1980b)

Henry’s law constant 

(atm-m3/mole at 25°C)

2.0 × 10−11 ATSDR (2012)

Vapor pressure (mm Hg at 20 ̊C) 4.10 × 10−9 Spanggord et al. (1980a)

Solubility in water (mg/L at 

25°C)

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1.1.2. Toxicokinetics

RDX is absorbed following exposure by oral and inhalation routes (see Appendix C, 

Section C.1.1). Studies in experimental animals indicate that oral absorption rates can range from 

approximately 50 to 90% (Krishnan et al., 2009; Guo et al., 1985; Schneider et al., 1978, 1977), with 

the rate and extent of absorption dependent on the physical form of RDX (i.e., the increased surface 

area associated with finely powdered RDX allows for increased absorption) and the dosing 

preparation or matrix (Bannon et al., 2009a; Krishnan et al., 2009; Crouse et al., 2008; Bannon, 

2006; Guo et al., 1985; MacPhail et al., 1985; Schneider et al., 1977). Dermal absorption of RDX has 

been demonstrated in in vitro studies using human and pig skin (Reddy et al., 2008; Reifenrath et 

al., 2008). 

RDX is systemically distributed, including to the brain (i.e., RDX can cross the blood:brain 

barrier), heart, kidney, liver, and fat (Musick et al., 2010; Bannon et al., 2006; MacPhail et al., 1985; 

Schneider et al., 1977). In rats, RDX can be transferred from dam to fetus across the placental:blood 

barrier, and has been identified in maternal milk (Hess-Ruth et al., 2007). 

The metabolism of RDX in humans has not been investigated. Studies in experimental 

animals indicate that metabolism of RDX is extensive and includes denitration, ring cleavage, and 

generation of CO2 possibly through cytochrome P450 (CYP450) (Musick et al., 2010; Major et al., 

2007; Fellows et al., 2006; Bhushan et al., 2003; Schneider et al., 1978, 1977). 

RDX and metabolites are eliminated primarily via urinary excretion and exhalation of CO2 

(Sweeney et al., 2012a; Musick et al., 2010; Krishnan et al., 2009; Major et al., 2007; Schneider et al., 

1977). Estimated elimination half-lives (t1/2) indicate that RDX is more rapidly metabolized in mice 

than in rats and humans; estimated t1/2 values were 1.2 hours for mice, 5−10 hours for rats, and 

15−29 hours for humans (Sweeney et al., 2012b; Krishnan et al., 2009; Özhan et al., 2003; Woody et 

al., 1986; Schneider et al., 1977). 

A more detailed summary of RDX toxicokinetics is provided in Appendix C, Section C.1. 

This document is a draft for review purposes only and does not constitute Agency policy.

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1.1.3. Description of Toxicokinetic Models 

1 A physiologically based pharmacokinetic (PBPK) model to simulate the pharmacokinetics of 

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RDX in rats was first developed by Krishnan et al. (2009) and revised to extend the model to 

humans and mice (Sweeney et al., 2012a; Sweeney et al., 2012b). The Sweeney et al. (2012a) model 

consists of six main compartments: blood, brain, fat, liver, and lumped compartments for rapidly 

perfused tissues and slowly perfused tissues, and can simulate RDX exposures via the intravenous 

(i.v.) or oral route. This model assumes that the distribution of RDX to tissues is flow-limited, and 

represents oral absorption as first-order uptake from the gastrointestinal (GI) tract into the liver, 

with 100% of the dose absorbed. RDX is assumed to be cleared by first-order metabolism in the 

liver. The model does not represent the kinetics of any RDX metabolites. The Sweeney et al. 

(2012a) and Sweeney et al. (2012b) PBPK models were evaluated and subsequently modified by

the U.S. Environmental Protection Agency (EPA) for use in dose-response modeling in this 

assessment (see Appendix C, Section C.1.5).

1.2. PRESENTATION AND SYNTHESIS OF EVIDENCE BY ORGAN/SYSTEM

In experimental animal studies, RDX test material administered in toxicology studies 

included formulations that ranged in purity (from 84 to 99.99%) and in particle size (from <66 to 

~200 μm particle size). Differences in test material purity and particle size were taken into 

consideration while evaluating RDX toxicity findings; this is discussed in the literature search

section and incorporated in the synthesis of evidence.

Mortality has been reported in the animal toxicology studies conducted for RDX. Due to the 

serious nature associated with a frank effect such as mortality, EPA specifically evaluated the 

database with respect to mortality (see Appendix C, Section C.3.1). In brief, mortality was observed 

following exposure to a range of doses in chronic-duration studies, in studies up to 6 months in 

duration, and during gestation. In further analyzing the available evidence, mortality occurred at 

lower doses in rats compared with mice and following gavage administration compared with 

dietary administration. Additionally, mortality occurred to a greater extent with administration of 

RDX in the form of relatively finer particle sizes, potentially due to the reduced ability of larger 

particles of RDX to enter the bloodstream. Some investigators attributed the mortality to RDXrelated cancer or noncancer effects (e.g., kidney or nervous system effects); others identified no 

cause for the animal deaths. Typically, evidence related to various hazards is presented and 

synthesized in distinct organ- or system-specific sections. However, in this case, the assessment 

does not present mortality in a hazard section by itself due to the likelihood that events leading to 

mortality fall under other specific hazards. Mortality evidence is considered in discussions of the 

evidence for organ/system-specific hazards where applicable.

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1.2.1. Nervous System Effects

1 In humans, nervous system effects following RDX exposure have been observed in multiple 

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case reports, and the association between RDX exposure and neurobehavioral effects has been 

examined in a single cross-sectional occupational epidemiology study. Information relevant to an 

examination of the association between RDX exposure and nervous system effects also comes from 

experimental animal studies involving chronic, subchronic, and gestational exposure to ingested 

RDX. A summary of nervous system effects associated with RDX exposure is presented in 

Tables 1-2 and 1-3 and Figure 1-1. Experimental animal studies are ordered in the evidence table

and exposure-response array by duration of exposure and then species. 

Observational Studies in Humans

In a cross-sectional study by Ma and Li (1993), neurobehavioral effects were evaluated in 

Chinese workers occupationally exposed to RDX. Memory retention and block design scores5 were 

significantly lower among exposed workers (mean concentrations of RDX in two exposed groups: 

0.407 and 0.672 mg/m3) compared to unexposed workers from the same plant. However, no 

significant differences were observed between the groups on other neurobehavioral tests (e.g., 

simple and choice reaction times, and letter cancellation test) (Table 1-2). This study did not

consider potential confounders such as alcohol consumption or co-exposure to trinitrotoluene 

(TNT), and there was limited information characterizing exposure to RDX. 

Case reports suggest an association between RDX exposure (via ingestion, inhalation, and 

possibly dermal exposure) and neurological effects (see Appendix C, Section C.2). Severe 

neurological disturbances include tonic-clonic seizures (formerly known as grand mal seizures) in 

factory workers (Testud et al., 1996a; Testud et al., 1996b; Kaplan et al., 1965; Barsotti and Crotti, 

1949), seizures and convulsions in exposed soldiers serving in Vietnam (Ketel and Hughes, 1972; 

Knepshield and Stone, 1972; Hollander and Colbach, 1969; Stone et al., 1969; Merrill, 1968), 

seizures, dizziness, headache, and nausea following nonwartime/nonoccupational exposures

(Kasuske et al., 2009; Davies et al., 2007; Küçükardali et al., 2003; Hett and Fichtner, 2002; HarrellBruder and Hutchins, 1995; Goldberg et al., 1992), and seizures in a child following ingestion of 

plasticized RDX from the mother’s clothing (Woody et al., 1986). 

Studies in Experimental Animals

Nervous system effects in experimental animals include a wide array of behavioral changes 

consistent with the induction of seizures by RDX exposure, and have been observed in the majority 

of chronic, subchronic, and developmental studies examining oral exposure to RDX (see Table 1-3 

5The memory quotient index measured short-term hearing memory, visual memory, combined hearing and 

visual memory, and learning ability. The block design index measured visual perception and design 

replication, and the ability to analyze spatial relationships.

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and Figure 1-1). Although study authors interchangeably used the terms seizures and convulsions, 

seizures, which result from abnormal electrical activity in the brain, can outwardly manifest in a 

variety of ways. While seizures can be detected in the form of convulsions, they can also manifest 

as facial twitches, tremors, or increased irritability, or they may go unnoticed. While behavioral 

methods exist to capture a spectrum of responses known to occur as a result of this aberrant 

neuronal activity, the most reliable detection methods are electrophysiological (Racine, 1972). 

Convulsions have been reported in studies with different animal species and experimental 

designs. In every study that reported convulsions, the incidence of convulsions increased with 

dose. In 2-year dietary studies in rats (F344 and Sprague-Dawley) and mice (B6C3F1), convulsions 

were observed beginning at doses of 35−40 mg/kg-day, but not at lower doses (Lish et al., 1984; 

Levine et al., 1983b; Hart, 1976).6 Subchronic dietary exposure to RDX was also associated with 

convulsions in the rat, although doses reported to increase convulsive activity were inconsistent 

across studies. Convulsions were reported in RDX-exposed rats at subchronic doses as low as 8 and 

25 mg/kg-day (Crouse et al., 2006; von Oettingen et al., 1949). In three other rat studies involving 

exposure durations of 30−90 days, no evidence of seizures, convulsions, or tremors was reported at 

doses ranging from 1 to 50 mg/kg-day (MacPhail et al., 1985; Cholakis et al., 1980) (both 

unpublished technical reports). Levine et al. (1990) reported convulsions in rats following 

subchronic exposure only at a dose of 600 mg/kg-day (a dose associated with 100% mortality); 

however, the unpublished technical report of this study (Levine et al., 1981a) inconsistently 

reported convulsions at 600 and ≥30 mg/kg-day, thereby reducing confidence in the identification 

of the dose level at which nervous system effects were observed in this study. RDX exposure (by 

gavage) during gestation in the rat was associated with induction of seizures or convulsions in the 

dams at doses ranging from 2 to 40 mg/kg-day (Angerhofer et al., 1986; Cholakis et al., 1980) 

(unpublished technical reports), demonstrating that effects on the nervous system can be observed 

following exposure durations as short as 10−14 days. Convulsions were also reported in dogs 

exposed to 50 mg/kg-day RDX for 6 weels (von Oettingen et al., 1949), but not 10 mg/kg-day for 

the 13 weeks (Hart, 1974) (unpublished technical report), and in five of six monkeys following a 

gavage dose of 10 mg/kg-day (Martin and Hart, 1974) (unpublished technical report). 

1 In the only study addressing susceptibility to seizures, Burdette et al. (1988) found that 

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seizure occurrence was more frequent in Long Evans rats exposed to a single dose of 50 or 

60 mg/kg RDX by gavage when challenged with an audiogenic stimulus 8 and 16 hours after 

treatment. However, no audiogenic seizures were observed at the earlier 2- and 4-hour post6The 2-year dietary studies in F344 rats by Levine et al. (1983b) and B6C3F1 mice by Lish et al. (1984) were

available only as a laboratory reports. An external peer review was conducted by EPA in July 2012 to

evaluate the accuracy of experimental procedures, results, and interpretation and discussion of the findings 

presented. A report of this peer review performed by Versar, Inc. is available through the EPA’s IRIS Hotline 

at (202) 566-1676 (phone) or hotline.iris@epa.gov (e-mail address), and on the Health and Environmental 

Research Online (HERO) database (U.S. EPA, 2012d). The 2-year dietary study in Sprague-Dawley rats by 

Hart (1976) is available as an unpublished technical report.

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1 dosing test periods even though RDX plasma concentrations were elevated throughout the testing 

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period. In a complementary experiment, Long Evans rats treated daily with 6 mg/kg-day RDX for 

up to 18 days required fewer stimulation trials to exhibit amygdaloid kindled seizures compared to 

controls. Neither the purity nor the specific particle size of the RDX used in the experiments by 

Burdette et al. (1988) was reported. 

The majority of animal studies reported convulsions and/or seizures as clinical 

observations; interpretation of these observations is limited to some extent because the nature and 

severity of convulsions and seizures were not more fully characterized. The 90-day study by

Crouse et al. (2006)7 was one of the few studies that collected and reported incidence data for 

convulsions and tremors, and demonstrated a clear dose-related increase in convulsions and 

tremors in male and female F344 rats associated with RDX exposure via gavage (see Table 1-3). 

Tremors were reported following administration of ≥12 mg/kg-day, persisting throughout the 

90-day study. Convulsions were observed at ≥8 mg/kg-day in male and female rats; information on 

duration and onset was not reported (Crouse et al., 2006). 

In general, gavage dosing induced convulsions at lower doses than did dietary 

administration. For example, in the gavage studies by Crouse et al. (2006) and Cholakis et al. 

(1980), convulsions were observed in 1−3 rats/group at doses of 2−8 mg/kg-day; at doses of 

15−20 mg/kg-day, convulsions were observed in approximately 60−70% of the animals. In 

contrast, in a 2-year dietary study by Levine et al. (1983b), convulsions were reported only at a 

dose of 40 mg/kg-day; no convulsions were observed at lower doses (≤8 mg/kg-day). The 

difference in response between gavage and dietary administration may be due to the bolus dosing 

resulting from gavage administration and the comparatively faster absorption and higher peak 

blood concentrations of RDX. 

Several experimental animal studies documented that unscheduled deaths were frequently 

preceded by convulsions or seizures. In a 2-year study in rats, Levine et al. (1983b) noted that

tremors and/or convulsions were often seen in high-dose animals prior to their death. In a rat 

developmental toxicity study (Cholakis et al., 1980), investigators concluded that early deaths in 

dams were preceded by convulsions based on the observation of convulsions in one rat prior to 

death, and a similar appearance (e.g., dried blood around the mouth and nose) in other dams that 

died during the study. Convulsions preceding death were also observed in pregnant SpragueDawley rats exposed to RDX during gestation (Angerhofer et al., 1986). 

7The 13-week gavage study in F344 rats by Crouse et al. (2006) was available only as a laboratory report. An 

external peer review was conducted by Versar, Inc. in July 2012 to evaluate the accuracy of experimental 

procedures, results, and interpretation and discussion of the findings presented. The U.S. EPA (2012d) report 

of this peer review is available through the EPA’s IRIS Hotline at (202) 566-1676 (phone) or 

hotline.iris@epa.gov (e-mail address), and on the HERO database. 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 The 90-day Crouse et al. (2006) study provides the most detailed information on the 

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relationship between convulsions and mortality (see Appendix C, Table C-10 for additional 

information on evidence of mortality associated with RDX exposure). Convulsions (3/20) and preterm deaths (2/20)8 were observed in male and female rats exposed to 8 mg/kg-day RDX; the 

incidences of both convulsions and pre-term deaths were higher in dose groups with greater 

exposures. Investigators stated that nearly all observed pre-term deaths in rats exposed to the 

three higher doses (10, 12, and 15 mg/kg-day RDX) for 90 days were preceded by neurotoxic signs 

such as rearing behavior, tremors and convulsions; however, pre-term death did not occur in all 

animals that convulsed. Some uncertainty exists in that convulsions were not typically observed 

during a functional observational battery (FOB) test conducted after exposure, likely due to the 

time needed to complete exposures prior to beginning behavioral testing. Of the 100 RDX-treated 

rats in the Crouse et al. (2006) study, convulsions were documented in 34 male and female rats 

across the five dose groups (with convulsions initially observed anywhere from day 7 to 87); based 

on additional information provided as a memorandum by study investigators (Johnson, 2015a), 26 

of these 34 rats (76%) survived to the end of the 90-day study. In general, higher doses of RDX 

were associated with fewer days of exposure before the first convulsion was observed. Of the eight 

rats that exhibited convulsions prior to pre-term death, convulsions were documented anywhere 

from the same day that the animal died to 8 weeks prior to death. Of the 26 rats that seized and 

survived to day 90, the first seizures were observed as early as day 10 and as late as day 87. Thus, 

while an increase in mortality was observed in the Crouse et al. (2006) study at the same dose as 

convulsions, the additional information provided by Johnson (2015a) do not show as clear a 

correspondence between convulsions (and other neurotoxic signs) and mortality. Analysis of these 

data is limited to the extent that convulsions may have occurred at times when animals were not 

observed and are therefore undercounted in the individual animal data; however, Johnson (2015a)

noted that it is unlikely that seizure observations were missed, since seizures generally occurred 

soon after dosing. 

A few studies reported mortality that was not specifically or directly associated with 

neurological effects (see Appendix C, Table C-10) (Angerhofer et al., 1986; Levine et al., 1981a; von 

Oettingen et al., 1949); however, in these studies, animals may not have been monitored for clinical 

observations with sufficient frequency to have observed convulsive activity prior to death. In case 

reports of convulsions and other nervous system effects in workers exposed to RDX during 

manufacture and in individuals exposed acutely as a result of accidental or intentional ingestion, 

there were no reports of mortality subsequent to the convulsions (see Appendix C, Section C.2). 

Additional neurobehavioral effects associated with RDX exposure in rats included increased 

hyperactivity, hyper-reactivity to approach, fighting, and irritability at doses similar to those that 

8At the 8 mg/kg-day dose level, the three rats that convulsed survived to the end of the study; no convulsions 

were observed in the two rats that died before study termination.

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1 induced tremors, convulsions, and seizures (20−100 mg/kg-day) (Levine et al., 1990; Angerhofer et 

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al., 1986; Levine et al., 1983b; Levine et al., 1981a, b; von Oettingen et al., 1949). Hyperactivity and 

nervousness were also reported in male mice that received a subchronic exposure to 320 mg/kgday RDX (Cholakis et al., 1980). No changes in motor activity, flavor aversion, scheduled-controlled 

behavior, or acoustic startle response were observed in a 30-day gavage study in rats (changes in 

acoustic startle response in acute exposures at higher doses [12.5−50 mg/kg] were noted), but 

doses were relatively low (≤10 mg/kg-day) (MacPhail et al., 1985), and no significant changes in 

behavioral or neuromuscular activity were observed in rats following exposure to ≤15 mg/kg-day 

for 90 days (Crouse et al., 2006). Crouse et al. (2006) observed that stained haircoats and increased 

barbering in female F344 rats receiving 15 mg/kg-day may have been caused by the oral dosing 

procedure (gavage) alone.

Changes in absolute and relative brain weight were mixed across studies. Elevated absolute 

brain weights were reported in subchronic assays in B6C3F1 mice and F344 rats (Crouse et al., 

2006; Levine et al., 1990; Levine et al., 1981a, b; Cholakis et al., 1980); however, the changes were 

not consistently observed across studies. Relative brain weights in some studies showed 

correspondingly greater increases compared to absolute brain weight (Crouse et al., 2006; Levine 

et al., 1983b; Cholakis et al., 1980), but these changes were likely a result of changes in body weight 

in the study, and were not a useful measure of effects of RDX on brain weights. In 2-year oral 

studies, a decrease in absolute brain weight of female B6C3F1 mice (3−4% relative to control) was 

reported at doses ≥35 mg/kg-day (Lish et al., 1984), where as an increase in absolute brain weight 

(2% relative to control) was observed in F344 rats at a dose of 40 mg/kg-day (Levine et al., 1983b). 

Less weight is placed on evidence of organ weight changes from chronic (2-year) studies because 

normal physiological changes associated with aging and intercurrent disease may contribute to 

inter-animal variability that could confound organ weight interpretation (Sellers et al., 2007). 

In some studies, seizures appeared soon after dosing, suggesting that seizure induction was 

more strongly correlated with dose level than with duration of exposure. Consistent with this 

observation are the findings of Williams et al. (2011), who demonstrated that RDX is rapidly 

absorbed and crosses the blood:brain barrier following oral administration in rats, and that 

distribution of RDX (8 μg/g wet weight) to the brain correlated with seizure onset. 

While a dose-response relationship was observed consistently within studies, a dose that 

induced convulsions in animals in one study did not necessarily induce convulsions at the same 

dose in another study. This lack of consistency may be attributed, at least in part, to differences in 

the purity or particle size of the test material across studies. Assuming that increased particle size 

(and the corresponding reduction in available surface area compared with smaller particle sizes) 

results in slowed absorption and distribution to the brain, studies that used a larger particle size 

may be expected to produce less neurotoxicity in test animals. The mouse study by Cholakis et al. 

(1980) used a relatively large RDX particle size (200 μm) compared to the rat study by Levine et al. 

(1983b) that used a smaller (<66 μm) particle size. This could contribute to why the Cholakis et al. 

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1 (1980) subchronic dietary study in the mouse (doses up to 320 mg/kg-day RDX) and rat (doses up 

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to 40 mg/kg-day) failed to report seizures or convulsions. Finally, differences in study design may 

have contributed to differences in reported neurological responses in subchronic and chronic 

duration studies. In particular, the protocols for observation for clinical signs (e.g., observations

performed once daily in the morning in Levine et al. (1983b)) may not have been sufficiently 

frequent to accurately measure the incidence of seizures or other nervous system effects. 

Table 1-2. Evidence pertaining to nervous system effects in humans

Reference and study design Results

Ma and Li (1993) (China)

Cross-sectional study, 60 workers from 

the same plant exposed to RDX (30 in 

Group A [26 males; 4 females]; 30 in

Group B [24 males; 6 females]), 

compared to 32 workers with similar 

age, education level, and length of 

employment from same plant with no 

exposure to RDX (27 males; 5 females). 

Exposure measures: Details of 

exposure measurement were not 

provided; two groups of workers

exposed to the following mean RDX 

concentrations in air (basis for dividing 

workers into two exposure groups was 

not provided). 

Concentration (mg/m3

) (mean 

± standard deviation): 

Group A 0.407 (± 0.332)

Group B 0.672 (± 0.556)

Effect measuresa

: Five neurobehavioral 

function tests and five additional 

memory subtests. 

Analysis: Variance (F-test); unadjusted 

linear regression, multiple regression, 

and correlation analysis. 

Neurobehavioral function tests, scaled scores (mean, standard 

deviation)

Test Control Group A Group B

Memory retention* 111.3 (9.3) 96.9 (9.6) 91.1 (10.3)

Simple reaction time

(milliseconds)

493 (199) 539 (183) 578 (280)

Choice reaction time

(milliseconds)

763 (180) 775 (161) 770 (193)

Block design*

(elapsed time)

18.0 (5.4) 16.0 (4.3) 13.5(6.7)

Letter cancellation

(quality per unit time)

1,487 (343) 1,449 (331) 1,484 (443)

*p < 0.01 (overall F-test); no statistically significant differences between 

Group A and Group B. 

Lower score indicates worse performance.

Memory retention subtests, scaled scores (mean, standard deviation)

Subtest Control Group A Group B

Directional memory* 23.5 (3.6) 17.2 (4.9) 18.1 (5.7)

Associative learning* 24.9 (5.1) 20.0 (4.3) 18.5 (4.6)

Image free recall* 24.1 (3.8) 20.9 (4.1) 20.4 (3.3)

Recognition of 

nonsense pictures*

26.3 (3.6) 23.2 (4.9) 21.6 (4.3)

Associative recall of 

portrait characteristics*

26.3 (3.3) 20.3 (4.4) 18.5 (4.3)

*p < 0.01 (overall F-test); no statistically significant differences between 

Group A and Group B. 

Lower score indicates worse performance.

Total behavioral score negatively correlated with exposure index (high 

exposure correlated with poor performance).

a

Symptom data were not included in evidence table because of incomplete reporting.

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Table 1-3. Evidence pertaining to nervous system effects in animals

Reference and study design Results

Convulsions and neurobehavioral effects

Lish et al. (1984)

Mice, B6C3F1, 85/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles <66 μm

0, 1.5, 7.0, 35, or 175/100 mg/kg-d (high 

dose reduced to 100 mg/kg-d in wk 11 

due to excessive mortality)

Diet

2 yrs

One male in the 35 mg/kg-d dose group and one female in the 

175/100 mg/kg-d group convulsed near the end of the study.

Hart (1976)

Rats, Sprague-Dawley, 100/sex/group

Purity and particle size not specified

0, 1.0, 3.1, or 10 mg/kg-d 

Diet

2 yrs

No nervous system effects, as evidenced by clinical signs or changes 

in appearance or behavior, were reported.

Levine et al. (1983b)

Rats, F344, 75/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles <66 μm

0, 0.3, 1.5, 8.0, or 40 mg/kg-d 

Diet

2 yrs

Tremors, convulsions, and hyper-responsiveness to stimuli were 

noted in males and females at 40 mg/kg-d; no incidence data were 

reported.

Cholakis et al. (1980)

Mice, B6C3F1, 10−12/sex/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

0, 40, 60, or 80 mg/kg-d for 2 wks 

followed by 0, 320, 160, or 80 mg/kg-d 

(TWA doses of 0, 79.6, 147.8, or 

256.7 mg/kg-d for males and 0, 82.4, 

136.3, or 276.4 mg/kg-d for females)a

Diet

13 wks

Hyperactivity and/or nervousness observed in 50% of the high-dose 

males; no signs observed in femalesb; no incidence data were 

reported.

Cholakis et al. (1980)

Rats, F344, 10/sex/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

0, 10, 14, 20, 28, or 40 mg/kg-d 

Diet

13 wks

No nervous system effects, as evidenced by clinical signs or changes 

in appearance or behavior, were reported. 

This document is a draft for review purposes only and does not constitute Agency policy.

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Reference and study design Results

Cholakis et al. (1980)

Rats, CD, two-generation study; 

F0: 22/sex/group; F1: 26/sex/group; 

F2: 10/sex/group 

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

F0 and F1 parental animals: 0, 5, 16, or 

50 mg/kg-d 

Diet

F0 exposure: 13 wks pre-mating, and 

during mating, gestation, and lactation of 

F1; F1 exposure: 13 wks after weaning, 

and during mating, gestation, and 

lactation of F2; F2 exposure: until weaning

No nervous system effects were reported.

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d 

Gavage

13 wks

Doses 0 4 8b 10 12 15

Convulsions (incidence)

M 0/10 0/10 1/10 3/10 8/10 7/10

F 0/10 0/10 2/10 3/10 5/10 5/10

Tremors (incidence)

M 0/10 0/10 0/10 0/10 2/10 3/10

F 0/10 0/10 0/10 0/10 0/10 1/10

(Levine et al. (1990); Levine et al. (1981a), 

1981b))

c

Rats, F344, 10/sex/group; 30/sex for 

control

84.7 ± 4.7% purity, ~10% HMX, median 

particle diameter 20 μm, ~90% of particles 

≤66 μm

0, 10, 30, 100, 300, or 600 mg/kg-d 

Diet

13 wks

Hyper-reactivity to approach was observed in rats (sex not specified) 

receiving ≥100 mg/kg-d; no incidence data were reported.

Tremors and convulsions were observed prior to death in one female 

and two male rats receiving 600 mg/kg-d.d (600 mg/kg-d was lethal 

to all rats.)

von Oettingen et al. (1949)

Rats, sex/strain not specified, 20/group

90−97% pure, with 3−10% HMX; particle 

size not specified

0, 15, 25, or 50 mg/kg-d 

Diet

13 wks

Hyperirritability and convulsions were observed in the 25 and 

50 mg/kg-d groupsb; no incidence data were reported.

This document is a draft for review purposes only and does not constitute Agency policy.

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Reference and study design Results

Hart (1974)

Dogs, Beagle, 3/sex/group

Pre-mix with ground dog chow containing

20 mg RDX/g-chow, 60 g dog food; purity 

and particle size not specified

0, 0.1, 1, or 10 mg/kg-d 

Diet

13 wks

No nervous system effects, as evidenced by clinical signs or changes 

in appearance or behavior, were reported.

Martin and Hart (1974)

Monkeys, Cynomolgus or Rhesus e, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d 

Gavage

13 wks

Doses 0 0.1 1 10b

CNS effects characterized as depression, trembling, shaking, jerking, 

or convulsions (incidence)

M 0/3 0/3 0/3 2/3

F 0/3 0/3 0/3 3/3

von Oettingen et al. (1949)

Dogs, breed not specified, 

5 females/group (control); 

7 females/group (exposed)

90−97% pure, with 3−10% HMX; particle 

size not specified

0 or 50 mg/kg-d 

Diet

6 d/wk for 6 wks

Treated dogs exhibited convulsions, excitability, ataxia, and 

hyperactive reflexesb; no incidence data were reported.

MacPhail et al. (1985)

Rats, Sprague-Dawley derived CD, 

8−10 males or females/group 

Purity 84 ± 4.7%; ≤66 μm particle size

0, 1, 3, or 10 mg/kg-d 

Gavage

30 d

No changes in motor activity, flavor aversion, scheduled-controlled 

response, or acoustic startle-response were reported.

Cholakis et al. (1980)

Rats, F344, 24−25 females/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants

0, 0.2, 2.0, or 20 mg/kg-d 

Gavage

GDs 6−19

Doses 0 0.2 2.0 20

Convulsions (incidence)

F 0/24 0/24 1/24 18/25

Angerhofer et al. (1986) (range-finding 

study)

Rats, Sprague-Dawley, 6 pregnant 

females/group

Purity 90%; 10% HMX and 0.3% acetic acid 

occurred as contaminants

0, 10, 20, 40, 80, or 120 mg/kg-d 

Gavage

GDs 6−15

Convulsions preceding death were observed at ≥40 mg/kg-d; 

no incidence data were reported.

This document is a draft for review purposes only and does not constitute Agency policy.

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Reference and study design Results

Angerhofer et al. (1986)

Rats, Sprague-Dawley, 39−51 mated 

females/group

Purity 90%; 10% HMX and 0.3% acetic acid 

occurred as contaminants

0, 2, 6, or 20 mg/kg-d 

Gavage

GDs 6−15

Convulsions and hyperactivityb were observed at 20 mg/kg-d; 

no incidence data were reported.

Burdette et al. (1988)

Rats, Long Evans, 10−21 males/group 

0, 10, 12.5, 20, 25, 50, or 60 mg/kg-d 

Study conducted as two experiments with

the same study design, each with a 

control group

Gavage

8 hrs (single exposure)

After an 8-hr exposure, rats placed in 

observation chamber; 0−64 kHz, 95 dB 

ultrasonic cleaner turned on for 1 min or 

until seizure initiated with uncontrolled 

running (whichever occurred first)

Doses 0 10 12.5 20 25 50 60

Prevalence of audiogenic seizures (%)†

M 0 9 0 29 40 82* 78*

†

Values estimated from graph using Grab It! Software. Statistical 

significance indicated by study authors; p < 0.017. 

Brain weight

Lish et al. (1984)

Mice, B6C3F1, 85/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles <66 μm

0, 1.5, 7.0, 35, or 175/100 mg/kg-d (high 

dose reduced to 100 mg/kg-d in wk 11 

due to excessive mortality)

Diet

2 yrs

Doses 0 1.5 7 35 175/100

Absolute brain weight (percent change compared to control)

M 0% −0.2% 0.61% 0.81% −1%

F 0% −2% −2% −4%* −3%*

Relative brain weight (percent change compared to control)

M 0% 4% 2% 2% 5%

F 0% −4% −1% −3% 18%*

Levine et al. (1983b)

Rats, F344, 75/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles <66 μm

0, 0.3, 1.5, 8.0, or 40 mg/kg-d 

Diet

2 yrs

Doses 0 0.3 1.5 8 40

Absolute brain weight (percent change compared to control)

M 0% 2% −1% 2% 2%

F 0% −0.3% −0.4% 1% 2%*

Relative brain weight (percent change compared to control)

M 0% 0% 8% 2% 22%*

F 0% −1% 3% 4% 20%*

This document is a draft for review purposes only and does not constitute Agency policy.

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Reference and study design Results

Cholakis et al. (1980)

Mice, B6C3F1, 10−12/sex/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

Experiment 1: 0, 10, 14, 20, 28, or 

40 mg/kg-d 

Diet

13 wks

Doses 0 10 14 20 28 40

Absolute brain weight (percent change compared to control)

M 0% − − − 2% 2%

F 0% − − − 4% 2%

Relative brain weight (percent change compared to control)

M 0% − − − 6% 2%

F 0% − − − 0% 3%

Experiment 2: 0, 40, 60, or 80 mg/kg-d for 

2 wks followed by 0, 320, 160, or 

80 mg/kg-d (TWA doses of 0, 79.6, 147.8, 

or 256.7 mg/kg-d for males and 0, 82.4, 

136.3, or 276.4 mg/kg-d for females)a

Diet

13 wks

Doses 0 80 160 320

Absolute brain weight (percent change compared to control)

M 0% 0% 2% 10%

F 0% 0% 4% 2%

Relative brain weight (percent change compared to control)

M 0% −3% 1% 8%

F 0% 0% 3% −4%

Cholakis et al. (1980)

Rats, F344, 10/sex/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

0, 10, 14, 20, 28, or 40 mg/kg-d 

Diet

13 wks

Doses 0 10 14 20 28 40

Absolute brain weight (percent change compared to control)

M 0% − − − 3% 0%

F 0% − − − 0% 0%

Relative brain weight (percent change compared to control)

M 0% − − − 7%* 10%*

F 0% − − − 5% 6%

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d 

Gavage

13 wks

Doses 0 4 8 10 12 15

Absolute brain weight (percent change compared to control)

M 0% −1% −0.3% 2% 5%* 7%*

F 0% −2% 6% 1% 4% 6%

Relative brain weight (percent change compared to control)

M 0% 6% 10% 5% 3% 4%

F 0% −2% −2% −12%* −12%* −15%*

This document is a draft for review purposes only and does not constitute Agency policy.

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Reference and study design Results

(Levine et al. (1990); Levine et al. (1981a), 

1981b))

c

Rats, F344, 10/sex/group; 30/sex for 

control

84.7 ± 4.7% purity, ~10% HMX, median 

particle diameter 20 μm, ~90% of particles 

≤66 μm

0, 10, 30, 100, 300, or 600 mg/kg-d 

Diet

13 wks

Doses 0 10 30 100 300 600

Absolute brain weight (percent change compared to control)

M 0% 1% 0.53% −6% − −

F 0% −1% 1% 2% − −

Relative brain weight (percent change compared to control)

M 0% 4% 7% 14% − −

F 0% 0.3% 2% 5% − −

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*Statistically significant (p < 0.05) based on analysis by study authors.

a

Doses were calculated by the study authors.

b

Mortality was reported in some RDX-treated groups in this study. c

Levine et al. (1981a) is a laboratory report of a 13-week study of RDX in F344 rats; two subsequently published 

papers (Levine et al., 1990; Levine et al., 1981b) present subsets of the data provided in the full laboratory report. d

Discrepancies in the doses at which convulsions occurred were identified in the technical report. The nervous 

system effects reported in this table and in the corresponding exposure-response array are those provided in the 

results section of the technical report (Levine et al., 1981a) and in the published paper (Levine et al., 1990). In 

other sections of the technical report, the authors reported that hyperactivity to approach and convulsions were 

observed in rats receiving ≥30 mg/kg-day (abstract and executive summary), or that mortality was observed in 

rats receiving 100 mg/kg-day and that hyperactivity to approach, tremors, and convulsions were observed in 

animals exposed to lethal doses (discussion).

e

The species of monkey used in this study was inconsistently reported in the study as either Cynomolgus (in the 

methods section) or Rhesus (in the summary).

CNS = central nervous system; GD = gestational day; HMX = octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine; 

TWA = time-weighted average

Note: A dash (“‒”) indicates that the study authors did not measure or report a value for that dose group.

This document is a draft for review purposes only and does not constitute Agency policy.

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Figure 1-1. Exposure response array of nervous system effects following oral exposure. 1 1 

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3 

4 

1Because convulsions and seizures are rare in experimental animals, any occurrence in an RDX-exposed group was considered treatment-related. 

Given the severity of this endpoint, a response in treated groups was determined to be significant (filled circles) in the array where there was any 

occurrence of convulsions and/or seizures reported in the study, whether or not the incidence was statistically significantly elevated over the control. 

This document is a draft for review purposes only and does not constitute Agency policy.

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Mechanistic Evidence

1 Studies that have explored the mode of action (MOA) of RDX on the central nervous system

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(CNS) have focused on the potential impacts on neurotransmission. These studies implicate a MOA 

for RDX-induced seizures and convulsions involving distribution to the brain (across the 

blood:brain barrier) and subsequent effects on neurotransmitters, including gamma-amino butyric 

acid (GABA) and glutamate. There is significant evidence from the scientific literature to suggest 

that RDX neurotoxicity results from interactions of RDX with the GABAA receptor. GABA is a major 

inhibitory neurotransmitter in the brain, and the GABAA receptor has been implicated in 

susceptibility to seizures (Galanopoulou, 2008). 

In research conducted by the U.S. Army Center for Health Promotion and Preventative 

Medicine, Williams et al. (2011) and Bannon et al. (2009a) showed a correlation between blood and 

brain concentrations of RDX in rats that received a single oral dose of RDX (>98−99.5% purity) by 

gavage, which closely correlated with the time of seizure onset. RDX (75 mg/kg) was distributed to 

the brain in direct proportion to levels found in the blood, while time to seizure onset was reduced 

as RDX brain levels increased (Williams et al., 2011). Similarly, oral exposure to RDX (via a gel 

capsule: 3 or 18 mg/kg) resulted in quick absorption followed by transport to the brain and 

subsequent alterations in neurotransmission (Bannon et al., 2009a). 

In receptor binding studies, RDX has only showed affinity for GABAA receptors (Williams et 

al., 2011; Williams and Bannon, 2009). Specifically, RDX showed a significant affinity for the 

picrotoxin convulsant site of the GABA channel. RDX treatment in brain slices from the basolateral 

amygdala inhibited GABAA-mediated inhibitory postsynaptic currents and initiated seizure-like 

neuronal discharges. RDX exposure may reduce the inhibitory effects of GABAergic neurons, 

resulting in enhanced excitability that could lead to seizures (Williams et al., 2011; Williams and 

Bannon, 2009). The limbic system, and the amygdala and hippocampus in particular, are known to 

be critical to the development of seizures in various human conditions (e.g., epilepsy) and animal 

models (Jefferys et al., 2012; Gilbert, 1994). Burdette et al. (1988) hypothesized that the limbic 

system was involved in seizures caused by RDX exposure, given that amygdaloid kindled rats (rats 

subjected to patterns of electrical stimulation to promote the development of seizures) exhibited 

pro-convulsant activity at a dose that was approximately half of the dose necessary for RDX to 

induce spontaneous seizures (rats treated with RDX also required fewer electrical stimulations to 

trigger kindled seizures). Potential limbic system involvement is also suggested given its role in 

integrating emotional and behavioral responses (including aggression) and the anecdotal 

observations of hyperactivity, hyper-responsiveness to approach, and irritability noted across 

several studies of RDX toxicity (Levine et al., 1990; Levine et al., 1983b; Levine et al., 1981a, b; 

Cholakis et al., 1980; von Oettingen et al., 1949). 

RDX binding at the picrotoxin convulsant site of the GABA channel may also inform the 

relationship between exposure to the chemical and the time when a seizure was observed. In 

general across the RDX database, induction of convulsions and seizures appears to be more strongly 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 correlated with dose than duration of exposure. However, Gerkin et al. (2010) demonstrated that 

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young C57/Bl6 mice injected intraperitoneally (i.p.) with picrotoxin to induce seizures had a 

significantly increased frequency of elevated neuronal activity (“Up state”), and firing rates were 

significantly increased in neocortical neurons up to 24 hours after exposure. It is possible that this 

period of elevated neuronal activity could increase the likelihood that a subsequent stimulus could 

trigger a seizure. While the study authors did not look at longer durations post exposure, it is 

possible in a chronic exposure scenario that repeated exposure to RDX binding at the same site as 

picrotoxin, through a general increase in brain tissue with elevated neuronal activity, could 

increase the likelihood of seizures developing over time, or have other longer-term effects on 

normal brain function. 

It is possible to construct a hypothetical MOA for RDX-induced neurotoxicity based on the 

evidence summarized above. Following absorption and transport to the brain:

1) Parent RDX acts as a receptor antagonist (supported by Schneider et al. (1977) and 

Williams et al. (2011)), binding noncompetitively to the picrotoxin convulsant site of the 

GABAA receptor (supported by Williams and Bannon (2009) and Williams et al. (2011)).

2) RDX binding to the GABAA receptor results in decreased conduction of chloride through 

the ion channel. 

3) Reduced chloride conduction results in depolarization of the neuronal membrane, 

thereby reducing spontaneous inhibitory postsynaptic currents (sIPSCs). Williams et al. 

(2011) observed a reduction in the amplitude and frequency of sIPSCs in whole-cell in 

vitro recordings of neurons in brain slices from the rat basolateral amygdala after 

exposure to RDX. In addition, RDX treatment of slices inhibited GABA-induced currents. 

4) Reduction in sIPSCs results in an overall reduction in inhibitory inputs to the nervous 

system. Williams et al. (2011) observed a pattern of seizure-like neuronal discharges in 

vitro from slices of the basolateral amydgala in rats after adding RDX and noted that the 

effects were not reversible after 40 minutes of washout. 

The steps above provide a biologically plausible sequence of mechanistic events that result 

in the generation of seizure-like neuronal activity. Reduction of the inhibitory GABAergic signaling 

is common to many convulsants, as summarized in Kalueff (2007). Some organochlorine 

insecticides, including alpha-endosulfan, dieldrin, and lindane, also exert neurotoxic effects through 

interaction with the GABAA receptor, and can produce a range of hyperexcitability effects (including 

convulsions) in mammals (Vale et al., 2003; Bloomquist, 1992; Suñol et al., 1989). The interaction 

of RDX with the GABAA receptor is directly supported by receptor-binding assays (Williams et al., 

2011). Although these binding assays were performed on rat receptors, it is plausible that the 

results are relevant to human neurotoxicity. Seizures have been observed in many species, 

including humans, rats, mice, dogs, lizards, and birds at varying dosages and durations of exposure

This document is a draft for review purposes only and does not constitute Agency policy.

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1 (Quinn et al., 2013; McFarland et al., 2009; Johnson et al., 2007; Bruchim et al., 2005; Küçükardali et 

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al., 2003; Woody et al., 1986; Lish et al., 1984; Berry et al., 1983; Levine et al., 1983b). A more 

recent meta-analysis of toxicogenomic data across a phylogenetically diverse set of organisms (rat, 

quail, fathead minnow, earthworm, and coral) demonstrated that neurotoxic responses are 

conserved in more highly-related species and that binding to the GABAA receptor is a common 

molecular initiating event (Garcia-Reyero et al., 2011). While these lines of evidence do not 

preclude a role of other receptors as yet unscreened for RDX binding affinity, they support 

involvement of the GABAergic pathway described above in the development of RDX neurotoxicity. 

The GABAA receptor is also a target of many anticonvulsant therapies (e.g., benzodiazepines, 

propofol, barbiturates) (Meldrum and Rogawski, 2007; Möhler, 2006). Additional support for the 

involvement of GABAergic signaling in the neurotoxicity of RDX comes from human case reports. In 

multiple case reports, medical intervention included treatment with benzodiazepines (commonly 

diazepam or lorazepam) to treat seizing patients (Kasuske et al., 2009; Davies et al., 2007; 

Küçükardali et al., 2003; Hett and Fichtner, 2002; Woody et al., 1986). Benzodiazepines act in large 

part by enhancing the effects of GABA at the GABAA receptor by increasing chloride conductance, 

resulting in anticonvulsant and relaxant effects (Goodman et al., 1996). 

Some other pro-convulsant agents with minimal direct toxicity to nerve cells, such as sarin 

and some organophosphate pesticides, are known to act through inhibition of acetylcholinesterase 

(AChE) activity (McDonough and Shih, 1997). Some of the clinical signs observed following RDX 

exposure are similar to the clinical signs associated with organophosphate pesticides and nerve 

agents (Crouse et al., 2006; Burdette et al., 1988; Barsotti and Crotti, 1949). However, the limited 

data available for RDX do not support AChE inhibition as a primary mechanism because: (1) blood 

and brain levels of AChE are unaffected by RDX (Williams et al., 2011; Williams and Bannon, 2009); 

and (2) in vitro neurotransmitter receptor binding studies do not reveal any affinity of RDX for 

acetylcholine receptors (Williams et al., 2011; Williams and Bannon, 2009). Additionally, common 

AChE-induced symptoms (salivation and lacrimation) have not routinely been observed (Williams 

et al., 2011). RDX showed no affinity for other receptors that are known targets of convulsants, 

including the glutamate family of receptors, nicotinic receptors, glycine receptors, and several 

monoamine receptors (Williams et al., 2011; Williams and Bannon, 2009). 

In a microarray experiment, Bannon et al. (2009a) found that RDX caused a down 

regulation of an abundance of genes in the cerebral cortex related to neurotransmission, including 

those encoding proteins involved in synaptic transmission and vesicle transport. Genes encoding 

proteins involved in the glutamate pathway were also underexpressed, indicating a possible 

mechanism of action for RDX via excessive glutamate stimulation. The authors speculated that this 

depression of the major excitatory neurotransmitter system could be a negative response to the 

increase in seizure likelihood from RDX influx into the brain. Molecular changes in response to RDX 

have been described by Zhang and Pan (2009b), who observed significant changes in micro-RNA 

(miRNA) expression in the brains of B6C3F1 mice fed 5 mg RDX/kg diet (estimated dose: 0.75−1.5 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 mg/kg-day; Bannon et al., 2009b) for 28 days. One miRNA, miR-206, was upregulated 26-fold in 

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RDX-exposed brains; brain-derived neurotrophic factor (BDNF) was identified as a downstream 

gene target of this miRNA, along with two other miRNAs that were upregulated in RDX-exposed 

brains (miR-30a and miR-195) (Zhang and Pan, 2009a, b). BDNF is a member of the neurotrophin 

family of growth factors, and promotes the survival and differentiation of existing and new neurons. 

Deng et al. (2014) conducted miRNA and mRNA profiling in rats to identify targets up or 

downregulated after 48-hour exposure to RDX, finding that many of the gene targets of these 

miRNAs were associated with nervous system function, and may contribute to the neurotoxicity of 

RDX. However, while effects of RDX on BDNF expression or other downstream targets may play a 

role in RDX neurotoxicity, the utility of miRNAs as predictors of toxicity has not been demonstrated 

and downstream targets of miRNA require verification (Bannon et al., 2009b). Thus, the 

contribution, if any, of aberrant expression of a suite of miRNAs to the MOA for RDX neurotoxicity is 

unknown.

Recent research has provided greater insight to inform a mechanistic basis of RDX 

neurotoxicity. While other possible MOA(s) may contribute to the overall neurotoxicity of RDX, the 

demonstrated affinity of RDX for the GABAA receptor, evidence of supportive electrophysiological 

changes with direct application of RDX, and toxicokinetic evidence of distribution of RDX to the 

brain provides a mechanistic basis for the association of seizures with exposure to RDX. The 

available information supports that RDX-induced hyperactivity and seizures likely result from 

inhibition of GABAergic signaling in the limbic system.

Integration of Nervous System Effects

Evidence for nervous system effects associated with exposure to RDX comes from studies in 

both humans and animals. One occupational study reported memory impairment and decrements 

in certain neurobehavioral tests in workers exposed to RDX compared to controls (Ma and Li, 

1993), and human case reports provide other evidence of an association between acute RDX 

exposure and neurological effects. There was consistent evidence of neurotoxicity associated with 

exposure to RDX; 11 of 16 repeat-dose animal studies (of varying design) reported neurological 

effects (some severe), including seizures, convulsions, tremors, hyperirritability, hyper-reactivity, 

and behavioral changes, associated with RDX exposure (Crouse et al., 2006; Angerhofer et al., 1986; 

Levine et al., 1983b; Levine et al., 1981b; Cholakis et al., 1980; von Oettingen et al., 1949). In most 

of these studies, the occurrence of neurological effects was dose-related. In those studies that 

found no evidence of RDX-associated neurotoxicity (MacPhail et al., 1985; Cholakis et al., 1980; 

Hart, 1976, 1974), differences in dosing, particle size, and purity of the RDX administered could 

possibly account for the lack of effect. Seizures resulting from RDX exposure likely result from 

inhibition of GABAergic signaling due to the interaction of RDX with the GABAA receptor. 

Convulsant receptor binding leading to a decreased seizure threshold, considered with kindling 

studies, suggests that the effect is specific to CNS toxicity. 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 Together, toxicological information in animals and humans, supported by toxicokinetic and 

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mechanistic information, provides a coherent identification of nervous system effects as a human

hazard of RDX exposure. 

1.2.2. Kidney and Other Urogenital System Effects

The association between RDX exposure and effects on clinical measures of kidney function

was examined in one occupational epidemiology study. Case reports, involving accidental exposure 

to ingested or inhaled RDX, offer some information on the potential for acute exposure to RDX to 

affect the kidney in humans. Organ weight and histopathology findings from experimental animal 

studies involving subchronic and chronic exposure to ingested RDX also provide data relevant to an 

examination of the association between RDX exposure and kidney and other urogenital system 

effects. A summary of these effects associated with RDX exposure is presented in Tables 1-4 to 1-8 

and Figure 1-2. Experimental animal studies are ordered in the evidence table and exposureresponse array by duration of exposure and then by species.

Human case reports of individuals accidently exposed to unknown amounts of RDX by 

ingestion or inhalation provide some evidence that RDX may affect the kidney and the urogenital 

system. Reported symptoms included decreased urine output (Ketel and Hughes, 1972; Knepshield 

and Stone, 1972; Hollander and Colbach, 1969; Merrill, 1968), blood in urine (Kasuske et al., 2009; 

Knepshield and Stone, 1972; Hollander and Colbach, 1969; Merrill, 1968), proteinuria (Kasuske et 

al., 2009; Küçükardali et al., 2003; Ketel and Hughes, 1972; Hollander and Colbach, 1969; Merrill, 

1968), glucosuria (Küçükardali et al., 2003), elevated blood urea nitrogen (BUN) levels (Hollander 

and Colbach, 1969; Merrill, 1968), and one case of acute renal failure requiring hemodialysis 

following accidental inhalation of RDX (Ketel and Hughes, 1972). In many of these case reports, 

renal parameters returned to normal within a few days following exposure. No changes in renal 

parameters were reported in other individuals exposed to unknown amounts of RDX (Stone et al., 

1969; Kaplan et al., 1965). In a cross-sectional epidemiologic study of workers from five U.S. Army 

munitions plants (69 exposed to RDX alone and 24 exposed to RDX and octahydro1,3,5,7-tetranitro-1,3,5,7-tetrazocine (HMX); RDX exposure range: undetectable [<0.01 mg/m3] to 

1.6 mg/m3), no statistically significant differences in BUN or total serum protein between 

nonexposed and RDX-exposed groups were observed (Hathaway and Buck, 1977) (Table 1-4). As it 

is a cross-sectional study, no information was provided on the length of employment or other 

proxies that could be used to indicate cumulative exposure concentrations. 

Studies in experimental animals provide some evidence that RDX exposure is associated 

with kidney and other urogenital system effects (Table 1-5 and Figure 1-2). Histopathological 

changes in the urogenital system were associated with exposure to RDX in a 2-year bioassay. 

Specifically, increased incidences of kidney medullary papillary necrosis and pyelitis, uremic 

mineralization, bladder distention and/or cystitis, and suppurative prostatitis were observed in 

high-dose (40 mg/kg-day) male rats that died spontaneously or were sacrificed in moribund 

condition (Levine et al., 1983b). These renal effects were considered the principal cause of 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 treatment-related morbidity and mortality in these high-dose males. Similar kidney lesions were 

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not observed in female rats in this study. An increased incidence of tubular nephrosis was 

observed in male B6C3F1 mice exposed to 320 mg/kg-day RDX in feed for 90 days, but not in female 

mice in this study (Cholakis et al., 1980). In other chronic and subchronic oral studies in rats and 

mice, no histopathological changes in the kidney were associated with RDX exposure (Crouse et al., 

2006; Levine et al., 1990; Lish et al., 1984; Levine et al., 1981a, b; Cholakis et al., 1980; Hart, 1976). 

Increased incidence of minimal to mild mineralization of the medulla was observed in male and 

female monkeys exposed to 10 mg/kg-day RDX for 90 days by gavage (Martin and Hart, 1974), but 

the study authors did not identify this as treatment related. No dose-related histopathological 

changes were reported in a subchronic study in dogs (Hart, 1974), and no histological alterations 

were noted in the kidneys of rabbits exposed dermally to a cumulative dose of 165 mg/kg RDX in 

dimethylsulfoxide (DMSO) received over a 4-week period (5 days/week) (McNamara et al., 1974). 

Measurement of serum chemistry parameters that may indicate effects on renal function, including 

BUN and uric acid, in studies of RDX in mice, rats, dogs, and monkeys (Crouse et al., 2008; Levine et 

al., 1990; Lish et al., 1984; Levine et al., 1981a, b; Cholakis et al., 1980; Hart, 1976, 1974; Martin and 

Hart, 1974) revealed variations (increases or decreases) from the respective control groups that 

were not dose-related. 

The findings of suppurative prostatitis provide the strongest evidence of urogenital toxicity. 

A significant, dose-related increase in the total incidence of suppurative prostatitis was reported in 

male F344 rats exposed to ≥1.5 mg/kg-day RDX in the diet for 2 years (Levine et al., 1983b). 

Suppurative prostatitis was not observed in 90-day studies in the rat involving oral (dietary or 

gavage) exposure to RDX (Crouse et al., 2006; Levine et al., 1990; Levine et al., 1981a, b). Similarly, 

prostate effects were not observed in a 2-year dietary study in mice (Lish et al., 1984). The Levine 

et al. (1983b) report is the only 2-year study that reported examination of the prostate in rats. 

Some reports have hypothesized that the observation of prostate inflammation in Levine et al. 

(1983b) is secondary to a bacterial infection unrelated to RDX toxicity (ATSDR, 2012; Sweeney et 

al., 2012a; Crouse et al., 2006). For example, in describing the results from the 2-year dietary study 

in rats, Crouse et al. (2006) observed that the inflammation reflects a common condition in rodents, 

noting that since 85% of the incidence occurred in rats found at spontaneous death or moribund 

sacrifice (SDMS), it was most likely that the condition was a result of an incidental bacterial 

infection. However, Levine et al. (1983b) distinguished between nonsuppurative and suppurative 

inflammation (the latter being characterized by the formation of pus and a high concentration of 

neutrophils). Although the proportion of suppurative prostatitis was higher in SDMS rats, there 

was an increasing trend with dose in both the scheduled sacrifice (SS) and SDMS groups; the 

incidence of suppurative prostatitis in the control group was 4% when the SS and SDMS groups 

were combined. Additionally, the dose-related nature of the increased incidence suggests that the 

primary cause (potentially leading to bacterial infection) was treatment-related, as a more uniform 

distribution of rats with suppurative prostatitis would be expected with a spontaneous or ageThis document is a draft for review purposes only and does not constitute Agency policy.

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1 related lesion. The dose-responsiveness could be explained if the infections were secondary to 

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treatment-related immunotoxicity, but there is no evidence from Levine et al. (1983b) to support 

this possibility. A more thorough analysis of immune endpoints in a 90-day gavage exposure of 

F344 rats did not identify any immunotoxic effects associated with RDX (Crouse et al., 2006). 

As noted above, Levine et al. (1983b) documented an array of kidney and other urogenital 

lesions in their 2-year dietary exposure of F344 rats to RDX. However, the sequence by which those 

effects may have occurred is unclear. Renal medullary necrosis, bladder distension, and cystitis 

were observed mainly in the male rats exposed to 40 mg/kg-day RDX for 24 months, although one 

rat in the 0.3 mg/kg-day dose group also exhibited these lesions. Treatment-related effects on the 

kidney (necrosis) and bladder (distension/obstruction and hemorrhagic cystitis) were also 

identified in the 12-month pathology report (see Tables 1-6 to 1-8). The absence of these 

observations in the 6-month interim pathology report suggests that an exposure duration 

>6 months may be required before RDX-induced effects on the urogenital system are observed. 

Suppurative prostatitis was observed with increasing incidence in each dose group in the study at 

24 months. Considered as a group, treatment-related kidney and urogenital lesions may have led to 

a blockage that resulted in urinary stasis. Reduced urinary flow and/or retrograde flow may have 

contributed to an environment that allowed bacterial infection of the prostate. Thus, while an 

opportunistic bacterial infection could be the proximal cause of the suppurative prostatitis (ATSDR, 

2012; Sweeney et al., 2012a; Crouse et al., 2006), it may have been secondary to the effects of RDX 

on the urogenital system. This hypothesis is consistent with the observed dose-related increase in 

incidence of suppurative prostatitis.

Although the ultimate sequence of effects in the urogenital system is unclear, even from 

review of the scheduled sacrifices at 6 or 12 months on study, it is plausible that suppurative 

prostatitis would occur after other kidney or bladder lesions that resulted in the initial blockage 

and urinary stasis. The incidence of suppurative prostatitis reported in Levine et al. (1983b) was 

increased at doses lower than the doses associated with an increased incidence of other urogenital 

lesions. However, the incidence of bladder lesions may have been underreported, as the bladders 

were only examined following observation of a gross abnormality. Bladder distension was 

reported sporadically among the lower dose groups (0.3, 1.5, or 8.0 mg/kg-day), but the bladder 

was not routinely examined in these groups (Levine et al., 1983b). Although the pathogenesis of 

kidney and urogenital system effects cannot be established, the available evidence is consistent 

with suppurative prostatitis as an indirect effect of RDX exposure and as a marker for the broader 

array of kidney and urogenital system effects observed by Levine et al. (1983b). 

Changes in kidney weights in subchronic oral toxicity studies in rats, dogs, and monkeys did 

not show a clear pattern of increase or decrease associated with RDX exposure. Kidney weight 

changes were either not dose-related or were inconsistent across sexes when absolute and relative 

weights were compared (see Table 1-5). Less weight is placed on evidence of organ weight changes 

from chronic (2-year) studies (Lish et al., 1984; Hart, 1976) because normal physiological changes 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 associated with aging and intercurrent disease may contribute to inter-animal variability that could 

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confound organ weight interpretation (Sellers et al., 2007). 

Exposure to HMX, the major contaminant in many of the available RDX studies, was 

associated with histopathological changes in the kidney and alterations in renal function in female, 

but not male, rats fed doses ≥450 mg/kg-day HMX for 13 weeks (see the Integrated Risk 

Information System [IRIS] assessment of octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine [HMX] at 

http://www.epa.gov/iris). No effects were observed at doses ≤115 mg/kg-day. Because the 

percentage of HMX as an impurity ranged from 3 to 10%, resulting in HMX exposures 

≤60 mg/kg-day in the studies of RDX toxicity, the contribution of HMX to the observed kidney 

toxicity in studies of RDX is expected to be negligible. Further, differences in the pattern of toxicity 

(i.e., kidney effects observed only in RDX-exposed males and HMX-exposed females) also suggest 

that HMX contaminants were not responsible for kidney effects in rats exposed to RDX.

Table 1-4. Evidence pertaining to kidney effects in humans

Reference and study design Results

Hathaway and Buck (1977)

Cross-sectional study, 2,022 workers, 

1,491 participated (74% response rate). 

Analysis group: limited to whites; 

69 workers exposed to RDX alone and 

24 workers exposed to RDX and HMX, 

compared to 338 workers not exposed to 

RDX, HMX, or TNT.

Exposure measures: Exposure 

determination based on job title and 

industrial hygiene evaluation; exposed 

subjects assigned to two groups: 

undetected (<LOD) or ≥0.01 mg/m3

(mean for employees with exposures 

≥LOD: 0.28 mg/m3

).

Effect measures: Renal function tests 

(blood)

Analysis: Types of statistical tests were 

not reported (assumed to be t-tests for 

comparison of means and χ2 tests for 

comparison of proportions). 

Renal function tests: mean (standard deviation not reported)

Test

Referent

(n = 237)

RDX exposed males*

Undetected (<LOD)

(n = 22)

>0.01 mg/m3

(n = 45)

BUN 15.5 15.6 16.4

Total protein 7.2 7.2 7.3

RDX exposed females*

Referent

(n = 101)

Undetected (<LOD)

(n = 1)

>0.01 mg/m3

(n = 25)

BUN

Total protein

13.2

7.3

8

7.6

12.6

7.2

*Includes both workers exposed to RDX alone and RDX and HMX.

No differences were statistically significant in men or women. 

LOD = limit of detection

This document is a draft for review purposes only and does not constitute Agency policy.

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1 Table 1-5. Evidence pertaining to kidney and other urogenital system effects 

2 in animals 

Reference and study design Results

Histopathological lesions

Lish et al. (1984)

Mice, B6C3F1, 85/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles <66 μm

0, 1.5, 7.0, 35, or 175/100 mg/kg-d (high 

dose reduced to 100 mg/kg-d in wk 11 

due to excessive mortality)

Diet

2 yrs

The incidence of cytoplasmic vacuolization of renal tubules was 

greater for RDX-treated males than the control group males after 

6 mo of treatment. However, at 12 and 24 mo of treatment, this 

lesion was observed as frequently in controls as males treated with 

RDX. There was no increase in incidence of this lesion in females at 

any time point.

Hart (1976)

Rats, Sprague-Dawley, 100/sex/group

Purity and particle size not specified

0, 1.0, 3.1, or 10 mg/kg-d 

Diet

2 yrs

Histopathological examination of kidney did not reveal any significant 

differences compared to controls; lesions observed were not 

attributed to RDX treatment; incidence data were reported only for 

control and 10 mg/kg-d groups.

Levine et al. (1983b)

Rats, F344, 75/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles <66 μm

0, 0.3, 1.5, 8.0, or 40 mg/kg-d 

Diet

2 yrs

Note: More detailed histopathological 

results, including interim sacrifice data at 

6 and 12 mo, are provided in Tables 1-6 to 

1-8. 

Data for male rats sacrificed on schedule (SS) and those that died 

spontaneously or were sacrificed moribund (SDMS) (summarized 

below) were analyzed separately; incidence data were not reported 

for females.

Doses 0 0.3 1.5 8.0 40

Kidney, medullary papillary necrosis; 24 mo (incidence)

(SS) 0/38 0/36 0/25 0/29 0/4

(SDMS) 0/17 1/19 0/27 0/26 18/27*

(Sum) 0/55 1/55 0/52 0/55 18/31*

Kidney, suppurative pyelitis; 24 mo (incidence)

(SS) 0/38 0/36 0/25 0/29 0/4

(SDMS) 0/17 1/19 0/27 1/26 5/27*

(Sum) 0/55 1/55 0/52 1/55 5/31*

Kidney, uremic mineralization; 24 mo (incidence)

(SS) 1/38 0/36 0/25 0/29 0/4

(SDMS) 0/17 1/19 2/27 0/26 13/27

(Sum) 1/55 1/55 2/52 0/55 13/31

This document is a draft for review purposes only and does not constitute Agency policy.

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Reference and study design Results

Urinary bladder, luminal distention; 24 mo (incidence)

(SS) 0/38 0/36 0/25 0/29 1/4*

(SDMS) 0/16 2/19 1/27 3/22 24/28*

(Sum) 0/54 2/55 1/52 3/51 25/32*

Urinary bladder, cystitis hemorrhagic/suppurative; 24 mo 

(incidence)

(SS) 0/38 0/36 0/25 1/29 0/4

(SDMS) 0/16 2/19 1/27 0/22 18/27*

(Sum) 0/54 2/55 1/52 1/51 18/31*

Prostate, suppurative inflammation (prostatitis); 24 mo (incidence)

SS 0/38 1/36 2/25* 4/29* 0/4

SDMS 2/16 3/19 7/27* 8/26 19/27*

(Sum) 2/54 4/55 9/52* 12/55* 19/31*

Cholakis et al. (1980)

Mice, B6C3F1, 10−12/sex/group 

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

0, 80, 60, 40 mg/kg-d for 2 wks followed 

by 0, 80, 160, or 320 mg/kg-d (TWA doses 

of 0, 79.6, 147.8, or 256.7 mg/kg-d for 

males and 0, 82.4, 136.3, or 

276.4 mg/kg-d for females)a

Diet

13 wks

Doses 0 80 160 320

Tubular nephrosis (incidence)

M 0/10 − − 4/9*

F 0/11 − − 1/11

Cholakis et al. (1980)

Rats, F344, 10/sex/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

0, 10, 14, 20, 28, or 40 mg/kg-d

Diet

13 wks

Histopathological examination of kidney did not reveal any significant 

differences compared to controls; incidence data were reported only 

for control and 40 mg/kg-d groups. 

This document is a draft for review purposes only and does not constitute Agency policy.

1-26 DRAFT—DO NOT CITE OR QUOTE
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Reference and study design Results

Cholakis et al. (1980)

Rats, CD, two-generation study; 

F0: 22/sex/group; F1: 26/sex/group; 

F2: 10/sex/group 

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

F0 and F1 parental animals: 0, 5, 16, or 

50 mg/kg-d

Diet

F0 exposure: 13 wks pre-mating, and 

during mating, gestation, and lactation of 

F1; F1 exposure: 13 wks after weaning, 

and during mating, gestation, and 

lactation of F2; F2 exposure: until weaning

Data were reported only for F2 generation controls and 5 and 

16 mg/kg-d groups.

Doses 0 5 16 50

Cortical cysts (incidence)

M 4/10 4/10 8/10 −

F 3/10 4/10 8/10 −

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d

Gavage

13 wks

Doses 0 4 8 10 12 15

Prostate, mild subacute inflammation (incidence)

M 0/10 − − − − 1/8

Histopathological examination of kidney did not reveal any significant 

differences compared to controls; incidence data were reported only 

for control and 15 mg/kg-d groups. 

(Levine et al. (1990); Levine et al. (1981a), 

1981b))

b

Rats, F344, 10/sex/group; 30/sex for 

control

84.7 ± 4.7% purity, ~10% HMX, median 

particle diameter 20 μm, ~90% of particles 

≤66 μm

0, 10, 30, 100, 300, or 600 mg/kg-d

Diet

13 wks

Histopathological examination of kidney did not reveal any significant 

differences compared to controls. No histopathology findings

available for the 300 or 600 mg/kg-d dose groups because all rats in 

these groups died before the 13-wk necropsy. 

Hart (1974)

Dogs, Beagle, 3/sex/group

Pre-mix with ground dog chow containing 

20 mg RDX/g-chow, 60 g dog food; purity 

and particle size not specified

0, 0.1, 1, or 10 mg/kg-d

Diet

13 wks

Histopathological examination of kidney did not reveal any significant 

differences compared to controls; incidences were reported only for 

control and 10 mg/kg-d groups.

Martin and Hart (1974)

Monkeys, Cynomolgus or Rhesusc, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d

Gavage

13 wks

Doses 0 0.1 1 10

Medulla; mineralization, minimal to mild (incidence)

M + F 0/6 1/6 0/6 4/6

This document is a draft for review purposes only and does not constitute Agency policy.

1-27 DRAFT—DO NOT CITE OR QUOTE
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Reference and study design Results

Kidney weightd

Lish et al. (1984)

Mice, B6C3F1, 85/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles <66 μm

0, 1.5, 7.0, 35, or 175/100 mg/kg-d (high 

dose reduced to 100 mg/kg-d in wk 11 

due to excessive mortality)

Diet

2 yrs

Doses 0 1.5 7.0 35 175/100

Absolute kidney weight at 104 wks (percent change compared to 

control)

M 0% −1% 4% 9%* 19%*

F 0% 3% 1% 1% −2%

Relative kidney weight at 104 wks (percent change compared to 

control)

M 0% 3% 6% 11%* 27%*

F 0% 1% 1% 2% 19%*

Hart (1976)

Rats, Sprague-Dawley, 100/sex/group

Purity and particle size not specified

0, 1.0, 3.1, or 10 mg/kg-d

Diet

2 yrs

Doses 0 1.0 3.1 10

Absolute kidney weight (percent change compared to control)

M 0% −3% −7% 2%

F 0% 14% −4% 8%

Relative kidney weight (percent change compared to control)

M 0% −1% −4% 4%

F 0% 22% 3% 18%

Levine et al. (1983b)

Rats, F344, 75/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles <66 μm

0, 0.3, 1.5, 8.0, or 40 mg/kg-d

Diet

2 yrs

Doses 0 0.3 1.5 8.0 40

Absolute kidney weight at 105 wks (percent change compared to 

control)

M 0% 2% −7% 1% 0%

F 0% 3% 3% 2% 2%

Relative kidney weight at 105 wks (percent change compared to 

control)

M 0% 1% 0% 2% 20%*

F 0% 3% 6% 5% 21%*

Cholakis et al. (1980)

Mice, B6C3F1, 10−12/sex/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

Experiment 1: 0, 10, 14, 20, 28, or 

40 mg/kg-d

Diet

13 wks

Doses 0 10 14 20 28 40

Absolute kidney weight (percent change compared to control)

M 0% − − − 18% 2%

F 0% − − − −8% −5%

Relative kidney weight (percent change compared to control)

M 0% − − − 29% 0%

F 0% − − − −8% −3%

This document is a draft for review purposes only and does not constitute Agency policy.

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Reference and study design Results

Experiment 2: 0, 40, 60, or 80 mg/kg-d for 

2 wks followed by 0, 320, 160, or 

80 mg/kg-d (TWA doses of 0, 79.6, 147.8, 

or 256.7 mg/kg-d for males and 0, 82.4, 

136.3, or 276.4 mg/kg-d for females)a

Diet

13 wks

Doses 0 80 160 320

Absolute kidney weight (percent change compared to control)

M 0% 8% 11% 13%

F 0% −5% −3% 0%

Relative kidney weight (percent change compared to control)

M 0% 5% 9% 10%

F 0% −5% −4% −5%

Cholakis et al. (1980)

Rats, F344, 10/sex/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

0, 10, 14, 20, 28, or 40 mg/kg-d

Diet

13 wks

Doses 0 10 14 20 28 40

Absolute kidney weight (percent change compared to control)

M 0% − − − −2% −5%

F 0% − − − 1% 0%

Relative kidney weight (percent change compared to control)

M 0% − − − 1% 5%

F 0% − − − 6% 6%

Cholakis et al. (1980)

Rats, CD, two-generation study; 

F0: 22/sex/group; F1: 26/sex/group; 

F2: 10/sex/group 

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

F0 and F1 parental animals: 0, 5, 16, or 

50 mg/kg-d

Diet

F0 exposure: 13 wks pre-mating, and 

during mating, gestation, and lactation of 

F1; F1 exposure: 13 wks after weaning, 

and during mating, gestation, and 

lactation of F2; F2 exposure: until weaning 

Doses 0 5 16 50

Absolute kidney weight (percent change compared to control)

M 0% 6% −12% −

F 0% −4% −21%* −

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d

Gavage

13 wks

Doses 0 4 8 10 12 15

Absolute kidney weight (percent change compared to control)

M 0% −3% −4% −1% 3% 5%

F 0% 2% 5% 13%* 10% 15%*

Relative kidney weight (percent change compared to control)

M 0% 3% 6% 2% 1% 3%

F 0% 1% −3% −1% −6% −7%*

This document is a draft for review purposes only and does not constitute Agency policy.

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Reference and study design Results

(Levine et al. (1990); Levine et al. (1981a), 

1981b))

b

Rats, F344, 10/sex/group; 30/sex for 

control

84.7 ± 4.7% purity, ~10% HMX, median 

particle diameter 20 μm, ~90% of particles 

≤66 μm

0, 10, 30, 100, 300, or 600 mg/kg-d 

Diet

13 wks

Doses 0 10 30 100 300 600

Absolute kidney weight (percent change compared to control)

M 0% 1% 1% −9% − −

F 0% 1% 3% −1% − −

Relative kidney weight (percent change compared to control)

M 0% 5% 7% 10% − −

F 0% 3% 5% 2% − −

Hart (1974)e

Dogs, Beagle, 3/sex/group

Pre-mix with ground dog chow containing 

20 mg RDX/g-chow, 60 g dog food; purity 

and particle size not specified

0, 0.1, 1, or 10 mg/kg-d 

Diet

13 wks

Doses 0 0.1 1 10

Absolute kidney weight (percent change compared to control)

M 0% − − 38%

F 0% − − −18%

Martin and Hart (1974)e

Monkeys, Cynomolgus or Rhesuse, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d 

Gavage

13 wks

Doses 0 0.1 1 10

Absolute kidney weight (percent change compared to control)

M + F 0% −2% −3% 4%

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*Statistically significant (p < 0.05) based on analysis by study authors. a

Doses were calculated by the study authors.

b

Levine et al. (1981a) is a laboratory report of a 13-week study of RDX in F344 rats; two subsequently published 

papers (Levine et al., 1990; Levine et al., 1981b) present subsets of the data provided in the full laboratory report. c

The species of monkey used in this study was inconsistently reported in the study as either Cynomolgus (in the 

methods section) or Rhesus (in the summary). d

An analysis by Craig et al. (2014) found a statistically significant correlation between absolute, but not relative, 

kidney weights and renal histopathology. Therefore, only absolute kidney weight data from RDX studies are 

presented in Figure 1-2. e

Kidney weight data from the Hart (1974) and Martin and Hart (1974) studies were considered less informative 

than other studies. Hart (1974) reported organ weight data for high-dose dogs (3/sex/group) only, and the kidney 

weights from Martin and Hart (1974) were highly variable across monkeys (e.g., kidney weights for the control 

animals ranged from 4.9 to 13.1 g). Therefore, kidney weight data from these two studies were not presented in 

the exposure-response array for kidney and other urogenital system effects (Figure 1-2). 

Note: A dash (“‒”) indicates that the study authors did not measure or report a value for that dose group.

SDMS -- spontaneous death or moribund sacrifice; SS -- scheduled sacrifice

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1 Table 1-6. Six-, 12-, and 24-month incidence of kidney endpoints in male F344 

2 rats reported for statistical evaluation in Levine et al. (1983b) 

Doses (mg/kg-d) 0 0.3 1.5 8.0 40

Medullary papillary necrosis (incidence)

6 mo

SS 0/10 0/10 0/10 0/10 0/10

SDMS − − − − 0/5

Sum 0/10 0/10 0/10 0/10 0/15

12 mo 

SS 0/10 0/10 0/10 0/10 0/10

SDMS − − 0/3 − 15/19*

Sum 0/10 0/10 0/13 0/10 15/29*

24 mo 

SS 0/38 0/36 0/25 0/29 0/4

SDMS 0/17 1/19 0/27 0/26 18/27*

Sum 0/55 1/55 0/52 0/55 18/31*

Pyelitis (incidence)

6 mo 

SS 0/10 0/10 0/10 0/10 0/10

SDMS − − − − 0/5

Sum 0/10 0/10 0/10 0/10 0/15

12 mo 

SS 0/10 0/10 0/10 0/10 0/10

SDMS − − 0/3 − 1/19

Sum 0/10 0/10 0/13 0/10 1/29

24 mo 

SS 0/38 0/36 0/25 0/29 0/4

SDMS 0/17 1/19 0/27 1/26 5/27*

Sum 0/55 1/55 0/52 1/55 5/31*

Pyelonephritis (incidence)

6 mo 

SS 0/10 0/10 0/10 0/10 0/10

SDMS − − − − 0/5

Sum 0/10 0/10 0/10 0/10 0/15

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Doses (mg/kg-d) 0 0.3 1.5 8.0 40

12 mo 

SS 0/10 0/10 0/10 0/10 0/10

SDMS − − 0/3 − 1/19

Sum 0/10 0/10 0/13 0/10 1/29

24 mo 

SS 0/38 0/36 0/25 1/29 0/4

SDMS 0/17 0/19 2/27 1/26 1/27

Sum 0/55 0/55 2/52 2/55 1/31

1 

2 

3 

4 

5 

6 

7 

8 

*Statistically significant (p < 0.05) based on analysis by study authors.

Note: A dash (“‒”) indicates that the study authors did not measure or report a value for that dose group.

SDMS -- spontaneous death or moribund sacrifice; SS -- scheduled sacrifice

Source: Levine et al. (1983b). 

Table 1-7. Six-, 12-, and 24-month incidence of urinary bladder endpoints in 

male F344 rats reported for statistical evaluation in Levine et al. (1983b)

Doses (mg/kg-d) 0 0.3 1.5 8.0 40

Luminal distention (incidence)

6 mo 

SS 0/10 0/10 0/10 0/10 0/10

SDMS − − − − 0/5

Sum 0/10 0/10 0/10 0/10 0/15

12 mo 

SS 0/10 0/10 0/10 0/10 0/10

SDMS − − 0/3 − 18/19*

Sum 0/10 0/10 0/13 0/10 18/29

24 mo 

SS 0/38 0/36 0/25 0/29 1/4*

SDMS 0/16 2/19 1/27 3/22 24/28*

Sum 0/54 2/55 1/52 3/51 25/32*

Cystitis, hemorrhagic/suppurative (incidence)

6 mo 

SS 0/10 0/10 0/10 0/10 0/10

SDMS − − − − 0/5

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Doses (mg/kg-d) 0 0.3 1.5 8.0 40

Sum 0/10 0/10 0/10 0/10 0/15

12 mo 

SS 0/10 0/10 0/10 0/10 0/10

SDMS − − 0/3 − 17/19*

Sum 0/10 0/10 0/13 0/10 17/29

24 mo 

SS 0/38 0/36 0/25 1/29 0/4

SDMS 0/16 2/19 1/27 0/22 18/27*

Sum 0/54 2/55 1/52 1/51 18/31*

1 

2 *

3 

4 N

5 S

6 

7 S

8 

9 

Statistically significant (p < 0.05) based on analysis by study authors.

ote: A dash (“‒”) indicates that the study authors did not measure or report a value for that dose group.

DMS -- spontaneous death or moribund sacrifice; SS -- scheduled sacrifice

ource: Levine et al. (1983b).

Table 1-8. Six-, 12-, and 24-month incidence of prostate endpoints in male 

F344 rats reported for statistical evaluation in Levine et al. (1983b)

Doses (mg/kg-d) 0 0.3 1.5 8.0 40

Spermatic granuloma (incidence)

6 mo 

SS 0/10 2/10 2/10 1/10 6/10*

SDMS − − − − 2/5

Sum 0/10 2/10 2/10 1/10 8/15*

12 mo 

SS 0/10 0/10 1/10 1/10 0/10

SDMS − − 0/3 − 0/19

Sum 0/10 0/10 1/13 1/10 0/29

24 mo 

SS 0/38 0/36 0/25 0/29 0/4

SDMS 0/16 0/19 0/27 0/26 0/27

Sum 0/54 0/55 0/52 0/55 0/31

Suppurative inflammation (incidence)

6 mo 

SS 0/10 0/10 0/10 0/10 0/10

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Doses (mg/kg-d) 0 0.3 1.5 8.0 40

SDMS − − − − 0/5

Sum 0/10 0/10 0/10 0/10 0/15

12 mo 

SS 0/10 0/10 0/10 0/10 0/10

SDMS − − 0/3 − 0/19

Sum 0/10 0/10 0/13 0/10 0/29

24 mo 

SS 0/38 1/36 2/25* 4/29* 0/4

SDMS 2/16 3/19 7/27* 8/26 19/27*

Sum 2/54 4/55 9/52* 12/55* 19/31*

1 

2 

3 

4 

5 

6 

7 

*Statistically significant (p < 0.05) based on analysis by study authors.

Note: A dash (“‒”) indicates that the study authors did not measure or report a value for that dose group.

SDMS -- spontaneous death or moribund sacrifice; SS -- scheduled sacrifice

Source: Levine et al. (1983b). 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 Figure 1-2. Exposure-response array of kidney and other urogenital system effects.

0.1

1

10

100

1000

Lish et al., 1984 (mouse, male)

Lish et al., 1984 (mouse, female)

Hart, 1976 (rat)

Levine et al., 1983b (rat)

Cholakis et al., 1980 (mouse)Cholakis et al., 1980 (mouse)

Cholakis et al., 1980 (rat)

Cholakis et al., 1980 (F0 rat, male)

Cholakis et al., 1980 (F0 rat, female)

Crouse et al., 2006 (rat, male)

Crouse et al., 2006 (rat, female)

Levine et al., 1990, 1981a,b (rat)

Lish et al., 1984 (mouse)

Hart, 1976 (rat)

uremic mineralization; Levine et al., 1983b (rat)

necrosis, pyelitis; Levine et al., 1983b (rat) (1)

luminal distention; Levine et al., 1983b (rat)

cystitis; Levine et al., 1983b (rat) (2)

suppurative prostitis: Levine et al., 1983b (rat)

tubular nephrosis, Cholakis et al., 1980 (mouse)

Cholakis et al., 1980 (rat)

cortical cysts, Cholakis et al., 1980 (rat)

prostate inflammation, Crouse et al., 2006 (rat)

Levine et al., 1990, 1981a,b (rat)

Hart, 1974 (dog)

mineralization, Martin & Hart, 1974 (monkey)

Chronic Subchronic Chronic Subchronic

Absolute kidney weight Histopathological lesions

Dose (mg/kg-day)

significantly changed

not significantly changed

Note: Filled circle indicates that response was statistically significantly different from the control.

(1) Statistical significance determined from incidence at time of scheduled sacrfice. (2) Statistical significance determined from incidence at spontaneous death. 

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Mechanistic Evidence

1 No MOA information is available for RDX-induced kidney and other urogenital effects. 

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However, mechanistic information underlying the neurotoxicity observed with RDX exposure, and 

the specific affinity of RDX to the GABAA receptor-convulsant site (Williams et al., 2011; Williams 

and Bannon, 2009), suggests a biologically plausible role for the GABAA receptor in RDX-related 

effects on the urogenital system, and provides some potential MOA hypotheses for the effects 

reported in Levine et al. (1983b). 

One hypothesis is that urogenital effects of RDX are caused by interactions with GABAA

receptors mediating inputs to the urogenital system. GABA and GABA receptors have been 

identified in a number of peripheral tissues (Erdö et al., 1991; Ong and Kerr, 1990; Erdo, 1985). 

Brar et al. (2014) demonstrated that pretreatment with picrotoxin reduced the renoprotective 

effects of sodium valproate (which acts on both GABAA and GABAB receptors) in a rat model of 

ischemia-induced acute kidney injury, suggesting that GABAA receptors may be important in renal 

function. GABA is believed to play a role in the regulation of urination and bladder capacity 

(reviewed in Fowler et al. (2008) and Yoshimura and de Groat (1997)). In rats, injection of a GABAA

receptor agonist inhibits the urination reflex (Igawa et al., 1993; Kontani et al., 1987). GABAA

agonists injected into the periaqueductal gray area in rats inhibited reflex bladder activity, while 

injection of an antagonist reduced bladder capacity and increased the frequency of bladder reflex 

activity (Stone et al., 2011). RDX would be expected to act like an antagonist and increase bladder 

activity (which would not result in urinary stasis), although the impact of chronic exposure to RDX 

acting as a GABAA receptor antagonist is not known. Evidence of GABAergic signaling regulating 

bladder function, and the hypothesized disruption of that regulation by RDX via interaction with 

GABAA receptors, may plausibly account for the kidney and other urogenital lesions, including 

suppurative prostatitis, observed by Levine et al. (1983b); however, no evidence to support this 

hypothesized MOA is available.

Other potential mechanisms by which RDX, through GABAA binding, may lead to kidney and 

urogenital effects are less apparent. Alterations in hormonal signaling or circulating levels of 

estrogen or prolactin may lead to prostatitis. Prostate inflammation has been associated with 

endocrine disruptors in the environment (Cowin et al., 2010), and increased prolactin has been 

shown to cause lateral lobe prostatitis (Stoker et al., 1999b; Stoker et al., 1999a; Tangbanluekal and 

Robinette, 1993; Robinette, 1988). Typically, the inflammation seen is chronic and does not reverse 

over time (Robinette, 1988). Functional GABAA receptors have been identified in the anterior 

pituitary (Zemkova et al., 2008; Mayerhofer et al., 2001), which also serves as the primary source of 

prolactin. Thus, the prostate inflammation observed in the rat in the 2-year study by Levine et al. 

(1983b) could have been produced by disruption of pituitary prolactin or another hormonal signal 

via interference with normal regulatory GABA-related hormonal control. However, no direct 

evidence for this hypothesized MOA is available. Levine et al. (1983b) did not evaluate serum 

endocrine measures or pituitary weights, and pituitary adenomas that could account for higher 

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1 prolactin levels were not observed. A MOA hypothesis based on pituitary-mediated alterations in 

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endocrine signaling also does not explain the other urogenital lesions observed by Levine et al. 

(1983b). 

Another hypothesis is that the prostate effects could be mediated through an autoimmune 

inflammatory response. GABAA receptor transcripts have been identified in immune cells of mouse 

models (Reyes-García et al., 2007; Tian et al., 2004), and GABAA receptor agonists have decreased 

cytotoxic immune responses and hypersensitivity reactions (Tian et al., 1999; Bergeret et al., 1998). 

In a mouse autoimmune model of multiple sclerosis, Bhat et al. (2010) found that treatment of 

macrophages challenged with lipopolysaccharide with various GABA agonists decreased cytokine 

production; addition of picrotoxin (which may have effects similar to those of RDX, as it binds to the 

same site) was able to reduce this effect. However, picrotoxin on its own did not significantly alter 

cytokine production, suggesting that the effects are limited to reversal of agonist-induced 

GABAergic activity (Johnson, 2015b). If an autoimmune mechanism was contributing to the effects 

observed with RDX exposure, it is unclear why inflammation would be limited to the prostate. RDX 

has also tested negative in the only battery of immunotoxicity tests to which it was subjected

(Crouse et al., 2006). 

If the urogenital effects are mediated through localized interaction with GABAA receptors, 

another possibility would be that that effects would result from direct interactions with GABAA

receptors located on the prostate. GABAA receptors have been identified on the prostate

(Napoleone et al., 1990), providing a potential mechanism by which RDX could interact directly 

with the prostate. However, this would require that the prostate is actively maintained in a noninflamed state, mediated by GABA; RDX binding to GABAA receptor-convulsant sites on the prostate 

would result in a reduction of the inhibitory effects of the GABA receptor, leading to increased 

inflammation (Johnson, 2015b). No evidence was found to support this potential pathway leading 

to prostate inflammation. 

In summary, there are no studies available that inform mechanistically how RDX might lead 

to kidney and other urogenital effects. There is evidence that RDX binds to GABAA receptors in 

neuronal tissues (Williams et al., 2011; Williams and Bannon, 2009), and it is biologically plausible 

that binding to the GABA receptor could occur in other tissues as well, contributing to the observed 

kidney and urogenital effects. Among the mechanistic information presented above, MOAs that 

require direct action on the prostate are considered less likely because the available information 

suggests that the prostatitis is a secondary effect. However, the ways that GABAA receptors work in 

non-neuronal tissues and organs is still not well understood, and the MOA by which RDX induces 

kidney and other urogenital effects is unknown. 

Integration of Kidney and Urogenital System Effects

Evidence for kidney effects resulting from RDX exposure consists of human case reports and 

findings of histopathological changes in rodents. In humans, evidence for kidney effects (including 

decreased urine output, blood in urine, and proteinuria) is limited to individuals with acute 

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1 accidental exposure (ingestion and inhalation) to unknown amounts of RDX. No RDX-related 

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changes in kidney parameters were found in a small cross-sectional study of RDX-exposed workers

(Hathaway and Buck, 1977). 

A dose-related increase in the incidence of suppurative prostatitis in male rats (Levine et al., 

1983b) provides the strongest evidence of RDX-associated kidney and other urogenital system 

effects. As discussed above, the incidence of suppurative prostatitis is considered to be an indicator

for the broader array of kidney and other urogenital effects seen in this study. Levine et al. (1983b)

identified other histopathological effects (papillary necrosis, pyelitis, luminal distension, and 

cystitis) in the kidney and bladder, but at the highest dose only. A second 2-year study in SpragueDawley rats found no histopathological changes in the kidney or urogenital system (Hart, 1976), 

but exposure levels used in this study were low compared to Levine et al. (1983b). Other measures 

of kidney effects, specifically kidney weights and serum chemistry parameters, did not provide 

consistent evidence of dose-related changes associated with RDX exposure. In light of the doserelated increase in suppurative prostatitis and the lack of support for an alternative (i.e., non-RDXrelated) basis for this effect, kidney and urogenital effects are a potential human hazard of RDX

exposure. 

1.2.3. Reproductive and Developmental Effects 

No human studies were identified that evaluate the potential of RDX to cause reproductive 

or developmental effects. Information relevant to an examination of the association between RDX 

exposure and reproductive and developmental effects comes from a 2-generation reproductive 

toxicity study in rats and developmental studies in rats and rabbits involving oral administration of 

RDX during gestation. In addition, oral subchronic and chronic studies in experimental animals 

provide information useful for examining the association between RDX exposure and effects 

specifically on the male reproductive system. A summary of the reproductive and developmental 

effects associated with RDX exposure is presented in Tables 1-9 and 1-10 and Figures 1-3 and 1-4. 

Studies are ordered in the evidence tables and exposure-response arrays by duration of exposure 

and then by species.

Reproductive Effects

Evidence of male reproductive toxicity is provided by the finding of testicular degeneration 

in male mice. An increased incidence of testicular degeneration (10−11%) was observed in male 

B6C3F1 mice exposed to ≥35 mg/kg-day RDX for 2 years in the diet compared to concurrent (0%) 

and historical (1.5%) controls (Lish et al., 1984). Reductions in absolute testicular weight were 

observed, but the magnitude of this effect was small (≤6% compared to controls) and not doserelated. An increased incidence of germ cell degeneration was observed in rats exposed to 

40 mg/kg-day (40%) compared with controls at 12 months (0%); by 24 months, almost all male 

rats (including controls) had testicular masses (interstitial cell tumors), and no instances of germ 

cell degeneration were identified in control or RDX-treated groups (Levine et al., 1983b). No doseThis document is a draft for review purposes only and does not constitute Agency policy.

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1 related histopathological changes in the testes were identified in other studies in rats (Crouse et al., 

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2006; Levine et al., 1990; Levine et al., 1981a, b; Hart, 1976) or dogs (Hart, 1974). Changes in 

testicular weight were inconsistent across studies, with an equivalent number of studies identifying 

decreases (Crouse et al., 2006; Lish et al., 1984; Cholakis et al., 1980) or increases (Levine et al., 

1990; Levine et al., 1981a, b; Cholakis et al., 1980; Hart, 1976, 1974) in testicular weight; in most 

cases, the changes in testicular weight were small (≤10% change compared to control) and not 

dose-related. 

Reproductive function was assessed in two separate studies reported by Cholakis et al. 

(1980). No specific effects on reproductive function were observed in F0 and F1 CD rats exposed to 

≤16 mg/kg-day RDX in the Cholakis et al. (1980) two-generation study. The highest dose tested, 

50 mg/kg-day, was associated with reductions in fertility (specifically a decreased number of 

pregnancies) in the F0 generation, although these changes were not statistically significant. The 

finding of lower fertility rates only at the 50 mg/kg-day dose, a dose associated with reduced body 

weight and feed consumption and increased mortality (9% in male rats and 27% in female rats), 

suggests that effects on reproductive function were likely due to the general toxicity of RDX rather 

than a direct effect of RDX on reproduction. In the dominant lethal mutation study, which used the 

F0 males from the two-generation reproductive toxicity study, no effects on fertility were observed 

in male rats exposed to ≤16 mg/kg-day RDX. Pregnancy rates were lower in untreated females 

mated to males exposed to 50 mg/kg-day RDX for 15 weeks prior to mating; the authors attributed

this effect to a treatment-related decrease in the well-being of the males in this high-dose group

(Cholakis et al., 1980). 

Developmental Effects

Animal studies have reported decreases in offspring survival following administration of 

RDX. Pup survival rates in the F0 and F1 generations (including both stillborn pups and postnatal 

deaths through the age of weaning) were statistically significantly decreased in RDX-exposed CD 

rats compared to controls in the only available two-generation reproductive toxicity study of RDX 

(Cholakis et al., 1980). This observation was noted only at the highest dose tested (50 mg/kg-day) 

that also produced toxicity in adults (mortality [18%], reduced body weights [8−14%], and reduced 

food consumption [10−17%]). Decreased fetal viability was observed at the highest dose tested, 20 

mg/kg-day, in a developmental toxicity study in F344 rats (Cholakis et al., 1980), although no effect 

on live fetuses was observed in a developmental toxicity study in Sprague-Dawley rats at the same 

dose (Angerhofer et al., 1986); both of these studies reported significant mortality (29−31%) in 

dams at 20 mg/kg-day. Increased resorptions were similarly limited to the highest dose tested (20 

mg/kg-day) (Cholakis et al., 1980). Both of these studies started treatment with RDX on gestational

day (GD) 6, which may contribute to the incidence of resorptions observed in the control and 

treated groups. As noted in EPA’s Guidelines for Developmental Toxicity Risk Assessment (U.S. EPA, 

1991), treatment beginning around the time of implantation may result in an increase in 

implantation loss that reflects variability that is not treatment related. There was no evidence of 

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1 maternal toxicity, embryotoxicity, or decreased fetal viability in a teratology study of pregnant New 

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Zealand White (NZW) rabbits administered RDX by gavage from GD 7 to 29 at doses up to 20 

mg/kg-day (Cholakis et al., 1980), suggesting that rabbits may be less sensitive to RDX toxicity than 

rats.

Statistically significant, dose-related reductions in fetal body weight and length were 

reported in Sprague-Dawley rats administered RDX by gavage from GD 6 to 15 (Angerhofer et al., 

1986).9 Decreased fetal body weight (9%) and body length (5%), with statistically significant 

trends, were observed at 20 mg/kg-day, a dose that produced significant (31%) mortality in the 

dams. A similar reduction in fetal body weight of 7% (not statistically significant) was observed in 

F344 rats exposed to RDX at 20 mg/kg-day, a dose associated with 29% maternal mortality 

(Cholakis et al., 1980). Dose-related reductions in fetal body weight were not observed in NZW 

rabbits at doses up to 20 mg/kg-day (Cholakis et al., 1980). 

No treatment-related effects on morphological development have been reported in rats 

exposed to a dose as high as 20 mg/kg-day RDX, a dose that resulted in 29−31% maternal mortality 

(Angerhofer et al., 1986; Cholakis et al., 1980). Examination of rabbits administered RDX at doses 

up to 20 mg/kg-day from GD 7 to 29 also provided no evidence of treatment-related developmental 

anomalies (Cholakis et al., 1980). Although increased incidences of enlarged frontal fontanel and 

unossified sternebrae were observed in fetuses of all groups of NZW rabbits administered RDX 

(Cholakis et al., 1980), these developmental anomalies did not exhibit a dose-related increase in the 

number of either fetuses or litters affected, and were thus interpreted as not being treatmentrelated by the study authors (Cholakis et al., 1980). This interpretation is supported by the 

following additional considerations. Neither individual litter data nor historical control data from 

the performing laboratory were available to assist in the interpretation of these findings. A report 

of historical control incidences of fetal skeletal observations in NZW rabbits for 224 prenatal 

developmental toxicology studies conducted in 8 contract research laboratories during the period 

of 1988−1992 (MTA, 1992) included findings from 26,166 fetuses of 3,635 litters. Background 

control incidences of enlarged anterior fontanel were observed in 8 fetuses (0.031%) of 7 litters 

(0.193%), while sternebrae agenesis (which may not be entirely comparable to the finding of 

unossified sternebrae in Cholakis et al. (1980) was found in 10 fetuses (0.038%) of 5 litters 

(0.138%). Although the use of concurrent control data is preferable for the interpretation of 

developmental toxicity data, this historical information supports the low control incidences of these 

findings in the Cholakis et al. (1980) study as being within typical historical parameters. It is also 

noted that the non-dose-related pattern of increased enlarged fontanel and unossified sternebrae 

across treated groups in Cholakis et al. (1980) was similar to the pattern of decreases in fetal body 

9The statistical analyses presented by the study authors were performed on a per fetus basis; EPA’s Guidelines 

for Developmental Toxicity Risk Assessment (U.S. EPA, 1991) recommend that fetal data be analyzed on a per 

litter (rather than per fetus) basis. In a reanalysis of the Angerhofer et al. (1986) data by EPA on a per litter 

basis, fetal body weight and length showed statistically significant decreasing trends.

This document is a draft for review purposes only and does not constitute Agency policy.

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1 weight in the same study, suggesting a possible link between these particular sternebral and 

2 

3 

4 

5 

6 

7 

fontanel anomalies with fetal growth status. Given the lack of dose-related increases in the 

incidences of these anomalies, and patterns that mirrored fetal body weight decreases (which were 

also not dose-related), the findings of enlarged frontal fontanel and unossified sternebrae were not 

considered treatment-related. Gestational administration of RDX to NZW rabbits did not result in 

any other dose- and treatment-related skeletal abnormalities. 

Table 1-9. Evidence pertaining to male reproductive effects in animals 

Reference and study design Results

Lish et al. (1984)

Mice, B6C3F1, 85/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles <66 μm

0, 1.5, 7.0, 35, or 175/100 mg/kg-d (high 

dose reduced to 100 mg/kg-d in wk 11 

due to excessive mortality)

Diet

2 yrs

Doses 0 1.5 7.0 35 175/100

Testicular degeneration (incidence)

0/63 2/60 2/62 6/59 3/27a

Absolute testes weight; wk 105 (percent change compared to 

control)

0% −6% 0% −2% −6%

Relative testes weight; wk 105 (percent change compared to control)

0% −4% 2% −2% −2%

Hart (1976)

Rats, Sprague-Dawley, 100/sex/dose

Purity and particle size not specified

0, 1.0, 3.1, or 10 mg/kg-d 

Diet

2 yrs

Doses 0 1.0 3.1 10

Absolute testes (with epididymis) weight; wk 104 

0% −2% 2% 5%

Relative testes (with epididymis) weight; wk 104 

0% −1% 7% 9%

Testes were examined microscopically in control and 10 mg/kg-d 

groups; no degeneration or other treatment-related effects were 

observed. 

Levine et al. (1983b)

Rats, F344, 75/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles <66 μm

0, 0.3, 1.5, 8.0, or 40 mg/kg-d 

Diet

2 yrs

Doses 0 0.3 1.5 8.0 40

Testes, germ cell degeneration; 12 mob (incidence)

SS 0/10 0/10 0/10 0/10 4/10*

SDMS − − 1/3 − 4/19

Testes, germ cell degeneration; 24 mo (incidence)

SS 0/38 0/36 0/25 0/29 0/4

SDMS 0/16 0/19 0/27 0/26 0/27

Testes weights were not measured at termination due to testicular 

masses in nearly all males.

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Reference and study design Results

Cholakis et al. (1980)

Mice, B6C3F1, 10−12/sex/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

Experiment 1: 0, 10, 14, 20, 28, or 

40 mg/kg-d

Diet

13 wks

Doses 0 10 14 20 28 40

Absolute testes weight (percent change compared to control)

0% − − − −4% −4%

Relative testes weight (percent change compared to control)

0% − − − 2% −1%

Experiment 2: 0, 40, 60, or 80 mg/kg-d for 

2 wks followed by 0, 320, 160, or 

80 mg/kg-d (TWA doses of 0, 79.6, 147.8, 

or 256.7 mg/kg-d for males and 0, 82.4, 

136.3, or 276.4 mg/kg-d for females)c

Diet

13 wks

Doses 0 80 160 320

Absolute testes weight (percent change compared to control)

0% 4% −4% −8%

Relative testes weight (percent change compared to control)

0% 1% −4% −9%

Testes were examined microscopically in control and 320 mg/kg-d 

groups; no effects were observed.

Cholakis et al. (1980)

Rats, F344, 10/sex/dose

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

0, 10, 14, 20, 28, or 40 mg/kg-d

Diet

13 wks

Doses 0 10 14 20 28 40

Absolute testes weight (percent change compared to control)

0% − − − −2% 0%

Relative testes weight (percent change compared to control)

0% − − − 2% 9%

Testes were examined microscopically in control and 40 mg/kg-d 

groups; no effects were observed. 

Cholakis et al. (1980)

Rats, CD, two-generation study; F0: 

22/sex/group; F1: 26/sex/group; F2: 

10/sex/group 

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

F0 and F1 parental animals: 0, 5, 16, or 

50 mg/kg-d

Diet

F0 exposure: 13 wks pre-mating, and 

during mating, gestation, and lactation of 

F1; F1 exposure: 13 wks after weaning, 

and during mating, gestation, and 

lactation of F2; F2 exposure: until weaning

In F2 weanling offspring of 0, 5, and 16 mg/kg-d groups. No highdose F2 animals available.

Doses 0 5 16 50

Absolute testes weight (percent change compared to control)

0% 3% −31% −

Testes of F2 weanlings were examined microscopically in all groups; 

no effects were observed.

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Reference and study design Results

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d 

Gavage

13 wks

Doses 0 4 8 10 12 15

Absolute testes weight (percent change compared to control)

0% −3% −5% −4% −4% −8%

Relative testes weight (percent change compared to control)

0% 4% 5% 0% −6% −10%*

Levine et al. (1981a); Levine et al. (1990); 

Levine et al. (1981b)d

Rats, F344, 10/sex/group; 30/sex for 

control

84.7 ± 4.7% purity, ~10% HMX, median 

particle diameter 20 μm, ~90% of particles 

≤66 μm

0, 10, 30, 100, 300, or 600 mg/kg-d 

Diet

13 wks

Doses 0 10 30 100 300 600

Testes, germ cell degeneration (incidence)

0/10 0/10 0/10 0/10 1/9 1/10

Absolute testes weight (percent change compared to control)

0% 1% 1% −2% − −

Relative testes weight (percent change compared to control)

0% 4% 5% 19%* − −

Hart (1974)e

Dogs, Beagle, 3/sex/dose

Pre-mix with ground dog chow containing 

20 mg RDX/g-chow, 60 g dog food; purity 

and particle size not specified

0, 0.1, 1, or 10 mg/kg-d 

Diet

13 wks

Doses 0 0.1 1 10

Absolute testes (with epididymis) weight (percent change compared 

to control)

0% − − 51%

Testes were not examined microscopically.

1 

2 

3 

4 

5 

6 

7 

8 

9 

10

11

12

13

14

15

16

17

18

*Statistically significant (p < 0.05) based on analysis by study authors. a

Although the study authors did not observe a statistically significant increase in the incidence of testicular 

degeneration, they determined that the incidences at the 35 and 175/100 mg/kg-day dose groups were “notable” 

when compared to concurrent (0%) and historical (1.5%) incidences. b

Testicular atrophy was observed at 12 months along with a statistically reduced mean testes weight (compared 

with controls). By 24 months, almost all male rats (including controls) had testicular masses (interstitial cell 

tumors); testes weights were not recorded, and an increased incidence of testicular degeneration was not 

observed. 

c

Doses were calculated by the study authors.

d

Levine et al. (1981a) is a laboratory report of a 13-week study of RDX in F344 rats; two subsequently published 

papers (Levine et al., 1990; Levine et al., 1981b) present subsets of the data provided in the full laboratory report. e

Because testes weight was reported for only three treated animals in this study, organ data from this study were 

considered less informative than other studies; therefore, testes weights from Hart (1974) were not presented in 

the exposure-response array for male reproductive effects.

Note: A dash (“‒”) indicates that the study authors did not measure or report a value for that dose group.

SDMS -- spontaneous death or moribund sacrifice; SS -- scheduled sacrifice

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1 Figure 1-3. Exposure response array of male reproductive effects following oral exposure.

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1 Table 1-10. Evidence pertaining to reproductive and developmental effects in 

2 animals 

Reference and study design Results

Offspring survival

Cholakis et al. (1980)

Rats, CD, two-generation study;

F0: 22/sex/group; F1: 26 sex/group; 

F2: 10 sex/group 

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

F0 and F1 parental animals: 0, 5, 16, or 

50 mg/kg-d 

Diet

F0 exposure: 13 wks pre-mating, and 

during mating, gestation, and lactation of 

F1; F1 exposure: 13 wks after weaning, 

and during mating, gestation, and 

lactation of F2; F2 exposure: until weaning 

Doses 0 5 16 50

Stillborn pups (incidence)

F1 8/207 6/296 4/259 16/92*

F2 6/288 6/290 2/250 24/46*

Offspring survival at birth (percent of fetuses)

F1 96% 98% 98% 83%*

F2 98% 98% 99% 48%*

Survival at weaning (percent of liveborn pups)

F1 87% 96% 90% 8%

F2 79% 86% 79% 0%

F0 maternal deaths occurred at 50 mg/kg-d. Only six F1 females in 

this group survived to serve as parental animals; none of the 

surviving six died during subsequent treatment.

Note: results on a per litter basis were not provided.

Cholakis et al. (1980)

Rabbits, NZW, 11−12/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

0, 0.2, 2.0, or 20 mg/kg-d 

Gavage

GDs 7−29

Doses 0 0.2 2 20

Early resorptions (mean percent per dam)

6% 5% 4% 1%

Late resorptions (mean percent per dam)

8% 5% 3% 3%

Viable fetuses (mean percent per dam)

85% 82% 77% 94%

Cholakis et al. (1980)

Rats, F344, 24−25 females/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants.

0, 0.2, 2.0, or 20 mg/kg-d 

Gavage

GDs 6−19

Doses 0 0.2 2.0 20

Early resorptions (mean percent per dam)

6.0% 2.5% 4.8% 15.3%

Late resorptions (mean percent per dam)

0.5% 0.5% 0.3% 1.6%

Complete litter resorptions (number of litters)

0 0 0 2

Viable fetuses (mean percent per dam)

93.2% 97.6% 94.9% 81.4%

Significant maternal mortality (7/24 dams) occurred at 20 mg/kg-d.

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Reference and study design Results

Angerhofer et al. (1986)

Rats, Sprague-Dawley, 39−51 mated 

females/group (25−29 pregnant 

dams/group)

Purity 90%; 10% HMX and 0.3% acetic acid 

occurred as contaminants

0, 2, 6, or 20 mg/kg-d

Gavage

GDs 6−15

Doses 0 2 6 20

Resorptions (percent of total implantations)

4.8% 6.1% 5.9% 6.4%

Early resorptions (percent of total implantations)

4.8% 6.1% 5.9% 6.2%

Late resorptions (percent of total implantations)

0% 0% 0% 0.27%

Live fetuses (mean percent per litter)

100% 100% 100% 100%

Significant maternal mortality (16/51) occurred at 20 mg/kg-d. 

Percent resorptions and live fetuses based on number of surviving 

females at time of necropsy. 

Offspring growth

Cholakis et al. (1980)

Rabbits, NZW, 11−12/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

0, 0.2, 2.0, or 20 mg/kg-d

Gavage

GDs 7−29

Doses 0 0.2 2.0 20

Fetal body weight (percent change compared to control)

0% −6.7% −2.3% −9.3%

Cholakis et al. (1980)

Rats, F344, 24−25 females/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants.

0, 0.2, 2.0, or 20 mg/kg-d

Gavage

GDs 6−19

Doses 0 0.2 2.0 20

Fetal body weight (percent change compared to control)

0% 2% 3% −7%

Significant maternal mortality (7/24 dams) occurred at 20 mg/kg-d.

Angerhofer et al. (1986)

Rats, Sprague-Dawley, 39−51 mated 

females/group (25−29 pregnant 

dams/group)

Purity 90%; 10% HMX and 0.3% acetic acid 

occurred as contaminants

0, 2, 6, or 20 mg/kg-d

Gavage

GDs 6−15

Doses 0 2 6 20

Fetal body weight (percent change compared to control)

0% −4% −2% −9%a

Fetal body length (percent change compared to control)

0% −1% −1% −5%a

Significant maternal mortality (16/51) occurred at 20 mg/kg-d.

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Reference and study design Results

Morphological development

Cholakis et al. (1980)

Rabbits, NZW, 11−12/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants; ~200 μm particle size

0, 0.2, 2.0, or 20 mg/kg-d 

Gavage

GDs 7−29

Doses 0 0.2 2.0 20

Spina bifida (incidence)

Fetuses 0/88 0/99 0/94 3/110

Litters 0/11 0/11 0/11 2/12

Misshapen eye bulges (incidence)

Fetuses 0/88 0/99 0/94 3/110

Litters 0/11 0/11 0/11 1/12

Cleft palate (incidence)

Fetuses 0/39 1/46 2/44 2/52

Litters 0/11 1/11 1/11 1/12

Enlarged front fontanel (incidence)

Fetuses 0/49 5/53 2/50 8/58

Litters 0/11 2/11 2/11 2/12

Unossified sternebrae (incidence)

Fetuses 4/49 12/53 8/50 12/58

Litters 4/11 7/11 4/11 6/12

Cholakis et al. (1980)

Rats, F344, 24−25 females/group

88.6% pure, with 9% HMX and 2.2% water 

as contaminants.

0, 0.2, 2.0, or 20 mg/kg-d 

Gavage

GDs 6−19

No gross or soft-tissue anomalies were seen in any exposure group. 

No treatment-related increase in the incidence of litters with skeletal 

anomalies was observed. Significant maternal mortality (7/24 dams) 

occurred at 20 mg/kg-d.

Angerhofer et al. (1986)

Rats, Sprague-Dawley, 39−51 mated 

females/group (25−29 pregnant 

dams/group)

Purity 90%; 10% HMX and 0.3% acetic acid 

occurred as contaminants

0, 2, 6, or 20 mg/kg-d 

Gavage

GDs 6−15

No treatment-related increase in the incidence of anomalies was 

observed.

Doses 0 2 6 20

Total malformations (percent of fetuses with malformations)

1% 1% 0% 2%

Significant maternal mortality (16/51) occurred at 20 mg/kg-d.

1 

2 

3 

4 

*Statistically significant (p < 0.05) based on analysis by study authors. a

Statistically significant dose-related trend (p < 0.05) by linear trend test, performed for this assessment. Average 

fetal weights or lengths for each litter comprised the sample data for this test.

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Figure 1-4. Exposure response array of reproductive and developmental effects following oral exposure. 

0.1

1

10

100

Cholakis et al., 1980 (rat, F1 and F2)

Cholakis et al., 1980 (rabbit)

Cholakis et al., 1980 (rat)

Angerhofer et al., 1986 (rat)

Cholakis et al., 1980 (rabbit)

Cholakis et al., 1980 (rat)

Angerhofer et al., 1986 (rat) (1)

Cholakis et al., 1980 (rabbit)

Cholakis et al., 1980 (rat)

Angerhofer et al., 1986 (rat)

↓ Offspring survival Offpsring growth Morphological

development

Dose (mg/kg-day)

signficantly changed

not signifcantly changed

Note: Filled circle indicates that response was statistically significantly different from the control.

(1) Statistically signficant dose-related trend (p <= 0.05) by linear trend test, performed for this assessment. 

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Integration of Reproductive and Developmental Effects

1 Testicular effects were reported in male B6C3F1 mice chronically exposed to RDX in the diet 

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5 

6 

7 

8 

9 

10

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for 24 months (Lish et al., 1984). No other studies of equivalent duration were performed in mice 

to determine the consistency of this effect. Germ cell degeneration was observed in F344 rats at 

12 months, but not at 24 months, in a 2-year study (Lish et al., 1984); therefore, the biological 

significance of the 12-month findings is uncertain. Other testicular effects were inconsistent across 

rat studies. Based on the evidence of testicular degeneration in male mice reported by Lish et al. 

(1984), there is suggestive evidence of male reproductive effects associated with RDX exposure. 

Developmental studies in rats (Angerhofer et al., 1986; Cholakis et al., 1980) demonstrated 

effects on offspring survival, growth, and morphological development only at doses associated with 

severe maternal toxicity and mortality. No dose-related developmental effects were observed in 

rabbits (Cholakis et al., 1980). As noted in EPA’s Guidelines for Developmental Toxicity Risk 

Assessment (U.S. EPA, 1991), where adverse developmental effects are produced only at doses that 

cause minimal maternal toxicity, developmental effects should not be discounted as being

secondary to maternal toxicity; however, at doses causing excessive toxicity, as is the case with 

RDX, information on developmental effects may be difficult to interpret and of limited value. 

Therefore, at this time, no conclusions are drawn regarding developmental effects as a human 

hazard of RDX exposure.

1.2.4. Liver Effects

One occupational epidemiology study examined the association between RDX exposure and 

changes in serum liver enzymes. Case reports involving accidental exposure to RDX provide 

information on the potential for acute exposure to RDX to affect the liver in humans. In addition, 

organ weight, histopathology, and serum chemistry findings from experimental animal studies 

involving subchronic and chronic exposure to ingested RDX provide data relevant to an 

examination of the association between RDX exposure and liver effects. A summary of the liver 

effects associated with RDX exposure is presented in Tables 1-11 and 1-12 and Figure 1-5. 

Experimental animal studies are ordered in the evidence table and exposure-response array by 

duration of exposure and then by species.

Reports in humans provide inconsistent evidence of liver toxicity associated with acute 

exposure to RDX. Elevated serum levels of aspartate aminotransferase (AST) and/or alanine 

aminotransferase (ALT) were reported in several case reports of individuals who ingested 

unknown amounts of RDX (Küçükardali et al., 2003; Woody et al., 1986; Knepshield and Stone, 

1972; Hollander and Colbach, 1969; Stone et al., 1969; Merrill, 1968) (see Appendix C, Section C.2). 

Liver biopsies did not reveal any abnormal observations (Stone et al., 1969). In other case reports, 

no significant changes in serum levels of liver enzymes were observed (Testud et al., 1996a; Ketel 

and Hughes, 1972). In a cross-sectional epidemiologic study of workers from five U.S. Army 

munitions plants (69 exposed to RDX alone and 24 to RDX and HMX; RDX exposure range: 

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undetectable (<0.01 mg/m3) to 1.6 mg/m3 1 ) (Hathaway and Buck, 1977), serum chemistry analysis 

2 

3 

4 

5 

6 

7 

8 

9 

10

11

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32

(including the serum liver enzymes AST, ALT, and alkaline phosphatase [ALP]) revealed no 

statistically significant differences between exposed and unexposed workers (Table 1-11). 

In experimental animals, some, but not all, subchronic studies reported increased liver 

weight associated with RDX exposure (Table 1-12 and Figure 1-5). Dose-related increases in 

relative liver weight10 (11−25% in high-dose groups) were observed in male and female B6C3F1

mice given RDX in the diet for 90 days (Cholakis et al., 1980) and in female F344 rats in two 

separate 90-day dietary studies of RDX (Levine et al., 1990; Levine et al., 1981a, b; Cholakis et al., 

1980); however, relative liver weights were not increased in female F344 rats in another 90-day 

gavage study (Crouse et al., 2006). Male F344 rats exhibited an increase in relative liver weight 

only in one of these subchronic studies (Levine et al., 1990; Levine et al., 1981a, b). In subchronic 

studies in other species, absolute liver weights were increased in male and female monkeys

(6−16% relative to control at 1 and 10 mg/kg-day) (Martin and Hart, 1974) and in male, but not 

female, beagle dogs (53% relative to control in male dogs at 10 mg/kg-day) (Hart, 1974). 

Chronic RDX exposures in B6C3F1 mice and F344 or Sprague-Dawley rats showed a less 

consistent pattern of liver weight increases. Interpretation of liver weight increases in the 2-year 

mouse study is complicated by the incidence of adenomas and carcinomas in each dose group; the 

apparent increase in liver weights in male and female mice exposed to RDX in diet (Lish et al., 1984) 

was reduced when mice with liver adenomas or carcinomas were removed from the analysis. In a 

2-year rat study (Levine et al., 1983b), relative liver weights were increased in high-dose 

(40 mg/kg-day) males and females (by 11 and 18% compared to controls, respectively), likely 

reflecting the depressed weight gain in the high-dose rats (2-30% in males and 10-15% in 

females). In evaluating organ weight data across studies of all durations, less weight is placed on 

evidence of organ weight changes from chronic (2-year) studies because normal physiological 

changes associated with aging and intercurrent disease contributes to inter-animal variability that 

could confound organ weight interpretation (Sellers et al., 2007), as is true of the mouse liver 

weight data for RDX. 

Nonneoplastic histopathological changes in the liver were not associated with RDX 

exposure in the majority of experimental animal studies (Crouse et al., 2006; Levine et al., 1990; 

Lish et al., 1984; Levine et al., 1983b; Levine et al., 1981a, b; Hart, 1976, 1974; Martin and Hart, 

1974), including 2-year oral studies in mice at doses up to 100 mg/kg-day (Lish et al., 1984) and in 

rats at doses up to 40 mg/kg-day (Levine et al., 1983b). The few findings of liver lesions were 

10Based on an evaluation of the relationship between organ weight and body/brain weight to determine 

which endpoint (organ weight, organ-to-body weight ratio, or organ-to-brain weight ratio) is likely to more 

accurately detect target organ toxicity, Bailey et al. (2004) concluded that evaluation of the effects of a test 

chemical on liver weight are optimally analyzed using organ-to-body weight ratios. Therefore, the analysis of 

liver weight here focuses on relative weight data where study authors reported both relative and absolute 

weights, although both relative and absolute data are summarized in the evidence table (Table 1-12).

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1 reported in studies with more limited histopathological analyses, and were not confirmed in the 

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37

studies with more complete histopathologic examination and longer exposure durations (Lish et al., 

1984; Levine et al., 1983b). For example, the incidence of liver portal inflammation was increased 

in female rats, but not male rats, exposed to 40 mg/kg-day in the diet for 90 days (Cholakis et al., 

1980). There was an increase in the incidence of mild liver microgranulomas in female mice only

(Cholakis et al., 1980) and karyomegaly of hepatocytes in male mice only exposed to 

320 mg/kg-day RDX in the diet for 90 days (Cholakis et al., 1980). Because both the rat and mouse 

studies by Cholakis et al. (1980) used relatively small group sizes (n = 10/sex/group) and provided 

histopathologic findings for the control and high-dose groups only, less weight is placed on these 

findings than on those from the 2-year bioassays. It should be noted that exposure to HMX, the 

primary contaminant in several of the RDX studies, was associated with histopathological changes 

in the livers of male rats fed doses ≥450 mg/kg-day for 13 weeks. However, similar findings were 

not observed in the RDX studies, where the doses of RDX employed in the studies would have 

resulted in HMX exposures of ≤60 mg/kg-day. The contribution of HMX exposure to the overall 

liver findings in the studies of RDX toxicity is therefore expected to be negligible. 

Clinical chemistry parameters, including serum ALT, AST, and ALP, showed no treatmentrelated changes indicative of liver toxicity. Statistically significant changes in these parameters in 

some subchronic and chronic toxicity studies in rats and mice were relatively small (generally 

<50% of the control mean), were not dose-related in most instances, and showed no consistent 

pattern of change between sexes or across studies.

Some subchronic and chronic oral toxicity studies in rats and mice reported dose-related 

changes in serum cholesterol and triglyceride levels; however, these changes were not consistently 

observed in males and females within the same study, and patterns of changes were not consistent 

across studies. Specifically, serum triglyceride levels were elevated (up to 41%) in female B6C3F1

mice exposed to RDX in the diet for 2 years, although increases were not dose-related (Lish et al., 

1984); male mice in the same study did not show a similar increase in triglycerides. In contrast, 

serum triglycerides showed dose-related decreases in male and female F344 rats (50−62% at the 

high doses) in a subchronic oral (dietary) study (Levine et al., 1990; Levine et al., 1981a, b). In a 

chronic toxicity study by the same investigators (Levine et al., 1983b), serum triglyceride levels 

were generally decreased in male and female rats (52 and 51%, respectively, at the highest dose of 

40 mg/kg-day); however, triglyceride levels across the four dose groups in this study did not show 

a dose-related response. 

Serum cholesterol levels showed a dose-related increase (38% at the high dose of 

100 mg/kg-day) in female B6C3F1 mice exposed to RDX in the diet for 2 years (Lish et al., 1984); 

however, changes in cholesterol in male mice in the same study were not dose related. Changes in 

serum cholesterol in male and female F344 rats exposed to RDX in the diet for 2 years at doses up 

to 40 mg/kg-day (Levine et al., 1983b), in rats exposed to RDX by gavage for 90 days at doses up to 

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1 15 mg/kg-day (Crouse et al., 2006), and in monkeys exposed to RDX in the diet for 90 days (Martin 

2 

3 

4 

and Hart, 1974) were relatively small (within 38% of control mean) and were not dose related.

Table 1-11. Evidence pertaining to liver effects in humans

Reference and study design Results

Hathaway and Buck (1977) (United States)

Cross-sectional study, 2,022 workers, 

1,491 participated (74% response rate). 

Analysis group: limited to whites; 

69 exposed to RDX alone and 24 exposed 

to RDX and HMX; 338 not exposed to RDX, 

HMX, or TNT.

Exposure measures: Exposure 

determination based on job title and 

industrial hygiene evaluation. Exposed 

subjects assigned to two groups: <LOD or 

≥0.01 mg/m3 (mean for employees with 

exposures ≥LOD: 0.28 mg/m3

).

Effect measures: Liver function tests.

Analysis: Types of statistical tests were not 

reported (assumed to be t-tests for 

comparison of means and χ2 tests for 

comparison of proportions). 

Mean laboratory values of liver enzymes in men (mean; standard 

deviation not reported)

Test

Referent

(n = 237)

RDX exposed‡

Undetected (<LOD)

(n = 22)

>0.01 mg/m3

(n = 45)

LDH 173 191 174

ALP 82 78 80

ALA (SGOT) 22 25 21

AST (SGPT) 21 26 18

Bilirubin 0.5 0.4 0.4

‡

Includes both workers exposed to RDX alone and RDX and HMX.

No differences were statistically significant as reported by study 

authors. Similar results in women.

Liver function tests in men (prevalence of abnormally elevated

values)

Test

(abnormal 

range) Referent

RDX exposed‡

Undetected (<LOD) >0.01 mg/m3

LDH (>250) 2/237 1/22 0/45

ALP (>1.5) 34/237 1/22 6/45

AST (SGOT) 

(>35)

20/237 4/22 2/45

ALT (SGPT) 

(>35)

15/237 2/22 0/45

Bilirubin 

(>1.0)

5/237 1/22 1/45

‡

Includes both workers exposed to RDX alone and RDX and HMX.

No differences were statistically significant as reported by study 

authors. Similar results in women.

LDH = lactate dehydrogenase; SGOT = glutamic oxaloacetic transaminase; SGPT = glutamic pyruvic transaminase

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1 Table 1-12. Evidence pertaining to liver effects in animals

Reference and study design Results

Liver weight

Lish et al. (1984)

Mice, B6C3F1, 85/sex/group; interim 

sacrifices (10/sex/group) at 6 and 

12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles 

<66 μm

0, 1.5, 7.0, 35, or 175/100 mg/kg-d 

(high dose reduced to 100 mg/kg-d in 

wk 11 due to excessive mortality)

Diet

2 yrs

Doses 0 1.5 7.0 35 175/100

Absolute liver weight at 104 wks (percent change compared to control)

M 0% 28%* 11% 12% 35%*

F 0% 7% 7% 15% 18%*

Relative liver weight at 104 wks (percent change compared to control)

M 0% 32%* 12% 14% 46%*

F 0% 6% 8% 18% 45%*

Note: Percent change in liver weights of male and female mice was reduced 

in all dose groups when mice with liver tumors were removed from the 

analysis, suggesting no real effect on liver weight.

Hart (1976)

Rats, Sprague-Dawley, 100/sex/group

Purity and particle size not specified

0, 1.0, 3.1, or 10 mg/kg-d 

Diet

2 yrs

Doses 0 1.0 3.1 10

Absolute liver weight (percent change compared to control)

M 0% −6% −6% −6%

F 0% 7% −11% 1%

Relative liver weight (percent change compared to control)

M 0% −5% −2% −3%

F 0% 17% −2% 13%

Levine et al. (1983b)

Rats, F344, 75/sex/group; interim 

sacrifices (10/sex/group) at 6 and 

12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of 

particles <66 μm

0, 0.3, 1.5, 8.0, or 40 mg/kg-d 

Diet

2 yrs

Doses 0 0.3 1.5 8.0 40

Absolute liver weight at 105 wks (percent change compared to control)

M 0% 3% −7% 1% −8%

F 0% 1% −4% 3% 0%

Relative liver weight at 105 wks (percent change compared to control)

M 0% 1% 0% 2% 11%

F 0% 1% −2% 6% 18%*

Cholakis et al. (1980)

Mice, B6C3F1, 10−12/sex/group

88.6% pure, with 9% HMX and 2.2% 

water as contaminants; ~200 μm 

particle size

Experiment 1: 0, 10, 14, 20, 28, or 

40 mg/kg-d 

Diet

13 wks

Doses 0 10 14 20 28 40

Absolute liver weight (percent change compared to control)

M 0% − − − −6% −5%

F 0% − − − −4% −1%

Relative liver weight (percent change compared to control)

M 0% − − − −4% −4%

F 0% − − − −6% 1%

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Reference and study design Results

Experiment 2: 0, 40, 60, or 80 mg/kg-d 

for 2 wks followed by 0, 320, 160, or 

80 mg/kg-d (TWA doses of 0, 79.6, 

147.8, or 256.7 mg/kg-d for males and 

0, 82.4, 136.3, or 276.4 mg/kg-d for 

females)a

Diet

13 wks

Doses 0 80 160 320

Absolute liver weight (percent change compared to control)

M 0% 2% 12% 26%*

F 0% 4% 9% 29%*

Relative liver weight (percent change compared to control)

M 0% 0% 9% 25%*

F 0% 4% 4% 22%*

Cholakis et al. (1980)

Rats, F344, 10/sex/group

88.6% pure, with 9% HMX and 2.2% 

water as contaminants; ~200 μm 

particle size

0, 10, 14, 20, 28, or 40 mg/kg-d

Diet

13 wks

Doses 0 10 14 20 28 40

Absolute liver weight (percent change compared to control)

M 0% − − − −2% −5%

F 0% − − − 6% 4%

Relative liver weight (percent change compared to control)

M 0% − − − 2% 3%

F 0% − − − 10% 11%

Cholakis et al. (1980)

Rats, CD, two-generation study; F0: 

22/sex/group; F1: 26/sex/group; 

F2: 10/sex/group 

88.6% pure, with 9% HMX and 2.2% 

water as contaminants; ~200 μm 

particle size

F0 and F1 parental animals: 0, 5, 16, or 

50 mg/kg-d

Diet

F0 exposure: 13 wks pre-mating, and 

during mating, gestation, and lactation 

of F1; F1 exposure: 13 wks after 

weaning, and during mating, gestation, 

and lactation of F2; F2 exposure: until 

weaning

Doses 0 5 16 50

Absolute liver weight (percent change compared to control)

M 0% 7% −16% −

F 0% 0% −14% −

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d

Gavage

13 wks

Doses 0 4 8 10 12 15

Absolute liver weight (percent change compared to control)

M 0% −6% −9% 0% 7% 5%

F 0% 1% 7% 18%* 15% 28%*

Relative liver weight (percent change compared to control)

M 0% 0% −1% 2% 5% 2%

F 0% 1% −2% 2% −3% 2%

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Reference and study design Results

(Levine et al. (1990); Levine et al. 

(1981a), 1981b))

b

Rats, F344, 3−4 wks old; 10/sex/group; 

30/sex/group for controls

84.7 ± 4.7% purity, ~10% HMX, median 

particle diameter 20 μm, ~90% of 

particles ≤66 μm

0, 10, 30, 100, 300, or 600 mg/kg-d

Diet

13 wks

Doses 0 10 30 100 300 600

Absolute liver weight (percent change compared to control)

M 0% 5% −1% −2% − −

F 0% 2% 4% 16%* − −

Relative liver weight (percent change compared to control)

M 0% 9% 6% 20% − −

F 0% 3% 5% 19%* − −

Hart (1974)c

Dogs, Beagle, 3/sex/group

Pre-mix with ground dog chow 

containing 20 mg RDX/g-chow, 60 g 

dog food; purity and particle size not 

specified

0, 0.1, 1, or 10 mg/kg-d

Diet

13 wks

Doses 0 0.1 1 10

Absolute liver weight (percent change compared to control)

M 0% − − 53%

F 0% − − 3%

Martin and Hart (1974)c

Monkeys, Cynomolgus or Rhesus d, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d

Gavage

13 wks

Doses 0 0.1 1 10

Absolute liver weight (percent change compared to control)

M + F 0% 2% 6% 16%

Histopathological lesions

Lish et al. (1984)

Mice, B6C3F1, 85/sex/group; interim 

sacrifices (10/sex/group) at 6 and 

12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles 

<66 μm

0, 1.5, 7.0, 35, or 175/100 mg/kg-d 

(high dose reduced to 100 mg/kg-d in 

wk 11 due to excessive mortality)

Diet

2 yrs

Histopathological lesions in liver other than adenomas and carcinomas 

were not significantly different compared to controls, as reported by study 

authors.

Hart (1976)

Rats, Sprague-Dawley, 100/sex/group

Purity and particle size not specified

0, 1.0, 3.1, or 10 mg/kg-d

Diet

2 yrs

Histopathological examination performed only for controls and 10 mg/kg-d 

rats; no significant differences compared to controls were reported by 

study authors. 

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Reference and study design Results

Levine et al. (1983b)

Rats, F344, 3−4 wks old; 75/sex/group; 

interim sacrifices (10/sex/group) at 

6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles 

<66 μm

0, 0.3, 1.5, 8.0, or 40 mg/kg-d

Diet

2 yrs

Doses 0 0.3 1.5 8.0 40

Microgranulomas (incidence)

M 0/38 0/36 0/25 0/29 0/4

F 10/43 19/45 12/42 17/41 4/28

Cholakis et al. (1980)

Mice, B6C3F1, 10−12/sex/group 

88.6% pure, with 9% HMX and 2.2% 

water as contaminants; ~200 μm 

particle size

0, 80, 60, or 40 mg/kg-d for 2 wks 

followed by 0, 80, 160, or 320 mg/kg-d 

(TWA doses of 0, 79.6, 147.8, or 

256.7 mg/kg-d for males and 0, 82.4, 

136.3, or 276.4 mg/kg-d for females)a

Diet

13 wks

Doses 0 80 160 320

Liver microgranulomas; mild (incidence)

M 2/10 − − 1/9

F 2/11 − − 7/11*

Increased karyomegaly of hepatocytes (incidence)

M 0/10 − − 5/9*

F − − − −

Cholakis et al. (1980)

Rats, F344, 10/sex/group

88.6% pure, with 9% HMX and 2.2% 

water as contaminants; ~200 μm 

particle size

0, 10, 14, 20, 28, or 40 mg/kg-d

Diet

13 wks

Doses 0 10 14 20 28 40

Liver granulomas; mild (incidence)

M 0/10 − − − − 1/10

F − − − − − −

Liver portal inflammation (incidence)

M 2/10 − − − − 3/10

F 1/10 − − − − 7/10

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d

Gavage

13 wks

Histopathology examination of the 15 mg/kg-d group showed one male 

with mild liver congestion and one female with a moderate-sized focus of 

basophilic cytoplasmic alteration; neither finding was attributed by study 

authors to RDX treatment.

(Levine et al. (1990); Levine et al. 

(1981a), 1981b))

b

Rats, F344, 10/sex/group; 30/sex for 

control

84.7 ± 4.7% purity, ~10% HMX, median 

particle diameter 20 μm, ~90% of 

particles ≤66 μm

0, 10, 30, 100, 300, or 600 mg/kg-d

Diet

13 wks

Histopathological examination of liver did not reveal any significant 

differences compared to controls, as reported by study authors. No 

histopathology findings available for the 300 or 600 mg/kg-d dose groups 

because all rats in these groups died before the 13-wk necropsy.

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Reference and study design Results

Hart (1974)

Dogs, Beagle, 3/sex/group

Pre-mix with ground dog chow 

containing 20 mg RDX/g-chow, 60 g 

dog food; purity and particle size not 

specified

0, 0.1, 1, or 10 mg/kg-d

Diet

13 wks

Histopathological examination performed only for controls and 10 mg/kg-d 

dogs; no significant differences compared to controls were reported.

Martin and Hart (1974)

Monkeys, Cynomolgus or Rhesusd, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d

Gavage

13 wks

An increase in the amount of iron-positive material in liver cord cytoplasm 

was reported in monkeys treated with 10 mg/kg-d RDX, which the study 

authors considered to be of uncertain toxicological significance. Because 

iron-positive stain was present in controls and no further characterization 

of the staining was provided in the study report, the toxicological 

significance of this finding could not be determined.

Serum chemistry

Lish et al. (1984)

Mice, B6C3F1, 85/sex/group; interim 

sacrifices (10/sex/group) at 6 and 

12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles 

<66 μm

0, 1.5, 7.0, 35, or 175/100 mg/kg-d 

(high dose reduced to 100 mg/kg-d in 

wk 11 due to excessive mortality)

Diet

2 yrs

Doses 0 1.5 7.0 35 175/100

Serum cholesterol at 105 wks (percent change compared to control)

M 0% 11% −11% 5% 39%

F 0% 5% 15% 25% 38%

Serum triglycerides at 105 wks (percent change compared to control)

M 0% 21% −20% 10% −25%

F 0% 34% 28% 41% 28%

Levine et al. (1983b)

Rats, F344, 75/sex/group; interim 

sacrifices (10/sex/group) at 6 and 

12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles 

<66 μm

0, 0.3, 1.5, 8.0, or 40 mg/kg-d

Diet

2 yrs

Doses 0 0.3 1.5 8.0 40

Serum cholesterol at 104 wks (percent change compared to control)

M 0% 15% 38% 19% −6%

F 0% 6% 3% −7% −9%

Serum triglycerides at 104 wks (percent change compared to control)

M 0% 14% −15% −12% −52%

F 0% 18% 5% −42% −51%*

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d

Gavage

13 wks

Doses 0 4 8 10 12 15

Serum cholesterol (percent change compared to control)

M 0% −3% −10%* −16%* −18%* −11%*

F 0% −1% −8% −4% −4% −1%

Serum triglycerides (percent change compared to control)

M 0% 1% 1% −7% −2% −19%

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Reference and study design Results

F 0% −16% −21% 7% −37% 18%

(Levine et al. (1990); Levine et al. 

(1981a), 1981b))

b

Rats, F344, 10/sex/group; 30/sex for 

control

84.7 ± 4.7% purity, ~10% HMX, median 

particle diameter 20 μm, ~90% of 

particles ≤66 μm

0, 10, 30, 100, 300, or 600 mg/kg-d 

Diet

13 wks

Doses 0 10 30 100 300 600

Serum triglyceride levels (percent change compared to control)

M 0% −14% −34% −62%* − −

F 0% −12% −29% −50%* − −

Martin and Hart (1974)

Monkeys, Cynomolgus or Rhesusd, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d 

Gavage

13 wks

Serum biochemistry analysis revealed scattered deviations, but study 

authors indicated they appear to have no toxicological significance.

Doses 0 0.1 1 10

Serum cholesterol (percent change compared to control)

M 0% −17% −2% −7%

F 0% 7% 7% 7%

1 

2 

3 

4 

5 

6 

7 

8 

9 

10

11

12

13

14

15

*Statistically significant (p < 0.05) based on analysis by study authors. a

Doses were calculated by the study authors.

b

Levine et al. (1981a) is a laboratory report of a 13-week study of RDX in F344 rats; two subsequently published 

papers (Levine et al., 1990; Levine et al., 1981b) present subsets of the data provided in the full laboratory report. c

Liver weight data from the Hart (1974) and Martin and Hart (1974) studies were considered less informative than 

other studies. Hart (1974) reported organ weight data for high-dose dogs (3/sex/group) only, and the liver weights 

from Martin and Hart (1974) were highly variable across monkeys (e.g., liver weights for the control animals 

ranged from 46 to 141 g). Therefore, liver weight data from these two studies were not presented in the 

exposure-response array for liver effects (Figure 1-5). d

The species of monkey used in this study was inconsistently reported in the study as either Cynomolgus (in the 

methods section) or Rhesus (in the summary).

Note: A dash (“‒”) indicates that the study authors did not measure or report a value for that dose group.

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1 Figure 1-5. Exposure response array of liver effects following oral exposure.

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Integration of Liver Effects

1 There is limited evidence from human studies and from studies in experimental animals 

2 

3 

4 

5 

6 

7 

8 

9 

10

11

12

13

14

15

16

17

18

19

20

21

22

23

24

25

26

27

28

29

30

31

32

33

34

that RDX may affect the liver. The observation of short-term elevations of serum liver enzymes in 

several human case reports of individuals who ingested unknown amounts of RDX suggests that 

RDX might target the liver; however, serum liver enzymes were not elevated in a small crosssectional study of munition plant workers exposed to RDX. In experimental animals, dose-related 

increases in liver weight were observed in some studies following subchronic oral exposure, but 

liver weight changes were not consistent across sexes within a study or across different studies. 

Changes in serum chemistry were not consistent across studies and the magnitude of change 

relative to concurrent controls was not indicative of liver damage. Nonneoplastic histopathologic 

lesions of the liver were also not consistently associated with RDX exposure. At this time, no 

conclusions are drawn regarding liver effects as a human hazard of RDX exposure.

1.2.5. Carcinogenicity

The relationship between exposure to RDX and cancer in human populations has not been 

investigated. The carcinogenicity of RDX has been examined in one oral chronic/carcinogenicity 

bioassay in mice (Lish et al., 1984) and two bioassays in rats (Levine et al., 1983b; Hart, 1976). The 

2-year studies by Lish et al. (1984) and Levine et al. (1983b) included comprehensive 

histopathological examination of major organs, multiple dose groups and a control, and 

>50 animals/dose group (plus additional interim sacrifice groups). In both studies, the maximum 

tolerated dose was reached or exceeded in high-dose animals (based on decreased terminal body 

weight in high-dose male and female mice of 5 and 19%, respectively, and decreased survival in 

male and female rats by approximately 50 and 25%, respectively, compared to the control). The

earlier Hart (1976) study is largely limited by the lack of characterization of the test material and 

the pathology examination in control and high-dose groups only. A temperature spike in the animal 

rooms on study day 76 resulted in significant mortality across all dose groups and control animals; 

however, there were still >80 rats/sex/group after the overheating incident and 

≥50 rats/sex/group at termination, and it seems unlikely that the mortality associated with the 

temperature spike would have affected a tumor response in the rats. A summary of the evidence 

for liver and lung tumors in experimental animals from these three bioassays is provided in 

Tables 1-13 and 1-14. 

Liver Tumors

An increased incidence of liver tumors was observed in one chronic mouse study (Lish et al., 

1984) and one of two chronic rat studies (Levine et al., 1983b). Incidences of hepatocellular tumors

are presented in Table 1-13 and discussed in further detail below.

The incidence of hepatocellular carcinomas and the combined incidence of hepatocellular 

adenomas or carcinomas showed a statistically significant positive trend with RDX dose in female, 

but not male, B6C3F1 mice as compared to concurrent controls in a 2-year dietary study (Lish et al., 

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1 1984). In female B6C3F1 mice, Lish et al. (1984) observed that the liver tumor incidence in the 

2 

3 

4 

5 

6 

7 

8 

9 

10

11

12

13

14

15

16

17

18

19

20

21

22

23

24

25

26

27

28

concurrent female control mice was relatively low (1/65), and significantly lower than the 

incidence from historical controls (historical incidence data not provided by study authors). The 

study authors also compared liver tumor incidence in RDX-exposed female mice to mean historical 

control incidence for female mice of the same strain from National Toxicology Program (NTP) 

studies conducted during the same time period (147/1,781 or 8%; range: 0−20%) (Haseman et al., 

1985).11 The combined incidence of hepatocellular adenomas or carcinomas in female mice at RDX 

doses ≥35 mg/kg-day (19% at both doses) was statistically significantly elevated when statistical 

analysis was performed using NTP historical control data; limitations associated with comparisons 

to historical control data originating from a different laboratory are acknowledged given crossstudy differences in diet, laboratory, pathological evaluation, and animal provider.

A Pathology Working Group (PWG) with substantial participation by NTP pathologists

reviewed the slides of female mouse liver lesions from the Lish et al. (1984) study (Parker et al., 

2006; Parker, 2001). Some malignant tumors were downgraded to benign status and several 

lesions initially characterized as adenomas were changed to non-neoplastic lesions based on more 

recent diagnostic criteria used by the PWG (Harada et al., 1999). There remained a statistically 

significant positive trend in the combined incidence of hepatocellular adenomas or carcinomas, 

consistent with the original findings of Lish et al. (1984). Because the PWG analysis reflects more 

recent histopathological criteria for the grading of tumors, the incidence of hepatocellular 

adenomas or carcinomas as reported by Parker et al. (2006) were considered the more reliable 

measure of liver tumor response in female mice from the Lish et al. (1984) bioassay. 

In male mice from the Lish et al. (1984) study, the incidences of hepatocellular carcinomas 

in treated groups were higher than in the control, and the combined incidences of hepatocellular 

adenomas or carcinomas of male mice were higher in three of four treated groups than in the 

control; however, there were no statistically significant trends in either case. The incidences of 

liver carcinoma in control (21%) and treated groups of male mice (22−33%) were generally within

the range for the same mouse strain reported by NTP (8−32%) (Haseman et al., 1985). Similarly, 

the combined incidences of liver adenoma or carcinoma in control (32%) and treated groups 

11Comparison of control incidences of hepatocellular adenomas or carcinomas between Lish et al. (1984) and 

Haseman et al. (1985) must be interpreted with caution because of cross-study differences in labs, diets, and 

sources of animals. Specifically, the labs used by NTP and analyzed by Haseman et al. (1985) did not include 

the lab contracted to perform the Lish et al. (1984) study, and it is not clear if the diet used in the Lish et al. 

(1984) study was included in the diets reported in the NTP studies. Further, the NTP studies included three 

different suppliers of mice; one supplier was also used in the Lish et al. (1984) study. EPA Guidelines for 

Carcinogenic Risk Assessment (U.S. EPA, 2005a) also note that, unless the tumor is rare, the standard for 

determining statistical significance of tumor incidence is a comparison of dosed animals with the concurrent 

controls.

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1 (27−48%) were within the range for the same mouse strain reported by NTP (14−58%) (Haseman 

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et al., 1985).12 The PWG did not re-analyze liver tumor slides from male mice. 

A statistically significant positive trend with dose was observed in the incidence of 

hepatocellular carcinomas in male, but not female, F344 rats exposed to RDX in the diet for 2 years 

(Levine et al., 1983b). In the Levine et al. (1983b) study, there were only a few tumors observed in 

the exposed groups of male rats (0/55, 0/52, 2/55, 2/31) relative to the control (1/55), and 

inferences made from such a sparse response are uncertain. Because liver tumors are rare tumors 

in the rat13, some perspective is obtained by considering historical control data. In a paper 

published concurrently with the Levine et al. (1983b) study, NTP reported an incidence of liver 

carcinomas in untreated control male F344 rats of 0.7% (12/1,719; range: 0−2%) (Haseman et al., 

1985). In Levine et al. (1983b), the incidence of liver carcinomas in control male rats (1/55 or 

1.8%) was at the upper end of this NTP range, and the incidence in RDX-treated male F344 rats in 

the highest two dose groups (3.6 and 6.4%) exceeded the NTP historical control range. Using 

incidence data from NTP historical controls, the trend for carcinoma in the RDX-treated F344 rats 

was statistically significant (p-value = 0.003; one-sided exact Cochrane-Armitage trend test). It 

should be noted that although the NTP historical controls (Haseman et al., 1985) are comparable 

with Levine et al. (1983b) in terms of the time period, they may not be directly comparable in terms 

of diet, laboratory, pathological evaluation, and animal provider. However, other historical control 

datasets from male F344 rats, both recent and of the time period of the Levine study, indicate 

similar low incidences of liver carcinomas (0.36%, (NTP, 2009); 0.31%, (Maita et al., 1987)). In the 

Levine et al. (1983b) study, the mortality in the highest dose group is substantially higher than in 

the other dose groups during the second year leading to uncertainty in the true cancer incidence in 

the high dose group. It was not possible to estimate mortality-adjusted incidences because no timeto-death information was available. 

Nonmalignant liver tumors (neoplastic nodules) in F344 male rats in NTP historical 

controls were reported more frequently than carcinomas, with an average incidence of 3.5% 

(61/1,719; range: 0−12%) (Haseman et al., 1985); Levine et al. (1983b) reported an incidence of 

neoplastic nodules of 7.3% in their control male rats, consistent with the NTP historical control 

data, and a decline in incidence with increasing RDX exposure. The combined incidence of liver 

neoplastic nodules or carcinomas did not show a significant trend with dose. 

In a second 2-year dietary study in the rat study using a different strain (Sprague-Dawley), 

the combined incidence of hepatocellular adenomas or carcinomas was not increased with dose in 

rats of either sex at doses up to 10 mg/kg-day (Hart, 1976). However, interpretation of results 

12Ibid. 13NTP historical control data for hepatocellular carcinomas F344 rats as reported in Haseman et al. (1985): 

12/1,719 (0.7%) in males; 3/1,766 (0.17%) in females. Historical control data for Charles River SpragueDawley rats as reported in Chandra et al. (1992): 6/1,340 (0.45%) in males; 1/1,329 (0.08%) in females. 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 from this study is limited by the comparatively lower doses employed in the study, and the 

2 

3 

recording of effects only at the control and high dose groups. 

Table 1-13. Liver tumors observed in chronic animal bioassays 

Reference and study design Resultsa

Lish et al. (1984)

Mice, B6C3F1, 85/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles <66 μm

0, 1.5, 7.0, 35, or 175/100 mg/kg-d (high 

dose reduced to 100 mg/kg-d in wk 11 

due to excessive mortality)

Diet

2 yrs

Doses 0 1.5 7.0 35 175/100b

Hepatocellular adenomas (incidence)c

M 8/63

(12.7)

6/60

(10.0)

1/62*

(1.6)

7/59

(11.9)

7/27

(25.9)

F 1/65

(1.5)

1/62

(1.6)

6/64

(9.4)

6/64

(9.4)

3/31

(9.7)

Hepatocellular carcinomas (incidence)c

M 13/63

(20.6)

20/60

(33.3)

16/62

(25.8)

18/59

(30.5)

6/27

(22.2)

F 0/65

(0.0)

4/62

(6.5)

3/64

(4.7)

6/64

(9.4)

3/31d

(9.7)

Hepatocellular adenoma or carcinoma combined (incidence)c

M 20/63

(31.7)

26/60

(43.3)

17/62

(27.4)

25/59

(42.4)

13/27

(48.1)

F 1/65

(1.5)

5/62

(8.1)

9/64*

(14.1)

12/64*

(18.8)

6/31*d

(19.4)

PWG reanalysis of liver lesion slides from female mice (Parker et al., 

2006; Parker, 2001).

e

Doses 0 1.5 7.0 35 175/100b

Hepatocellular adenomas (incidence)c

F 1/67

(1.5)

3/62

(4.8)

2/63

(3.2)

8/64

(12.5)

2/31

(6.5)

Hepatocellular carcinomas (incidence)c

F 0/67

(0.0)

1/62

(1.6)

3/63

(4.8)

2/64

(3.1)

2/31

(6.5)

Hepatocellular adenoma or carcinoma combined (incidence)c

F 1/67

(1.5)

4/62

(6.5)

5/63

(7.9)

10/64

(15.6)

4/31d

(12.9)

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Reference and study design Resultsa

Hart (1976)

Rats, Sprague-Dawley, 100/sex/group

Purity and particle size not specified

0, 1.0, 3.1, or 10 mg/kg-d 

Diet

2 yrs

Doses 0 1.0 3.1 10

Neoplastic nodules (incidence)c

M 0/82 − − 3/77

F 1/72 − − 1/81

Hepatocellular carcinomas (incidence)c

M 1/82 − − 1/77

F 1/72 − − 1/81f

Neoplastic nodules or hepatocellular carcinomas combined

(incidence)c 

M 1/82 − − 4/77

F 2/72 − − 2/81

Levine et al. (1983b)

Rats, F344, 75/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles <66 μm

0, 0.3, 1.5, 8.0, or 40 mg/kg-d 

Diet

2 yrs

Doses 0 0.3 1.5 8.0 40

Neoplastic nodules (incidence)c

M 4/55

(7.3)

3/55

(5.5)

0/52

(0.0)

2/55

(3.6)

1/31

(3.2)

F 3/53

(5.6)

1/55

(1.8)

1/54

(1.9)

0/55

(0.0)

4/48

(8.3)

Hepatocellular carcinomas (incidence)c

M 1/55

(1.8)

0/55

(0.0)

0/52

(0.0)

2/55

(3.6)

2/31d

(6.5)

F 0/53

(0.0)

1/55

(1.8)

0/54

(0.0)

0/55

(0.0)

0/48

(0.0)

Neoplastic nodules or hepatocellular carcinomas combined

(incidence)c

M 5/55

(9.1)

3/55

(5.5)

0/52

(0.0)

4/55

(7.3)

3/31

(9.7)

F 3/53

(5.6)

2/55

(3.6)

1/54

(1.9)

0/55

(0.0)

4/48

(8.3)

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*Statistically significant difference compared to the control group (p < 0.05), identified by the authors. a

Selected percent incidences are provided in parentheses below the incidences to help illustrate patterns in the 

responses.

b

The lower dose of 100 mg/kg-day was started in week 11, resulting in a duration-weighted average dose of 

107 mg/kg-day. c

The incidences reflect the animals surviving to month 12. d

Statistically significant trend (p < 0.05) was identified using a one-sided Cochran-Armitage trend tests performed 

by EPA.

e

The numbers of animals at risk (i.e., the denominators) in the control group (n = 67) and 7 mg/kg-day dose group 

(n = 63) as reported in the PWG reanalysis (Parker et al., 2006; Parker, 2001) differed from the numbers reported 

in the original study by Lish et al. (1984) (n = 65 and 64, respectively). Further investigation of these differences 

by the U.S. Army (sponsor of the mouse bioassay and subsequent PWG reevaluation) was unable to resolve the 

This document is a draft for review purposes only and does not constitute Agency policy.

1-64 DRAFT—DO NOT CITE OR QUOTE
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1 discrepancy (email to Louis D’Amico, U.S. EPA, from Mark Johnson, U.S. Army Public Health Command, February 

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13, 2015).

f

Hart (1976) distinguishes the single high-dose carcinoma in the liver from a hepatocellular carcinoma; the 

incidence of hepatocellular carcinomas in this dose group is shown as 0/81 (p. 119 of the publication).

Note: A dash (“‒”) indicates that the study authors did not measure or report a value for that dose group.

Lung Tumors

Lung tumors were observed in female and male B6C3F1 mice exposed to RDX in the diet for 

2 years (Lish et al., 1984) (see Table 1-14). Incidence of alveolar/bronchiolar carcinomas and the 

combined incidence of alveolar/bronchiolar adenomas or carcinomas showed a statistically 

significant positive trend (one-sided p-values of 0.016 and 0.009, respectively, for the CochranArmitage trend test) in female mice. Incidence of alveolar/bronchiolar carcinomas in male mice 

showed a statistically significant positive trend (p-value = 0.015; one-sided Cochran-Armitage trend 

test). However, the combined incidence of adenomas and carcinomas was not elevated in male 

mice. In such a case, NTP policy recommends analyzing the tumors both separately and in 

combination (McConnell et al., 1986). This recommendation arose out of concern that combining 

benign and malignant neoplasms can result in a false negative if the chemical shows a statistically 

significant increase in malignant tumors without an increase in the combined incidence. In an 

addendum to the study report that included results of additional examination and sectioning of 

lung specimens from the mid-dose groups in the mouse study, Lish et al. (1984) noted an increase 

in the combined incidences of primary pulmonary neoplasms in males of all dose groups and in 

females in the 7.0, 35, and 175/100 mg/kg-day dose groups, but regarded these neoplasms as 

random and not biologically significant (rationale for this conclusion not provided). 

Bioassays in rats provide no evidence of an association between RDX exposure and 

induction of lung tumors. The incidence of alveolar/bronchiolar adenomas or carcinomas was not 

increased in either sex of Sprague-Dawley rats exposed chronically to RDX at doses up to 

10 mg/kg-day (Hart, 1976) or in F344 rats of either sex exposed chronically to RDX at doses up to 

40 mg/kg-day (Levine et al., 1983b). Alveolar/bronchiolar carcinomas are rare tumors in both 

species of rats, male or female (Chandra et al., 1992; Haseman et al., 1985). 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 Table 1-14. Lung tumors observed in chronic animal bioassays 

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3 

4 

5 

6 

7 

Reference and study design Resultsa

Lish et al. (1984)

Mice, B6C3F1, 85/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles 

<66 μm

0, 1.5, 7.0, 35, or 175/100 mg/kg-d (high 

dose reduced to 100 mg/kg-d in wk 11 

due to excessive mortality)

Diet

2 yrs

Doses 0 1.5 7.0 35 175/100b

Alveolar/bronchiolar adenomas (incidence)c

M 6/63

(9.5)

5/60

(8.3)

5/62

(8.1)

7/59

(11.9)

1/27

(3.7)

F 4/65

(6.2)

2/62

(3.2)

5/64

(7.8)

9/64

(14.1)

3/31

(9.7)

Alveolar/bronchiolar carcinomas (incidence)c

M 3/63

(4.8)

6/60

(10.0)

3/62

(4.8)

7/59

(11.9)

5/27d

(18.5)

F 3/65

(4.6)

1/62

(1.6)

3/64

(4.7)

3/64

(4.7)

4/31d

(12.9)

Alveolar/bronchiolar adenoma or carcinoma combined (incidence)c

M 9/63

(14.3)

11/60

(18.3)

8/62

(12.9)

14/59

(23.7)

6/27

(22.2)

F 7/65

(10.8)

3/62

(4.8)

8/64

(12.5)

12/64

(18.8)

7/31d

(22.6)

Hart (1976)

Rats, Sprague-Dawley, 100/sex/group

Purity and particle size not specified

0, 1.0, 3.1, or 10 mg/kg-d 

Diet

2 yrs

Doses 0 1.0 3.1 10

Alveolar/bronchiolar adenoma (incidence)

M 2/83 − − 1/77

F 0/73 − − 0/82

No alveolar/bronciolar carcinomas reported by study authors.

Levine et al. (1983b)

Rats, F344, 75/sex/group; interim 

sacrifices (10/sex/group) at 6 and 12 mo

89.2−98.7% pure, with 3−10% HMX as 

contaminant; 83−89% of particles 

<66 μm

0, 0.3, 1.5, 8.0, or 40 mg/kg-d 

Diet

2 yrs

Doses 0 0.3 1.5 8.0 40

Alveolar/bronchiolar adenomas (incidence)c

M 1/55 0/15 1/17 0/16 1/31

F 3/53 0/7 0/8 1/10 0/48

Alveolar/bronchiolar carcinomas (incidence)c

M − − − − −

F 0/53 0/7 1/8 0/10 0/48

Alveolar/bronchiolar adenoma or carcinoma combined (incidence)c

M − − − − −

F 3/53 0/7 1/8 1/10 0/48

a

Selected percent incidences are provided in parentheses below the incidences to help illustrate patterns in the 

responses.

b

The lower dose of 100 mg/kg-day was started in week 11, resulting in a duration-weighted average dose of 

107 mg/kg-day. c

The incidences reflect the animals surviving to month 12.

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d 1 Statistically significant trend (p < 0.05) was identified using a one-sided Cochran-Armitage trend test performed by 

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EPA. 

Note: A dash (“‒”) indicates that the study authors did not measure or report a value for that dose group.

Mechanistic Evidence

There are few mechanistic data to inform a MOA determination for either liver or lung 

tumors induced by exposure to RDX. 

The available in vitro and in vivo genotoxicity assay results are largely negative for parent 

RDX or its oxidative metabolites (see Appendix C, Section C.3.2), supporting the hypothesis that 

parent RDX or its oxidative metabolites do not interact directly with deoxyribonucleic acid (DNA). 

In contrast, there are some positive genotoxicity results for the N-nitroso metabolites of RDX, 

specifically hexahydro-1-nitroso-3,5-dinitro-1,3,5-triazine (MNX) and hexahydro-1,3,5-trinitroso1,3,5-triazine (TNX). MNX and TNX have been identified from minipigs; minipigs were chosen as 

the animal model for investigation of RDX metabolism because the GI tract of pigs more closely 

resembles that of humans (Musick et al., 2010; Major et al., 2007). MNX has tested positive in some 

in vitro assays, including unscheduled DNA synthesis in primary rat hepatocytes and the mouse 

lymphoma forward mutation assay (Snodgrass, 1984), although MNX tested negative in the only in 

vivo test performed, a mouse dominant lethal mutation test (Snodgrass, 1984). MNX was not 

mutagenic in Salmonella typhimurium (strains TA98, TA100, TA1535, TA1537, and TA1538), with 

or without the addition of the S9 metabolic activating mixture (Pan et al., 2007b; Snodgrass, 1984). 

When S. typhimurium strains TA97a and TA102, strains sensitive to frame shift and oxidative DNA 

damage, were used in conjunction with elevated concentrations of the metabolizing system (S9), 

MNX and TNX were mutagenic. N-nitroso metabolites, including MNX and TNX, are generated 

anaerobically and are likely a result of bacterial transformation of parent RDX in the GI tract to 

various N-nitroso derivatives (Pan et al., 2007b). Exposure to potentially mutagenic N-nitroso 

metabolites of RDX generated in the GI tract of mice may occur in the liver (and subsequently in the 

systemic circulation) via enterohepatic circulation. However, in pigs, the N-nitroso metabolites of 

RDX have been identified only in trace amounts in urine compared to the major metabolites, 

4-nitro-2,4-diazabutanal and 4-nitro-2,4-diaza-butanamide (Major et al., 2007). Thus, the 

contribution of the N-nitroso metabolites to the overall carcinogenic potential of RDX is unclear. 

Aberrant expression of miRNAs was observed in the brains and livers of female B6C3F1

mice fed 5 mg RDX/kg in the diet for 28 days (Zhang and Pan, 2009b) (dose of 0.75−1.5 mg/kg-day 

estimated by Bannon et al. (2009b)), with several oncogenic miRNAs being upregulated, while 

several tumor-suppressing miRNAs were downregulated. However, the pattern of induction was 

not always consistent in the livers of RDX-treated mice (e.g., miR-92a was downregulated in liver 

tissue samples when it is typically upregulated in hepatocellular carcinomas) (Sweeney et al., 

2012b). miRNAs have been associated with several cancers (Wiemer, 2007; Zhang et al., 2007), but 

the utility of miRNAs as predictive of carcinogenesis has not been demonstrated (Bannon et al., 

2009b). Further, it is unknown whether or not aberrant expression of a specific miRNA (or suite of 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 miRNAs) plays a role in the MOA of RDX carcinogenicity. Microarray analysis of gene expression in 

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male Sprague-Dawley rats after exposure to a single oral (capsule) dose of RDX revealed a general 

upregulation in gene expression (predominantly genes involved in metabolism) in liver tissues 

(Bannon et al., 2009a); however, the relevance of this finding to the carcinogenicity of RDX is 

unclear. 

Sweeney et al. (2012b) hypothesized a set of MOAs for the liver tumors:

• Genotoxicity mediated by either: (1) RDX; (2) tissue-generated oxidative metabolites; or 

(3) N-nitroso metabolites generated anaerobically in the GI tract. The key events in this 

hypothesized MOA are: production of DNA damage, gene mutation, formation of neoplastic 

lesions, and promotion/progression of tumors. The largely negative results for genotoxicity 

led Sweeney et al. (2012b) to conclude that this MOA is not plausible for RDX or its 

oxidative metabolites. Although there are some positive results for the N-nitroso 

metabolites, the limited evidence to support systemic uptake and distribution of 

metabolites to the liver led Sweeney et al. (2012b) to conclude that this MOA is not 

sufficiently plausible.

• Cell proliferation. The key events in this hypothesized MOA are GI-tract generation of 

N-nitroso metabolites, absorption, distribution to the liver, cytoxocity (optional), and 

enhanced cell proliferation, leading to preneoplastic foci that progress to hepatocellular 

adenomas and carcinomas. Sweeney et al. (2012b) cited evidence of increased liver weights 

in mice as consistent with cell proliferation, but noted that increased liver weights were 

also observed in rats without proceeding to liver tumors. They considered this MOA 

“plausible, but not particularly well supported.”

In addition, there are other results that do not support a metabolite-based proliferative 

response as the MOA for carcinogenesis.

• The absence of significant liver histopathology in mice after subchronic or chronic exposure 

to RDX at doses that induced liver tumors (Lish et al., 1984; Cholakis et al., 1980) suggests 

that cellular toxicity is not a precursor to these tumors. 

• As discussed in Section 1.2.4, changes in liver weight showed no consistent pattern across 

studies or sexes, and did not correlate with tumor response.

• No studies were available that directly measured RDX-induced cell proliferation rates. 

• No information was available to rule out non-precancerous causes of liver weight increase.

In summary, the available evidence indicates that RDX is likely not mutagenic (see 

Appendix C, Section C.3.2), although anaerobically-derived N-nitroso metabolites have 

demonstrated some genotoxic potential. While these metabolites have been measured in the 

mouse (Pan et al., 2007b) and minipig (Musick et al., 2010; Major et al., 2007), they have not been 

identified in humans, and may not be the predominant metabolites of RDX. A MOA involving a 

proliferative response generated by tissue-derived oxidative metabolites of RDX has been 

proposed, but is not supported by the available data. In light of limited information on precursor 

This document is a draft for review purposes only and does not constitute Agency policy.

1-68 DRAFT—DO NOT CITE OR QUOTE
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1 events leading to the observed liver and lung tumor response in RDX-exposed rodents and lack of 

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toxicokinetic information on RDX metabolites, neither a cell proliferative MOA nor a mutagenic Nnitroso metabolite MOA is supported. Thus, the MOA leading to the increased incidence of liver and 

lungs tumors is not known. 

1.2.6. Other Noncancer Effects 

There are isolated reports of RDX inducing systemic effects in several organs/systems, 

including the eyes and the musculoskeletal, cardiovascular, immune, and GI systems. However, 

there is less evidence for these effects compared to organ systems described earlier in Section 1.2. 

Generally, evidence for toxicological effects in these organ systems was limited to human case 

reports, lacked reproduction or were not observed in other studies of similar duration in the same 

species, or lacked consistent, dose-related patterns of increasing or decreasing effect. A longer 

discussion of the evidence for each of the other noncancer effects noted above is provided in 

Appendix C.3.2. At this time, no conclusions are drawn regarding the other noncancer effects as 

human hazards of RDX exposure.

1.3. INTEGRATION AND EVALUATION

1.3.1. Effects Other Than Cancer

The majority of evidence for the health effects of RDX comes from oral toxicity studies in 

animals. The three epidemiology studies that document possible inhalation exposure are limited by 

various study design features, including inability to distinguish exposure to TNT (associated with 

liver and hematological system toxicity), inability to adequately characterize exposure levels, small 

sample sizes, and inadequate reporting. The single animal inhalation study identified in the 

literature search had deficiencies (e.g., lack of a control and incomplete exposure information) that 

precluded its inclusion in this assessment (see literature search section). 

The strongest evidence for hazards following exposure to RDX is for nervous system effects. 

Toxicity studies in multiple animal species involving chronic, subchronic, and gestational exposures

provide consistent evidence of nervous system effects following oral exposure. Effects included

dose-related increases in seizures and convulsions, as well as observations of tremors, 

hyperirritability, hyper-reactivity, and other behavioral changes (Crouse et al., 2006; Angerhofer et 

al., 1986; Levine et al., 1983b; Levine et al., 1981a, b; Cholakis et al., 1980; von Oettingen et al., 

1949). 

Human studies provide supporting evidence for RDX as a neurotoxicant and that the 

nervous system effects observed in experimental animals are plausible in, and relevant to, humans. 

A cross-sectional study described memory impairment and visual-spatial decrements in RDXexposed workers (Ma and Li, 1993), although confidence in these findings is relatively low because 

of issues with design and reporting. Several case reports provide additional evidence of 

associations between exposure to RDX (via ingestion, inhalation, and possibly dermal exposure)

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1 and seizures and convulsions (Kasuske et al., 2009; Küçükardali et al., 2003; Testud et al., 1996a; 

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Testud et al., 1996b; Woody et al., 1986 and others, see Appendix C.2). Other nervous system 

effects identified in human case reports include dizziness, headache, confusion, and 

hyperirritability. 

Additional support for an association between RDX exposure and nervous system effects 

comes from consistent evidence of neurotoxicity across taxa, including several species of wildlife 

(Quinn et al., 2013; Garcia-Reyero et al., 2011; McFarland et al., 2009; Gogal et al., 2003). The 

association between RDX and neurological effects is biologically plausible, with studies 

demonstrating a correlation between blood and brain concentrations of RDX and the time of 

seizure onset (Williams et al., 2011; Bannon et al., 2009a). Additionally, the affinity of RDX for the 

picrotoxin convulsant site of the GABAA channel suggests that the resulting disinhibition could lead 

to the onset of seizures (Williams et al., 2011). 

Induction of convulsions and seizures appears to be more strongly correlated with dose 

than with duration of exposure. However, there is some mechanistic information to suggest that 

repeated exposure to a chemical binding to the receptor convulsant site of GABAA may promote a 

state of increased neuronal activity that could increase the likelihood of subsequent neurological 

effects (Gerkin et al., 2010). It is unclear if nervous system effects progressed in severity (e.g., from 

behavioral change to seizures and convulsions) with increasing dose, as many of the studies that 

reported more subtle neurobehavioral changes did not provide detailed dose-response 

information, and the majority of studies were not designed to capture this information. 

The nervous system effects following oral exposure to RDX were observed in humans 

acutely exposed to RDX and in multiple experimental animal studies in rats, mice, monkeys, and 

dogs following exposures ranging from 10 days to 2 years in duration. Across the database, 

behavioral manifestations of seizure activity were the most consistently observed nervous system 

effect associated with RDX exposure. This most commonly included evidence of increased 

convulsions, as well as other related effects such as tremors, shaking, hyperactivity, or nervousness, 

which were generally observed at doses that were the same as or higher than doses that induced 

convulsions. Nervous system effects are a human hazard of RDX exposure and are carried forward 

for consideration for dose-response analysis. Convulsions, considered a severe adverse effect, were 

selected as a consistent and sensitive endpoint representative of nervous system effects. 

Evidence for kidney and other urogenital toxicity is more limited than evidence for

neurotoxicity. Histopathological changes in the urogenital system (suppurative prostatitis, 

medullary papillary necrosis, suppurative pyelitis, uremic mineralization, and luminal distention 

and cystitis of the urinary bladder) were reported in male rats exposed to RDX in the diet for 

2 years (Levine et al., 1983b). Similar histopathological changes of the urogenital system were not 

observed in mice, and no other rat studies of similar duration that examined the prostate were

available. As discussed earlier, among the lesions identified in the rat, the incidence of suppurative 

prostatitis is considered a marker for RDX-related urogenital effects as it is plausibly associated 

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1 with a progression of effects resulting from other kidney or bladder lesions. The plausibility of a 

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MOA that shares a common molecular initiating event (binding to the GABAA receptor convulsantsite) with the neurotoxic effects of RDX provides some support for an association between RDX 

exposure and kidney and other urogenital effects. Kidney and other urogenital system effects are a 

potential human hazard of RDX exposure and were carried forward for consideration for doseresponse analysis. Prostatitis, considered a marker for the kidney and urogenital effects, was 

selected as a sensitive endpoint representative of the urogenital system effects.

There is some evidence for male reproductive toxicity that comes from the finding of 

testicular degeneration in male B6C3F1 mice chronically exposed to RDX in the diet (Lish et al., 

1984) in the only mouse study conducted of that duration (24 months). The effect was noted by the 

study authors at both the penultimate and highest dose tested in the study. However, studies in 

different rat strains did not consistently observe testicular effects. Although the available data are

limited, given the dose-related findings of mouse testicular degeneration, there is suggestive 

evidence of male reproductive effects associated with RDX exposure; these effects were carried 

forward for consideration for dose-response analysis. Testicular degeneration, the only endpoint 

observed, was selected as the endpoint representative of male reproductive effects.

Evidence for developmental toxicity and liver toxicity was more limited than that for the 

endpoints discussed above. In animal studies, developmental effects, including offspring survival, 

growth, and morphological development, were observed only at doses associated with maternal 

mortality (Angerhofer et al., 1986; Cholakis et al., 1980). Evidence for potential hepatic effects 

comes from observations of increases (generally dose-related) in liver weight in some subchronic 

oral animal studies (Lish et al., 1984; Levine et al., 1983b; Levine et al., 1981a, b; Cholakis et al., 

1980; Hart, 1976). However, these elevations in liver weight were not consistently observed across 

studies nor were they accompanied by RDX-related histopathological changes in the liver or 

increases in serum liver enzymes. In addition, the interpretation of liver weight changes in the 

mouse bioassay by Lish et al. (1984) is complicated by the relatively high incidence of liver tumors

in this study. At this time, no conclusions are drawn regarding developmental and liver toxicity as 

human hazards of RDX exposure; these effects were not considered further for dose-response 

analysis and derivation of reference values. 

As discussed in Section 1.2, mortality is not addressed in this assessment as a hazard by 

itself, but rather in the context of nervous and urogenital system hazards. Histopathological 

changes in the urogenital system observed in male rats exposed to 40 mg/kg-day in the diet for 2 

years were considered the principal cause of treatment-related morbidity and mortality (Levine et 

al., 1983b). However, the incidence of suppurative prostatitis, considered a sensitive marker of the 

urogenital effects, was increased at doses of ≥1.5 mg/kg-day. Therefore, the mortality 

characterized as secondary to renal effects in Levine et al. (1983b) is a less sensitive endpoint (by 

more than 10-fold) than the effect that is the selected as the basis dose-response analysis (i.e., 

suppurative prostatitis).

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1 In a number of the animal studies reporting nervous system effects, unscheduled deaths 

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occurred at RDX doses as low as those that induced nervous system effects (Crouse et al., 2006; 

Angerhofer et al., 1986; Levine et al., 1983b; Levine et al., 1981a; Cholakis et al., 1980; von 

Oettingen et al., 1949). In a 90-day study that recorded nervous system effects and survival more 

thoroughly than earlier studies, Crouse et al. (2006) reported that nearly all pre-term deaths were 

preceded by neurotoxic signs such as tremors and convulsions. Convulsions did not, however, 

necessarily lead to early mortality; of the animals observed to have convulsed in the Crouse et al. 

(2006) study, approximately 75% survived to the end of the 90-day study. Most of the earlier 

studies provide a limited understanding of the association between mortality and nervous system 

effects because the frequency of clinical observations was likely insufficient to observe convulsions 

prior to death. In humans, mortality has not been reported in case reports involving workers with 

symptoms of neurotoxicity exposed to RDX during manufacture or in individuals exposed acutely as 

a result of accidental or intentional ingestion; however, survival has not been specifically evaluated 

in studies of worker populations exposed chronically to RDX. 

Regarding mortality, the preference, in general, is not to use a frank health effect as severe 

as mortality as the basis for a reference value. As noted in U.S. EPA (2002), a chemical may cause a 

variety of effects ranging from severe—such as death—to more subtle biochemical, physiological, 

or pathological changes; primary attention in assessing health risk should be given to those effects 

in the lower exposure range and/or the effects most biologically appropriate for a human health 

risk assessment. Where mortality occurs as a consequence of a chemical’s effects on a specific 

organ/system (e.g., in the case of RDX, evidence suggests some relationship between mortality and 

effects on the nervous or kidney/urogenital systems), the preference would be to develop a 

quantitative assessment based on the initial hazard and not on death. Because unscheduled deaths 

were observed with some consistency across studies and, in some studies, at doses as low as those 

associated with convulsions, two additional analyses of mortality data are presented in Chapter 2. 

In the first analysis, BMDs derived using mortality data sets are compared to the BMD used to 

derive the RfC (Section 2.1.6). In addition, the relationship between convulsions and mortality is 

not clear and raises concerns for the potential underreporting of convulsions (see Section 1.2.1). 

An analysis, described in Section 2.1.7, addresses the possibility that the analyses of convulsions 

brought forward for dose-response analysis resulted in an underestimate of the toxicity for RDX. 

1.3.2. Carcinogenicity

As presented in Section 1.2.5, dietary administration of RDX induced dose-related increases 

in the incidence of hepatocellular adenomas or carcinomas in male and female B6C3F1 mice mice 

(Parker et al., 2006; Lish et al., 1984). In the same study, RDX also induced dose-related increases 

in the incidence of alveolar/bronchiolar adenomas or carcinomas in both sexes. Some of these 

trends in liver and lung were statistically significant. In Fischer 344 rats, dietary administration of 

RDX yielded a statistically significant trend in the incidence of hepatocellular carcinomas in males, 

but not in females (Levine et al., 1983b). A 2-year dietary study in Sprague-Dawley rats was 

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1 negative in both sexes Hart (1976), although doses in this study were somewhat lower (no 

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carcinomas at doses up to 10 mg/kg-d in Hart (1976), versus hepatocellular carcinomas at 8 and 

40 mg/kg-d in the Levine et al. (1983b) study. The human studies are not informative. 

This evidence leads to consideration of two hazard descriptors under the EPA’s cancer 

guidelines (U.S. EPA, 2005a). The descriptor likely to be carcinogenic to humans is appropriate 

when the evidence is “adequate to demonstrate carcinogenic potential to humans” but does not 

support the descriptor carcinogenic to humans. One example from the cancer guidelines is “an 

agent that has tested positive in animal experiments in more than one species, sex, strain, site, or 

exposure route, with or without evidence of carcinogenicity in humans.” RDX matches the 

conditions of this example, having induced dose-related increases in tumors in two species (mouse 

and rat), in both sexes, and at two sites (liver and lung). 

Alternatively, the descriptor suggestive evidence of carcinogenic potential is appropriate 

when the evidence raises “a concern for potential carcinogenic effects in humans” but is not 

sufficient for a stronger conclusion. The hepatocellular carcinoma result in male F344 rats is based 

on a small number of tumors (two each at 8 and 40 mg/kg-day, versus one in controls), and RDX 

did not increase the incidence of carcinomas at any other site in F344 or Sprague-Dawley rats of 

either sex. 

As noted in the EPA’s cancer guidelines (U.S. EPA, 2005a), choosing a hazard descriptor 

cannot be reduced to a formula, as descriptors may be applicable to a variety of potential data sets 

and represent points along a continuum of evidence. In the case of RDX, there are plausible 

scientific arguments for more than one hazard descriptor. Overall, the considerations discussed 

above, interpreted in light of the cancer guidelines, lead to the conclusion that there is suggestive 

evidence of carcinogenic potential for RDX. Although the evidence includes dose-related tumor 

increases in two species, two sexes, and two sites, the evidence of carcinogenicity outside the 

B6C3F1 mouse is not robust, and this factor was decisive in choosing a hazard descriptor. Within 

the spectrum of results covered by the descriptor suggestive evidence, the evidence for RDX is 

strong. There are well-conducted studies that tested large numbers of animals at multiple dose 

levels, making the cancer response suitable for dose-response analysis (Section 2).

The descriptor suggestive evidence of carcinogenic potential applies to all routes of human 

exposure. Dietary administration of RDX to mice and rats induced tumors of the liver or lung, sites 

beyond the point of initial contact, and human case reports have demonstrated absorption and 

distribution of inhaled RDX into the systemic circulation. Under the cancer guidelines, this 

information provides sufficient basis to apply the cancer descriptor developed from oral studies to 

other exposure routes. 

1.3.3. Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes

Susceptibility refers to factors such as lifestage, genetics, sex, and health status that may 

predispose a group of individuals to greater response to an exposure. This greater response could 

be achieved either through differences in exposure to the chemical underlying RDX toxicokinetics 

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1 or differences in RDX toxicodynamics between susceptible and other populations. Little 

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information is available on populations that may be especially vulnerable to the toxic effects of RDX. 

Lifestage, and in particular childhood, susceptibility has not been observed in human or 

animal studies of RDX toxicity. Transfer of RDX from dam to the fetus during gestation has been 

reported, and the presence of RDX in the milk of dams administered 6 mg/kg-day by gavage has 

been documented (Hess-Ruth et al., 2007); however, reproductive and developmental toxicity 

studies generally did not identify effects in offspring at doses below those that also caused severe 

maternal toxicity (Angerhofer et al., 1986; Cholakis et al., 1980). Thus, the existing toxicity 

literature does not provide evidence of early lifestage susceptibility to RDX. 

Limited data suggest that male laboratory animals may be more susceptible to noncancer 

toxicity associated with RDX exposure. In general, male animals were more sensitive to RDX 

neurotoxicity than females (i.e., more convulsions; more hyperactive; greater brain weight 

changes). Urogenital effects have been observed in males at lower doses than in females (Levine et 

al., 1983b; Levine et al., 1981a, b; Cholakis et al., 1980), suggesting a possible sex-based difference 

in susceptibility to RDX toxicity. 

Data on the incidence of convulsions and mortality from gavage studies of RDX in the rat

provide some indication that pregnant animals may be a susceptible population. In the 

developmental toxicity study by Cholakis et al. (1980), deaths were observed in pregnant F344 rats 

only at a dose of 20 mg/kg-day, but convulsions were reported in a single rat at 2 mg/kg-day. In a 

range-finding developmental toxicity study (Angerhofer et al., 1986), mortality and convulsions 

were reported in pregnant Sprague-Dawley rats at a dose of ≥40 mg/kg-day, but not at ≤20 mg/kgday, although the relatively small group sizes in this study should be noted. In the main study by 

these investigators, convulsions were reported in pregnant rats only at 20 mg/kg-day, but one 

death (in dose groups of 40 rats) was reported at both 2 and 6 mg/kg-day (Angerhofer et al., 1986). 

In comparison, increased mortality and convulsions were reported at ≥8 mg/kg-day in a 90-day 

gavage study in F344 rats (Crouse et al., 2006). The instances of one convulsion and two deaths in 

pregnant rats in the Cholakis et al. (1980) and Angerhofer et al. (1986) studies at doses of 2 or 6 

mg/kg-day raise the possibility that pregnants animals may be more susceptible to the effects of 

RDX; however, direct comparison between the available gavage studies in pregnants and 

nonpregnant rats is uncertain because of differences in study design, including numbers of animals 

tested per group, test material characteristics, and rat strain. Overall, the available information is 

not considered sufficient to conclude that pregnant animals are a susceptible population.

There is limited evidence that CYP450 or similar enzymes are involved in the metabolism of 

RDX (Bhushan et al., 2003), indicating a potential for genetic polymorphisms in these metabolic 

enzymes to affect susceptibility to RDX. This susceptibility may also be influenced by differential 

expression of these enzymes during development. Individuals with epilepsy or other seizure 

syndromes, and in particular those that have their basis in genetic mutation to GABAA receptors, 

may represent another group that may be susceptible to RDX exposure. However, there is currently 

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1 no information to support predictions of how genetic polymorphisms or the presence of seizure 

2 syndromes may affect susceptibility to RDX exposure. 

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2.DOSE-RESPONSE ANALYSIS

2.1. ORAL REFERENCE DOSE FOR EFFECTS OTHER THAN CANCER

1 The oral reference dose (RfD, expressed in units of mg/kg-day) is defined as an estimate 

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(with uncertainty spanning perhaps an order of magnitude) of a daily exposure to the human 

population (including sensitive subgroups) that is likely to be without an appreciable risk of 

deleterious effects during a lifetime. It can be derived from a no-observed-adverse-effect level 

(NOAEL), lowest-observed-adverse-effect level (LOAEL), or the 95% lower bound on the 

benchmark dose (BMDL), with uncertainty factors (UFs) generally applied to reflect limitations of 

the data used.

2.1.1. Identification of Studies for Dose-Response Analysis of Selected Effects

As discussed in Section 1.3.1, based on findings from oral studies in experimental animals, 

nervous system effects are a human hazard of hexahydro-1,3,5-trinitro-1,3,5-triazine (RDX) 

exposure, and kidney and other urogenital system effects are a potential human hazard of RDX 

exposure. There is suggestive evidence of male reproductive effects associated with RDX exposure. 

Although animal mortality has been reported in a number of the toxicology studies conducted for 

RDX, it was not considered a hazard by itself or as the basis for the derivation of a reference 

value. Rather, the mortality evidence was evaluated in the context of that system-specific hazard

(see Sections 2.1.6 and 2.1.7 for further discussion). 

The effects selected to best represent each of the hazards (see discussion in Section 1.3.1) 

are noted below. In order to identify the stronger studies for dose-response analysis, several 

attributes of the studies reporting the endpoints selected for each hazard were reviewed (i.e., study 

size and design, relevance of the exposure paradigm, and measurement of the endpoints of 

interest). In considering the study size and design, preference was given to studies using designs 

reasonably expected to have power to detect responses of suitable magnitude. Exposure paradigms 

including a route of human environmental exposure (i.e., oral and inhalation) are preferred. When 

developing a chronic reference value, chronic or subchronic studies are preferred over studies of 

acute exposure durations. Studies with a broad exposure range and multiple exposure levels are 

preferred to the extent that they can provide information about the shape of the exposure-response 

relationship. Additionally, with respect to measurement of the endpoint, studies that can reliably 

distinguish the presence or absence (or degree of severity) of the effect are preferred. 

Human studies are generally preferred over animal studies as the basis for a reference value 

when quantitative measures of exposure are reported, and the reported effects are determined to 

be associated with exposure. The available epidemiological studies of worker populations exposed 

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1 to RDX examined the relationship between certain health endpoints and inhalation exposure; 

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however, no epidemiological studies of ingested RDX are available. Multiple case reports support 

the identification of hazards associated with RDX exposure but are inadequate for dose-response 

analysis because they do not yield incidence estimates, exposure durations are short, and 

quantitative exposure information is lacking. Therefore, human studies could not be used for oral 

dose-response analysis or to serve as the basis for the RfD. In the absence of human data, the 

animal studies were considered for dose-response analysis. 

Experimental animal studies considered for each health effect were evaluated using general 

study quality considerations discussed in Section 6 of the Preamble and in the literature search

section, and the attributes described above. The rationales for selecting the strongest studies that 

reported the health effects are summarized below.

Nervous System Effects

Convulsions were selected for dose-response analysis as a consistent and sensitive endpoint

of nervous system effects (see Section 1.3.1 for discussion). This endpoint was reported in seven

studies (Crouse et al., 2006; Lish et al., 1984; Levine et al., 1983b; Levine et al., 1981a; Cholakis et 

al., 1980; Martin and Hart, 1974; von Oettingen et al., 1949). Table 2-1 provides an overview of the 

information considered in the studies reporting nervous system effects (i.e., convulsions) evaluated 

for dose-response analysis.

Table 2-1. Information considered for evaluation of studies that examined 

convulsions

Study 

reference

Study design and size

Relevance of exposure paradigm Measurement 

of endpoint

Design

Number of 

animals Route Duration

Number of 

dose 

groups1

Levels

(mg/kg-d)

Purity

(%)

Incidence 

data reported

Crouse et al. 

(2006)

Toxicity study 10 rats/sex/ 

group

Gavage 13-wk 5 4−15 99.99 Yes

Cholakis et al. 

(1980)

Developmental 

study

24−25 female

rats/group

Gavage 14-d 3 0.2−20 89 Yes

Martin and 

Hart (1974)

Toxicity study 3 monkeys/

sex/group

Gavage 13-wk 3 0.1−10 Not 

specifie

d

Yes

Levine et al. 

(1983b)

Toxicity and 

carcinogenicity 

bioassay

75 rats/sex/ 

group

Diet 2-yr 4 0.3−40 89−99 No

Lish et al. 

(1984)

Toxicity and 

carcinogenicity 

bioassay

85 mice/sex/ 

group

Diet 2-yr 4 1.5−175 89−99 No

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Study 

reference

Study design and size

Relevance of exposure paradigm Measurement 

of endpoint

Design

Number of 

animals Route Duration

Number of 

dose 

groups1

Levels

(mg/kg-d)

Purity

(%)

Incidence 

data reported

Levine et al. 

(1981a)

Toxicity study 10 rats/sex/ 

group

diet 13-wk 5 10−600 85 No

von Oettingen 

et al. (1949)

Toxicity study 20 rats/group diet 13-wk 3 15−50 90−97 No

1

Excluding the control group.

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Incidence of convulsions was reported in three studies of RDX—all involving gavage 

administration: Crouse et al. (2006), Cholakis et al. (1980) (developmental toxicity study), and

Martin and Hart (1974). Qualitative findings of nervous system effects were reported in other 

chronic and subchronic studies—all involving dietary administration: Lish et al. (1984), Levine et 

al. (1983b), Levine et al. (1981a), and von Oettingen et al. (1949). Incidence data on neurotoxic 

effects of RDX were not collected in any of the dietary studies. For example, Levine et al. (1983b)

reported only that convulsions and other nervous system effects were noted in rats exposed to RDX 

for 2 years at the highest dose (40 mg/kg-day) tested. The studies that included incidence data (i.e., 

the gavage studies) were preferred over those studies only reporting qualitative results (i.e., the 

dietary studies). 

The three gavage studies reporting incidence data were further considered. Crouse et al. 

(2006) reported a dose-related increase in convulsions and tremors in both male and female F344 

rats following a 90-day oral (gavage) exposure to RDX. This study used a test material of high 

purity and six dose groups (including the control) that provided good resolution of the doseresponse curve. Cholakis et al. (1980) reported a dose-related increase in convulsions in a 

developmental toxicity study in F344 rats, following a 14-day exposure to RDX on gestational days 

(GDs) 6−19. Although this study was designed as a standard developmental toxicity study (i.e., not 

specifically to examine nervous system effects), it reported information on the identity of the test 

material and used three dose groups that adequately characterized the dose-response curve. 

Further, this study provided evidence of nervous system effects at a relatively low dose. The study 

in monkeys by Martin and Hart (1974) provides supporting evidence of nervous system effects 

(trembling, shaking, ataxia, hyperactive reflexes, and convulsions); however, this study was not 

selected for dose-response analysis because of small group sizes (n = 3/sex) and uncertainty in 

measures of exposures (e.g., purity of the test material was not specified, and reported emesis in 

some animals likely influenced the delivered dose).

Although the gavage studies reporting incidence data were preferred over four dietary 

studies (Lish et al., 1984; Levine et al., 1983b; Levine et al., 1981a; von Oettingen et al., 1949) that 

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1 did not provide incidence data, it is important to note that the reported neurotoxic effects in the 

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dietary studies were observed at dose levels higher than the doses at which effects were observed 

in the gavage studies (Crouse et al., 2006; Cholakis et al., 1980; Martin and Hart, 1974). Given this 

potential difference based on dosing method, the dietary studies were also considered for 

quantitative analysis, despite the lack of incidence data, to evaluate the influence of oral dosing 

method on candidate reference values. In the 2-year study by Levine et al. (1983b), a LOAEL for 

nervous system effects (convulsions, tremors, and hyper-irritability) of 40 mg/kg-day and a NOAEL 

of 8 mg/kg-day were identified. Other studies identified higher effect levels (i.e., 100 mg/kg-day in 

the 2-year mouse study by Lish et al. (1984) and 50 mg/kg-day in the 3-month rat study by von 

Oettingen et al. (1949)), and, with the exception of Lish et al. (1984), used shorter exposure 

durations. The unusual dosing regimen in the Cholakis et al. (1980) 13-week mouse study 

precluded identification of a NOAEL and LOAEL, and the single-dose design of the 6-week dog study 

by von Oettingen et al. (1949) did not allow identification of a NOAEL. As discussed in Section 1.2.1 

and Table 1-3, the technical report of the 13-week study by Levine et al. (1981a) inconsistently 

identified the dose level at which convulsions occurred; therefore, a reliable NOAEL and LOAEL 

from this study could not be identified. 

Therefore, two gavage studies, Crouse et al. (2006) and Cholakis et al. (1980), and one 

dietary study, Levine et al. (1983b), were selected for dose-response analysis.

Kidney and Other Urogenital System Effects

Suppurative prostatitis was selected for dose-response analysis. It is considered to be a 

sensitive marker for the broader range of urogenital effects observed in F344 male rats in a 2-year 

study by Levine et al. (1983b). The Levine et al. (1983b) study: (1) included a histopathological 

examination of the kidney and other urogenital system tissues at 6-, 12-, and 24-month time points; 

(2) included four dose groups and a control group, and adequate numbers of animals per dose 

group (75/sex/group, with interim sacrifice groups of 10/sex/group at 6 and 12 months); and

(3) reported individual animal data. This study, the only one to identify suppurative prostatisis,

was selected for dose-response analysis. 

Male Reproductive Toxicity

Testicular degeneration was selected for dose-response analysis. Lish et al. (1984)

observed a dose-related increase in the incidence of testicular degeneration in mice following 

chronic administration of RDX in the diet. This 2-year study: (1) included histopathological 

examination of male reproductive organs; (2) included four dose groups and a control group, and

adequate numbers of animals per dose group (85/sex/group, with interim sacrifice groups of 

10/sex/group at 6 and 12 months); and (3) reported individual animal data. This study, the only 

one to identify testicular degeneration, was selected for dose-response analysis. 

This document is a draft for review purposes only and does not constitute Agency policy.

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2.1.2. Methods of Analysis

1 No biologically based dose-response models are available for RDX. In this situation, the U.S. 

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Environmental Protection Agency (EPA) evaluates a range of dose-response models thought to be 

consistent with underlying biological processes to determine how best to empirically model the 

dose-response relationship in the range of the observed data. Consistent with this approach, EPA 

evaluated dose-response information with the models available in EPA’s Benchmark Dose Software 

(BMDS, versions 2.4 and 2.5). EPA estimated the benchmark dose (BMD) and BMDL using a 

benchmark response (BMR) selected for each effect. A conceptual model of the analysis approach 

used for RDX is provided in Figure 2-1. In this assessment, points of departure (PODs) are 

identified through BMD modeling (preferred) or identification of a NOAEL, and followed by animalto-human extrapolation through the use of physiologically based pharmacokinetic (PBPK) models 

or the application of a dosimetric adjustment factor, depending on the data available.

Figure 2-1. Conceptual approach to dose-response modeling for oral exposure. 

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Nervous System Effects

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amenable to BMD modeling. For Crouse et al. (2006), statistical analysis conducted by EPA 

indicated no significant difference in convulsion rates of male and female rats (Mantel-Haenszel test

for independence; see Table 2-2); thus, combined incidence data from male and female rats were 

used for modeling convulsion data from this study. A BMR of 1% extra risk (ER) for convulsions 

was used to address the severity of this endpoint; modeling with 5 and 10% ER is provided in 

Appendix D (see Section D.1.2, Tables D-3 to D-14) for comparative purposes. In general, for 

noncancer effects, severe endpoints are not typically used as the basis of a noncancer risk value 

because of relatively high uncertainty in extrapolating to a level of exposure likely to be without 

appreciable risk. The use of a 1% ER BMR for convulsions in Crouse et al. (2006) resulted in 

extrapolation below the range of the experimental doses. However, the BMD of 3.02 mg/kg-day 

was not far below the dose range of 4−15 mg/kg-day used in the study; thus, this extrapolation was 

considered moderate. In addition to uncertainty from extrapolation, model uncertainty from the 

use of the 1% ER BMR can be a concern. However, the BMDLs from Crouse et al. (2006) ranged 

from 0.54 to 2.90, a 5.4-fold difference, which is also not considered large, so the use of a 1% ER 

BMR did not result in substantial model uncertainty. 

Because incidence data for convulsions were not provided by Levine et al. (1983b), a 

NOAEL was used as the POD for this dataset rather than a BMDL. 

Table 2-2 summarizes the PODs derived for each data set. More detailed BMD modeling 

information is available in Appendix D.

Kidney/Urogenital System Effects

Incidence data on suppurative prostatitis as reported by Levine et al. (1983b) were 

amenable to BMD modeling. A BMR of 10% ER was applied under the assumption that it represents 

a minimally biologically significant level of change. Table 2-2 summarizes the POD derived using 

data on the incidence of suppurative prostatitis. More detailed BMD modeling information is 

available in Appendix D.

Male Reproductive Effects

Incidence data on testicular degeneration as reported by Lish et al. (1984) were amenable 

to modeling. A BMR of 10% ER was applied under the assumption that it represents a minimally 

biologically significant level of change. Table 2-2 summarizes the POD derived using date on the 

incidence of testicular degeneration. More detailed BMD modeling information is available in 

Appendix D.

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1 Table 2-2. Summary of derivation of PODs following oral exposure to RDX

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Endpoint and 

reference

(exposure 

duration/route) Species/sex Modela BMR

BMD

(mg/kg-d)

BMDL 

(mg/kg-d)

PODHED (mg/kg-d)

Administered 

doseb

RDX 

AUCc

RDX 

Cmaxd

Nervous system 

Incidence of 

convulsions

Crouse et al. (2006)

(90-d/gavage)

Male and 

female F344 

rat, combinede 

Multistage 2° 1% ER 3.02 0.57 0.14 0.28 0.37

Incidence of 

convulsions 

Cholakis et al. (1980)

(GDs 6−19/gavage)

Female F344 

rat

Quantallinear1% ER 0.18 0.12 0.03 0.06 0.08

Incidence of 

convulsions

Levine et al. (1983b)

(2-yr/diet)

Male and 

female F344 

rat

LOAEL = 40 mg/kg-d; NOAEL = 8 mg/kg-df 1.9 3.9 4.3

Kidney/urogenital system

Incidence of 

suppurative 

prostatitis

Levine et al. (1983b)

(2-yr/diet)

Male F344 rat LogProbit 10% ER 1.67 0.47 0.11 0.23 0.25

Male reproductive system

Incidence of 

testicular 

degeneration

Lish et al. (1984)

(2-yr/diet)

Male B6C3F1

mouse

LogProbit 10% ER 56.0 16.3 2.4 0.08 0.18 

a

For modeling details, see Appendix D.

b

POD was converted to an HED using a standard DAF based on BW3/4. c

POD was converted to an HED based on the equivalence of internal RDX dose (expressed as AUC for RDX 

concentration in arterial blood) derived using PBPK models. d

POD was converted to an HED based on the equivalence of internal RDX dose (expressed as peak RDX 

concentration in arterial blood, Cmax) derived using PBPK models. e

Exact Mantel-Haenszel test for independence between convulsion incidence and sex, stratified by dose, yielded 

p-value >0.05. f

Nervous system effects for male and female rats reported qualitatively; incidence of convulsions and other 

nervous system effects was not reported. Therefore, available data do not support BMD modeling.

AUC = area under the curve; BW = body weight; DAF = dosimetric adjustment factor; ER = extra risk; HED = human 

equivalent dose

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Human Extrapolation

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equivalent doses (HEDs) from data in laboratory animals, with the preferred approach being PBPK 

modeling. Other approaches can include using chemical-specific information in the absence of a 

complete PBPK model. In lieu of either reliable, chemical-specific models or data to inform the 

derivation of human equivalent oral exposures, a body weight scaling to the 3⁄4 power (i.e., BW3/4) 

approach is generally applied to extrapolate toxicologically equivalent doses of orally administered 

agents from adult laboratory animals to adult humans.

Candidate PODs for endpoints selected from rat and mouse bioassays were expressed as 

HEDs. HEDs were derived using both PBPK modeling (with alternative measures of internal dose), 

and a BW3/4 scaling approach. These approaches are outlined in Figure 2-1, and the resulting 

PODHED values are presented in Table 2-2. 

Extrapolation using PBPK modeling. PBPK models for RDX in rats, humans, and mice have 

been published (Sweeney et al., 2012a; Sweeney et al., 2012b; Krishnan et al., 2009) based on RDXspecific data. EPA evaluated and further developed these models for extrapolating doses from 

animals to humans (see Appendix C, Section C.1.5). In general, appropriately chosen internal dose 

metrics are expected to correlate more closely with toxic responses than external doses for effects 

that are not occurring at the point of contact (McLanahan et al., 2012). Therefore, PBPK modelderived arterial blood concentration of RDX is considered a better dose-metric for extrapolation of 

health effects than administered dose when there is adequate confidence in the estimated value. 

The PBPK models for RDX were used to estimate two dose metrics: the area under the curve (AUC) 

and the peak concentration (Cmax) for RDX concentration in arterial blood. The AUC represents the 

average blood RDX concentration for the exposure duration normalized to 24 hours and the Cmax

represents the maximum RDX concentration for the exposure duration. 

It appears logical to use RDX concentration levels in the brain as the internal dose metric for 

analyzing convulsion data. Nevertheless, the blood concentration of RDX was preferred as the dose 

metric due to greater confidence in modeling this variable. This is because of the substantially 

greater number of measurements of RDX blood levels used in calibrating model parameters. 

Additionally, predictions of RDX concentrations in the brain are highly correlated with RDX blood 

concentrations because the brain compartment does not have absorption, metabolism, or 

elimination of RDX. Greater confidence was placed in model estimates of blood AUC than peak 

blood concentrations because, as discussed in Appendix C, Section C.1.5, the rate constant for oral 

absorption (KAS) is uncertain, and peak concentrations are more sensitive to variations in this 

parameter than average values. RDX-induction of convulsions and seizures appears to be more 

strongly correlated with dose than exposure duration, which might argue for use of peak blood 

concentration as an appropriate dose metric; however, biological support for blood AUC, rather 

than peak blood concentration, comes from mechanistic information on RDX binding at the 

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1 picrotoxin convulsant site of the gamma-amino butyric acid (GABA) channel. There is evidence 

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from examination of picrotoxin binding to GABAA that a resulting period of elevated neuronal 

activity post-exposure could result in increased likelihood of seizures developing over time or other 

longer-term effects on normal brain function (see Section 1.2.1 for further discussion). Therefore, 

the AUC for RDX concentration in arterial blood was selected as the internal dose metric for 

analyzing dose-response data for convulsions. Nevertheless, the PODHED values based on both 

blood AUC and peak blood concentration (Cmax) are presented in Table 2-2 for completeness.

The rodent PBPK model was applied to the BMDLs generated from BMD modeling to 

determine the animal internal dose, expressed as the AUC of RDX blood concentration, and 

representing the cross-species toxicologically equivalent (internal) dose. The human PBPK model 

was then applied to derive the corresponding HEDs (see Figure 2-2). Because the AUC is linear 

with exposure level, at least in the exposure range of interest, the value of the HED would be the 

same whether the rat or mouse PBPK model is applied before or after BMD modeling is performed. 

Because the sequence of the calculation does not influence the results, applying the PBPK model 

after BMD modeling is more efficient—BMD modeling would not have to be redone if there were 

changes to the PBPK model, and it is easier to evaluate and show two dose metrics (as discussed 

above). Because of relatively high confidence in the PBPK models developed for the rat and human, 

these models were used to derive reliable internal dose metrics for extrapolation. For datasets 

selected from the rat bioassays, the candidate oral values were calculated assuming cross-species 

toxicological equivalence of the AUC of RDX blood concentration derived from PBPK modeling. A 

published PBPK model for the mouse was evaluated (Sweeney et al., 2012b); however, major 

uncertainties were identified in this model. The mouse model was based on fitting both the 

absorption and metabolic rate constants to a single set of blood concentration measurements. In 

this study, the lowest dose at which RDX was detected was 35 mg/kg, an exposure level high 

enough to manifest some toxicity in the chronic mouse bioassay, and except for measurements from 

a single animal, all other data points were non-detects or excluded as outliers (Sweeney et al., 

2012b). The type of additional data that increased confidence in the rat and human models (e.g., in 

vitro measurements of RDX metabolism and RDX elimination data) are not available for mice. 

Consequently, confidence in the mouse model parameter values is low. Further, there are no data 

to enable characterizing the fraction of RDX that is metabolized in the mouse; this is problematic 

considering evidence that indicates that the role of metabolism in RDX toxicity may differ across 

species (e.g., mice may have more efficient or higher expression of the cytochrome P450 [CYP450] 

enzymes). Given the high sensitivity of the model to the metabolic rate constant, the uncertainty in 

mouse toxicokinetics significantly decreases confidence in using the mouse PBPK model for 

predicting mouse blood RDX concentrations. (See Summary of Confidence in PBPK Models for RDX 

in Appendix C, Section C.1.5 for further discussion of confidence in the mouse model.) Given the 

low confidence in the mouse PBPK model, the preferred approach for determining candidate oral 

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1 values for endpoints selected from the mouse bioassay is that based on the administered dose of 

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RDX extrapolated to humans using allometric BW3/4 scaling.

Extrapolation using BW3/4 scaling. HEDs were also calculated using a BW3/4 scaling 

approach consistent with EPA guidance (U.S. EPA, 2011). PODs (BMDLs or NOAELs) based on the 

RDX dose administered in the experimental animal study were adjusted by a standard dosimetric 

adjustment factor (DAF) derived as follows: 

DAF = (BWa

1/4/BWh1/4),

where 

BWa = animal body weight 

BWh = human body weight 

Using BWa values of 0.25 kg for rats and 0.035 kg for mice and a BWh of 70 kg for humans 

(U.S. EPA, 1988), the resulting DAFs for rats and mice are 0.24 and 0.15, respectively. Applying the 

DAF to the POD identified for effects in adult rats or mice yields a PODHED as follows (see Table 2-2): 

PODHED = laboratory animal dose (mg/kg-day) × DAF 

Further details of the BMDL modeling, BMDS outputs, and graphical results for the best fit 

model for each dataset included in Table 2-2 can be found in Appendix D, Section D.1. Details of the 

PBPK model evaluation used for extrapolation from BMDL values can be found in Appendix C, 

Section C.1.5. Table 2-2 summarizes the results of the BMD modeling and the PODHED for each data 

set discussed above.

2.1.3. Derivation of Candidate Values

Under EPA’s A Review of the Reference Dose and Reference Concentration Processes (U.S. EPA, 

2002) (Section 4.4.5), and as described in the Preamble, five possible areas of uncertainty and 

variability were considered when determining the application of UFs to the PODs presented in 

Table 2-2. An explanation follows: 

An intraspecies uncertainty factor, UFH, of 10 was applied to all PODs to account for 

potential differences in toxicokinetics and toxicodynamics in the absence of information on the 

variability of response in the human population following oral exposure to RDX. The available 

human pharmacokinetic data are not sufficient to inform human kinetic variability and derive a 

chemical-specific UF for intraspecies uncertainty.

An interspecies uncertainty factor, UFA, of 3 (101/2 = 3.16, rounded to 3) was applied to all 

PODs to account for uncertainty in characterizing the toxicokinetic and toxicodynamic differences 

between rodents and humans. For the testicular degeneration dataset from the mouse bioassay, 

mouse to human extrapolation was accomplished using BW3/4 scaling (see rationale in 

Section 2.1.2—Human Extrapolation), which addresses predominantly toxicokinetic and some 

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1 toxicodynamic aspects of cross-species extrapolation; residual uncertainty in toxicokinetic and 

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toxicodynamic extrapolation remains. In the absence of chemical-specific data to quantify this 

uncertainty, EPA’s BW3/4 guidance (U.S. EPA, 2011) recommends use of an uncertainty factor of 3. 

For datasets from the rat bioassays, a PBPK model was used to convert internal doses in rats to 

external doses in humans (see rationale in Section 2.1.2—Human Extrapolation). This reduces 

toxicokinetic uncertainty in extrapolating from the rat to humans, but does not account for 

interspecies differences due to toxicodynamics. A UFA of 3 was applied to account for this 

remaining toxicodynamic and any residual toxicokinetic uncertainty not accounted for by the PBPK 

model. 

A subchronic to chronic uncertainty factor, UFs, of 1 was applied to all PODs. This is because 

(1) in studies of subchronic or gestational exposure used to derive a POD, effects were seen at 

lower doses in the studies of shorter duration than in chronic studies, and (2) other studies used to 

derive a POD were of 2-year duration. Although EPA guidance recommends a default UFS of 10 on 

the assumption that effects in a subchonic study would occur at approximately 10-fold higher 

concentration than in a corresponding (but absent) chronic study (U.S. EPA, 2002), the RDX 

database does not support a UFS of 10. As discussed in Section 1.2.1, although Gerkin et al. (2010)

introduces the possibility of effects developing over longer-term exposures to RDX, in general, 

seizure induction appears to be more strongly correlated with dose level than with exposure 

duration. The available bioassays suggest that chronic exposure would not lead to effects at lower 

doses than those induced by subchronic exposure. In addition, chronic dietary doses associated 

with convulsions were ≥35 mg/kg-day and were at least fourfold higher than gavage doses that 

induced convulsions in 14- and 90-day studies (i.e., 2 mg/kg-day in Cholakis et al. (1980) and 

8 mg/kg-day in Crouse et al. (2006)) (also see Table 1-3 and Figure 1-1). This may be due to 

differences between dietary and gavage administration (see Sections 2.1.1 and 2.1.7). Nevertheless, 

these studies do not support the default expectation of observing effects in chronic studies at 

approximately 10-fold lower exposure levels than in subchronic studies. Accordingly, a UFs of 1 

was applied to PODs derived from studies of less-than-chronic duration. 

A LOAEL to NOAEL uncertainty factor, UFL, of 1 was applied to all POD values because the 

POD was a BMDL or a NOAEL. When the POD is a BMDL, the current approach is to address this 

factor as one of the considerations in selecting a BMR for BMD modeling. In this case, the BMR for

modeled endpoints was selected under the assumption that the BMR represents a minimal, 

biologically significant change for these effects. 

A database uncertainty factor, UFD, of 3 was applied to all POD values. The oral toxicity 

database for RDX includes subchronic and chronic toxicity studies in the rat and mouse, a twogeneration reproductive toxicity study in the rat, developmental toxicity studies in the rat and 

rabbit, and subchronic studies (with study design limitations) in the dog and monkey. As discussed 

below, some uncertainty is associated with characterization of the RDX neurotoxicity. 

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1 Analyses presented in Section 2.1.6 note that reference values derived from mortality data 

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would be similar to the RfD for RDX based on convulsions. EPA prefers to identify reference values 

based on upstream (less severe) effects that would precede frank effects like convulsions and 

mortality. Some uncertainty remains in our understanding of RDX-induced neurotoxicity. In part, 

this is due to limitations in study design to assess neurotoxicity across the RDX database; the 

frequency of animal observations in the available studies raises concerns that there may be 

underreporting of the true incidence of convulsions, and in general the reporting of this effect does 

not include a measure of the severity at the time of observation. No follow-up studies were 

identified that employed more sensitive assays to assess more subtle neurotoxicity. Uncertainties 

in the database for RDX neurotoxicity could be addressed by:

• Analysis of “convulsions” using more detailed behavioral scoring methods. In the available 

studies, “convulsion” can indicate a range of observable behaviors in response to altered 

brain activity, ranging from involuntary limb and facial twitches to tonic-clonic seizures in 

which animals exhibit a sustained (seconds to hours) and widespread loss of muscle control 

sometimes resulting in respiratory arrest and/or death. As there are studies where 

convulsions occur at the same dose as mortality, the convulsive activity in these studies is 

interpreted as severe. Scoring methods quantifying the occurrence of different behavioral 

aspects of the RDX-induced convulsions, such as the Racine scale (Racine, 1972), employed 

in Burdette et al. (1988) would provide a much more accurate, complete, and possibly more 

sensitive measure of RDX neurotoxicity. 

• Additional electrophysiological measures of epileptiform activity. Well-established and 

sensitive methods for evaluating brain activity exist. These measures could not only better 

describe the profile of RDX-induced convulsant activity, but could also be used to identify 

and quantify sub-convulsive effects of RDX exposure (e.g., EEG spiking). 

Electrophysiological characterization of the effects of RDX in vitro and in vivo has already 

been demonstrated by Williams et al. (2011). Additional studies building on this work, 

looking at the effects of different concentrations of RDX, could potentially identify more 

sensitive measures of RDX neurotoxicity.

• A FOB conducted by Crouse et al. (2006) provides information on neurobehavioral effects 

associated with RDX exposure, yet the results of that study did not identify notable effects 

associated with RDX exposure. While some components of the FOB testing conducted by 

Crouse et al. (2006) would be expected to give a screening-level evaluation of some stimuliinduced behaviors that have the potential to be related (e.g., response to handling, touch, 

click or open field), additional studies addressing whether RDX exposure alters the 

susceptibility to seizures elicited by traditional means could be informative. Burdette et al. 

(1988) examined seizure susceptibility in gavaged male Long Evans rats, but at doses ≥10 

mg/kg. Further evaluation of seizure susceptibility at doses lower than 10 mg/kg, and with 

longer exposure durations, may identify additional measures of RDX neurotoxicity. 

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1 • Further evaluation of potential developmental neurotoxicity (and specifically seizure 

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induction) associated with RDX exposure. Models for examining seizure-related behaviors 

during development exist, mainly involving manipulation and analyses in pre-weanling 

rodents. Hess-Ruth et al. (2007) reported possible transfer of RDX to offspring during 

gestation, as well as the presence of RDX in the milk of dams, indicating a potential for 

lactational transfer of RDX to offspring. Although examination of specific developmental 

neurotoxicity endpoints has not been conducted in studies of RDX toxicity, the available 

testing, including a two-generation reproductive toxicity study in the rat (Cholakis et al., 

1980), did not report any evidence of neurobehavioral effects in offspring exposed during 

gestation or lactation. However, confidence in the observation is reduced as there is a 

question if the extent of observation in Cholakis et al. (1980) was sufficient to accurately 

characterize neurobehavioral effects. Additional developmental neurotoxicity studies could 

further rule out the possibility that RDX exposure during development might result in 

immediate or delayed seizure activity, or predispose animals to developing seizures as 

adults. 

Overall, while the RDX database adequately covers major systemic effects, including reproductive 

and developmental effects, uncertainties in the adequacy of the database were identified in 

characterization of the neurotoxicity hazard. There is some concern that additional studies 

described above may lead to identification of a more sensitive endpoint or a lower POD. 

Accordingly, a UFD of 3 was applied to all derived PODs. 

Table 2-3 is a continuation of Table 2-2 and summarizes the application of UFs to each 

PODHED to derive a candidate value for each data set. The candidate values presented in Table 2-3 

are preliminary to the derivation of the organ/system-specific reference values. These candidate 

values are considered individually in the selection of a representative oral reference value for a 

specific hazard and subsequent overall RfD for RDX.

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1 Table 2-3. Effects and corresponding derivation of candidate values 

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7 

Endpoint and reference PODHEDa

POD 

type UFA UFH UFL UFS UFD

Composite 

UF

Candidate

value 

(mg/kg-d)

Nervous system (rat)

Incidence of convulsions

Crouse et al. (2006)

0.28 BMDL01 3 10 1 1 3 100 2.8 × 10−3

Incidence of convulsions 

Cholakis et al. (1980)

0.06 BMDL01 3 10 1 1 3 100 6.0 × 10−4

Incidence of convulsions

Levine et al. (1983b)

3.9 NOAEL 3 10 1 1 3 100 3.9 × 10−2

Kidney/urogenital system (rat)

Incidence of prostate 

suppurative inflammation

Levine et al. (1983b)

0.23 BMDL10 3 10 1 1 3 100 2.3 × 10−3

Male reproductive system (mouse)

Incidence of testicular 

degeneration

Lish et al. (1984)

2.4 BMDL10 3 10 1 1 3 100 2.4 × 10−2

a

PODHED values based on data from the rat were derived using PBPK modeling; the PODHED based on data from the 

mouse was derived using BW3/4 adjustment (see Section 2.1.2 and discussion of the PBPK models above and in 

Appendix C, Section C.1.5).

Figure 2-2 presents graphically the candidate values, UFs, and PODHED values, with each bar 

corresponding to one data set described in Tables 2-2 and 2-3. 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 Figure 2-2. Candidate values with corresponding POD and composite UF.

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2.1.4. Derivation of Organ/System-Specific Reference Doses

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system. Organ- or system-specific reference values may be useful for subsequent cumulative risk 

assessments that consider the combined effect of multiple agents acting at a common site.

Table 2-4. Organ/system-specific RfDs and overall RfD for RDX

Effect Basis

RfD 

(mg/kg-d)

Study exposure 

description Confidence

Nervous system Incidence of convulsions

(Crouse et al., 2006)

3 × 10−3 Subchronic Medium

Kidney/urogenital system Incidence of suppurative prostatitis

(Levine et al., 1983b)

2 × 10−3 Chronic Low

Male reproductive system Incidence of testicular degeneration

(Lish et al., 1984)

2 × 10−2 Chronic Low

Overall RfD Nervous system 3 × 10−3 Subchronic Medium

Nervous System Effects

The organ/system-specific RfD for nervous system effects was based on the incidence of 

convulsions in F344 rats reported in Crouse et al. (2006), a well-conducted study that used a 

99.99% pure form of RDX, five closely-spaced dose groups that provided a good characterization of 

the dose-response curve for convulsions, and an endpoint (convulsions) that was replicated across 

multiple studies. Although the candidate value derived from the developmental toxicity study in 

F344 rats by Cholakis et al. (1980) is lower (by approximately fivefold), there is greater certainty in 

the value derived from Crouse et al. (2006). Crouse et al. (2006) was specifically designed to assess 

the nervous system effects of RDX (including a functional observational battery), whereas Cholakis 

et al. (1980) was designed as a developmental toxicity study with only routine monitoring of 

clinical signs (the methods section states that “Dams were monitored daily for toxic signs”). Crouse 

et al. (2006) used five dose groups (plus the control) that provided good resolution of the doseresponse curve for RDX-induced convulsions, whereas Cholakis et al. (1980) used only three dose 

group (plus the control) with order of magnitude dose spacing, resulting in a less well defined 

characterization of the dose-response curve for this endpoint. Further, Crouse et al. (2006) used a 

higher purity test material than did Cholakis et al. (1980) (99.99% versus 88.6%, respectively). 

Finally, the Crouse et al. (2006) study used a longer exposure duration (90 days) than did the 

Cholakis et al. (1980) study (14 days), and is more representative of a chronic exposure duration. 

The lower candidate reference value from the Cholakis et al. (1980) developmental toxicity study 

could indicate that pregnant animals are a susceptible population, which could support selection of 

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1 this study as the basis for the RfD; however, as discussed in Section 1.3.3, the available studies in 

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pregnant and nonpregnants rats cannot be directly compared, and the available information is not 

considered sufficient to identify pregnant animals as a susceptible population.

As discussed in Section 2.1.1, the 2-year dietary study by Levine et al. (1983b) was also 

considered for RfD derivation because the available oral studies suggest that bolus doses of RDX

received with gavage administration may induce nervous system effects at doses lower than those 

resulting from dietary administration (recognizing that differences in particle size and purity of the 

test material may confound direct comparisons between gavage and dietary administration). 

Convulsion data from Levine et al. (1983b) yielded a PODHED 14-fold higher than the PODHED derived 

from Crouse et al. (2006). The POD derived from the Levine et al. (1983b) study is considered less 

certain than that derived from Crouse et al. (2006). Levine et al. (1983b) did not provide 

information on the incidence of neurotoxic effects, and BMD analysis was thus not supported (i.e., 

the POD was based on a NOAEL). As discussed in Section 1.2.1, the frequency of daily observations 

in the Levine et al. (1983b) study may not have been sufficient to provide an accurate measure of 

the occurrence of nervous system effects, potentially leading to underestimation of convulsions and 

other nervous system effects. For these reasons, and in light of the fact that data from the Levine et 

al. (1983b) study yielded a higher POD, Levine et al. (1983b) was not used as the basis for the 

organ/system-specific RfD for nervous system effects.

Kidney/Urogenital Effects

A single data set for incidence of suppurative prostatitis in male F344 rats as reported in a 

2-year dietary study by Levine et al. (1983b) was brought forward for quantitative analysis as a 

sensitive marker for the broader array of RDX-associated effects observed in the urogenital system. 

The RfD for kidney and other urogenital effects is based on this dataset. 

Male Reproductive Effects

A single dataset for male reproductive effects, specifically the incidence of testicular 

degeneration as reported in male B6C3F1 mice exposed to RDX in diet for 24 months (Lish et al., 

1984), was brought forward for quantitative analysis. The RfD for male reproductive effects is 

based on this dataset.

2.1.5. Selection of the Overall Reference Dose

Multiple organ/system-specific reference doses were derived for effects identified as 

potential hazards from RDX exposure, including nervous system effects, kidney and other 

urogenital effects, and male reproductive effects. Evidence for nervous system effects, and 

specifically convulsions, was observed in multiple studies, in multiple species, and following a range 

of exposure durations. In addition, the organ/system-specific RfD for nervous system effects was 

the lowest among the organ/system-specific RfDs derived for RDX. Evidence for dose-related 

effects on the urogenital system comes primarily from a single 2-year toxicity study in male rats 

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1 (Levine et al., 1983b), and evidence for male reproductive effects comes primarily from a single

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2-year toxicity study in mice (Lish et al., 1984); neither a second chronic study in the rat that 

evaluated prostate histopathology nor a second mouse study was available to validate and replicate 

these findings. 

The organ/system-specific RfD of 3 × 10−3 mg/kg-day for nervous system effects in the rat 

as reported by Crouse et al. (2006) is selected as the overall RfD for RDX given the strength of 

evidence for the nervous system as a hazard of RDX exposure, and as the lowest organ/systemspecific RfD. This overall RfD provides an exposure level below which effects associated with RDX 

exposure are not expected to occur.

The overall RfD is derived to be protective of all types of effects for a given duration of 

exposure, and is intended to protect the population as a whole, including potentially susceptible 

subgroups (U.S. EPA, 2002). Decisions concerning averaging exposures over time for comparison 

with the RfD should consider the types of toxicological effects and specific lifestages of concern. 

Fluctuations in exposure levels that result in elevated exposures during these lifestages could 

potentially lead to an appreciable risk, even if average levels over the full exposure duration were 

less than or equal to the RfD. In the case of RDX, no specific lifestages have been identified as a 

potentially susceptible subgroup. 

2.1.6. Comparison with Mortality LD01s 

As previously discussed, mortality was considered in discussions of other organ/systemspecific toxicity (and in particular, effects on the nervous system and kidney). EPA did not develop 

a candidate RfC from mortality because EPA generally does not develop reference values based on 

frank effects such as mortality, rather, reference values are generally based on earlier (less severe) 

upstream events, where possible, in order to protect against all adverse outcomes. Nevertheless, 

additional analysis of mortality data was undertaken because some studies (see Table 2-5) 

identified mortality at the same RDX dose that induced nervous system effects (Crouse et al. (2006); 

Angerhofer et al. (1986); Cholakis et al. (1980); von Oettingen et al. (1949)). 

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1 Table 2-5. Comparison of dose levels associated with mortality and 

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convulsions in selected studies

Study

Doses associated with 

mortality

Doses associated with 

convulsions

Crouse et al. (2006)

Rats, F344, 10/sex/group

0, 4, 8, 10, 12, or 15 mg/kg-d 

13 wks/gavage

≥8 mg/kg-d ≥8 mg/kg-d

von Oettingen et al. (1949)

Rats, sex/strain not specified, 20/group

0, 15, 25, or 50 mg/kg-d 

13 wks/diet

≥25 mg/kg-day ≥25 mg/kg-d

Cholakis et al. (1980)

Rats, F344, 24−25 females/group

0, 0.2, 2.0, or 20 mg/kg-d 

GDs 6−19/gavage

20 mg/kg-d Primarily 20 mg/kg-d; 1 

convulsion at 2 mg/kg-d 

Angerhofer et al. (1986)

Rats, Sprague-Dawley, 39−51 mated females/group

0, 2, 6, or 20 mg/kg-d 

GDs 6−15/gavage

Primarily at 20 mg/kg-d, but 

one death each at 2 and 

6 mg/kg-d 

20 mg/kg-d

A discussion of mortality evidence for RDX is presented in Appendix C, Section C.3.1, and the 

relationship between mortality and nervous system effects in Sections 1.2.1 and 1.3.1. Unscheduled 

deaths were observed as early as day 8 of a 90-day gavage study (Crouse et al., 2006) and in 

development toxicity studies with exposure durations of two weeks (Angerhofer et al. (1986); 

Cholakis et al. (1980)). 

 Given the proximity in the dose at which mortality and nervous system effects were 

observed in several studies, the dose-response relationships for mortality were compared across 

studies with similar durations to those in Table 2-5 by comparing the LD01 (the dose expected to be 

lethal to 1% of the animals) or NOAELs derived from each study. In addition, these LD01s and 

NOAELs were compare to the BMD01 for convulsions used to derive the RfD.14 Interpretation of 

mortality data from chronic exposure studies in mice and rats is complicated by other treatmentrelated effects and pathology regularly observed in aging animals (e.g., kidney pathology, neoplastic 

lesions), and was not considered in this analysis. Other studies that were less informative and not 

considered in this analysis are not presented in Table 2-6.15 

14 BMDs were compared, as opposed to BMDLs, because, as stated on p. 20 of the BMD Technical Guidance 

(U.S. EPA, 2012b), “In general, it is recommended that comparisons across chemicals/studies/endpoints be 

based on central estimates; this is in contrast to using lower bounds for PODs for reference values...” 

15The following less informative studies were not included in the analysis of early mortality: 

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1 Table 2-6. Summary of dose-response evaluation for mortality following oral 

2 exposure to RDX

Reference

(exposure duration/route) Species/sex Modela BMR

LD01

(mg/kg-d)

LDL01

(mg/kg-d)

Diet studies

Lish et al. (1984)

(11-week data from 2-yr 

study/diet)

Male and female 

B6C3F1 mouse

Not amenable to 

modeling

NOAEL: 35 mg/kg-day

95% CI for response: 0-4%

Levine et al. (1981a)

(13-wk/diet)

Male and female 

F344 rat, combined

Multistage 40 1% ER 7.8 2.2

von Oettingen et al. (1949)

(13-wk/diet)

Rats, sex/strain not 

specified

Not amenable to 

modeling

NOAEL: 15 mg/kg-day

95% CI for response: 0-15%

Cholakis et al. (1980)

(2-generation design/diet)

Female CD rat Not amenable to 

modeling

NOAEL: 16 mg/kg-day

95% CI for response: 0–13%

Levine et al. (1983b)

(13-week data from 2-yr 

study/diet)

Male and female 

F344 rat

NA

(no mortality at 

highest dose tested)

NOAEL: 40 mg/kg-day

95% CI for response: 0-4%

Cholakis et al. (1980)

(13-wk/diet)

Male and female 

F344 rat

NA

(no mortality at 

highest dose tested)

NOAEL: 40 mg/kg-day

95% CI for response: 0-25%

Gavage studies

Crouse et al. (2006)

(90-d/gavage)

Male and female 

F344 rat, combined

Multistage 20 1% ER 2.1 0.46

Cholakis et al. (1980)

(GDs 6‒19/gavage)

Female F344 rat Not amenable to 

modeling

NOAEL: 2 mg/kg-day

95% CI for response: 0–12%

Angerhofer et al. (1986)

(GD 6‒15/gavage)

Female SD rat Multistage 30 1% ER 1.7 0.59

Cholakis et al. (1980)

(GDs 7‒29/gavage)

Female New 

Zealand white rabbit

NA

(no mortality at 

highest dose tested)

NOAEL: 20 mg/kg-day

95% CI for response: 0-22%

• 13-week dietary study in the mouse by Cholakis et al. (1980). Mortality was observed only in the highdose group (257−276 mg/kg-day TWA), and the unusual dosing regimen precluded identification of a 

NOAEL or LOAEL.

• 13-week dietary study in the dog by Hart (1974) and 13-week study in the monkey by Martin and Hart 

(1974). Both studies used small group sizes (3 animals/dose group), and no animals died on study 

(although one high-dose monkey was euthanized). 

• 6-week dietary study in the dog from the 1949 publication by von Oettingen et al. (1949). This dog study 

included only one treatment group and recorded only one death.

• 30-day gavage study in the rat by MacPhail et al. (1985). The authors did not identify treatment-related 

mortality, but reporting was limited.

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Reference

(exposure duration/route) Species/sex Modela BMR

LD01

(mg/kg-d)

LDL01

(mg/kg-d)

a

For modeling details, see Appendix D, Section D.1.3.

ER = extra risk

LD01 = dose expected to be lethal to 1% of the animals 

LDL01 = lower confidence limit on the LD01

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Of the studies in Table 2-6, dose-response analysis was conducted for all studies that 

showed an increased incidence of unscheduled deaths. LD values are provided in Table 2-6, and 

detailed modeling results are provided in Appendix D, Section D.1.3. Mortality was observed only 

at the highest dose tested at week 11 in the 2-year mouse study by Lish et al. (1984), in the 13-week 

rat study by von Oettingen et al. (1949), and in the two-generation reproductive and developmental 

toxicity studies by Cholakis et al. (1980). In these cases, data were not amenable to LD01 estimation, 

and a NOAEL (with its associated confidence interval, CI) was used in this comparative analysis 

instead.

LD01 values for mortality in Table 2-6 range from 1.7 mg/kg-day (10-day gavage exposure

in pregnant rats) to 7.8 mg/kg-day (13-week dietary exposure in rats), with the lower values 

generally from studies that administered RDX by gavage. These values may be compared to the 

BMD for convulsions from Crouse et al. (2006) that was used as basis for the overall RfD for RDX. 

The BMD for convulsions of 3.0 mg/kg-day is in the middle of the distribution of calculated LD01s, 

and the lowest LD01 of 1.7 mg/kg-day is within twofold of the convulsion BMD of 3.0 mg/kg-day. 

The NOAELs from studies where mortality was observed tend to be higher than the LDs. 

However, NOAELs are not directly comparable to BMD01s for several reasons. Confidence intervals 

for the observed responses of 0% characterize some statistical uncertainty for NOAELs from 

studies that could not be modeled (note that the upper bound on an observed response of 0% is not 

directly comparable to a lower bound on a benchmark dose). The confidence intervals suggest that 

comparable 1% levels for these datasets could be somewhat lower than the NOAELs. On the other 

hand, dose-spacing can affect the interpretation of NOAELs, such that from the Cholakis et al. 

(1980) developmental toxicity study because of the wide (order-of-magnitude) spacing between 

doses in that study (i.e., the reported NOAEL is 10-fold lower than the dose associated with 17% 

mortality at 20 mg/kg-day).

In general, this comparison indicates that reference values derived from mortality data 

would be similar to the final RfD for RDX based on convulsions, assuming the application of the 

same extrapolation procedures and uncertainty factors. However, the similarity in RDX doses 

associated with both mortality and nervous system effects should be taken into consideration when 

using this RfD, and in particular in evaluating exposures that exceed the RfD in light of the severity 

of mortality.

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2.1.7. Uncertainties in the Derivation of the Reference Dose

1 To derive the RfD, the UF approach (U.S. EPA, 2000a, 1994) was applied to a PODHED based 

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on nervous system effects in rats exposed to RDX for a subchronic duration. UFs were applied to 

the PODHED values to account for uncertainties in extrapolating from an animal bioassay to human 

exposure, the likely existence of a diverse human population of varying susceptibilities, and 

subchronic to chronic duration. For the most part, these extrapolations are carried out with default 

approaches given the lack of data to inform individual steps. One exception is the use of PBPK 

modeling to perform interspecies (i.e., rat to human) extrapolation. Uncertainties associated with 

the PBPK models are considered in Appendix C, Section C.1.5. 

Nervous system effects have been documented in multiple studies and animal species and 

strains; however, some uncertainty is associated with the incidence of reported neurological effects 

in studies that employed a study design that did not monitor animals with sufficient frequency to 

accurately record neurobehavioral effects, including convulsions. In the study used to derive the 

RfD (Crouse et al., 2006), Johnson (2015a) noted that convulsions were observed infrequently 

outside the dosing period; more often, seizures were observed during the 2-hour (gavage) dosing 

period, typically within 60−90 minutes of dosing. Similar information was not available for other 

studies to assess the likelihood that observations of convulsions were missed. However, animals 

were not monitored continuously during the Crouse et al. (2006) study, and investigators reported 

that nearly all observed pre-term deaths in rats exposed to the three higher doses were preceded 

by signs of neurotoxicity. If an animal died during the study as a result of effects on the nervous 

system, convulsions preceding death could have been missed, resulting in an underestimation of 

the incidence of convulsions. Conversely, attributing all mortality to neurotoxicity (i.e., all deaths 

were preceded by convulsions that may not have been observed) could result in an overestimation 

of the incidence of convulsions. A dose-response analysis of the combined incidence of seizures and 

mortality from Crouse et al. (2006) was conducted to evaluate the impact of these assumptions, as 

the true convulsion incidence would likely fall somewhere between the observed convulsion 

incidence and the combined incidence of convulsions and mortality. The PODHED of 0.24 mg/kg-day 

for a combined incidence of convulsions and mortality16 was compared to the PODHED of

0.28 mg/kg-day for convulsions alone, indicating that the addition of mortality incidence did not 

have a significant impact. Therefore, the RfD based on the incidence of convulsions alone does not 

appear to underestimate the toxicity associated with RDX. 

Some uncertainty is also associated with the influence of the method of oral dosing on the 

magnitude of dose required to induce nervous system effects. As noted in Section 1.2.1, gavage 

16BMD = 2.56 mg/kg-day; BMDL = 0.49 mg/kg-day (see Appendix D.1.2 for BMD modeling results). The 

PODHED value was derived using PBPK modeling (see Section 2.1.2 and discussion of the PBPK models in 

Appendix C, Section C.1.5).

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1 administration generally induced convulsions in experimental animals at lower doses than did 

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dietary administration, possibly due to the bolus dose resulting from gavage administration that 

could lead to comparatively faster absorption and higher peak blood concentrations of RDX. To 

some extent, this uncertainty is reflected in the 14-fold difference in the candidate PODHED values 

derived from the Crouse et al. (2006) (gavage administration) and Levine et al. (1983b) (dietary 

administration) studies. A more rigorous examination of the effect of oral dosing method cannot be 

performed because of the differences in test materials and study designs used in the available 

gavage and dietary studies that could also have contributed to differences in response (e.g., test 

article purity and particle size, number and spacing of dose groups, exposure duration, frequency of 

clinical observations, and thoroughness of the reporting of observations).

Although the database is adequate for reference value derivation, uncertainty is associated 

with the consistency in toxicity results across studies that used RDX test materials that differed in 

purity, formulation, and particle size. There is evidence that differences in test material 

formulation and particle size (i.e., the increased surface area associated with finely powdered RDX 

allows for increased absorption) can affect oral bioavailability of RDX and subsequent toxicity (see 

discussion in Appendix C, Section C.1.5, Absorption of RDX from the GI Tract). 

2.1.8. Confidence Statement

A confidence level of high, medium, or low is assigned to the study used to derive the RfD, 

the overall database, and the RfD itself, as described in Section 4.3.9.2 of EPA’s Methods for 

Derivation of Inhalation Reference Concentrations and Application of Inhalation Dosimetry (U.S. EPA, 

1994). The overall confidence in this RfD is medium. Confidence in the principal study (Crouse et 

al., 2006) is high. The study was well-conducted, utilized 99.99% pure RDX, and had five closelyspaced dose groups that allowed characterization of dose-response curves for convulsions in the 

dose range of interest. One limitation identified by study authors was the limited ability of the FOB 

to fully identify neurobehavioral effects at doses ≥8 mg/kg-day due to the timing of the dosing 

procedure and timing of the FOB screening. Confidence in the database is medium. The database 

includes three chronic studies in rats and mice; eight subchronic studies in rats, mice, dogs, and 

monkeys; two short-term studies; and four reproductive/developmental toxicity studies in rats and 

rabbits (including a two-generation reproductive study). Confidence in the database is reduced

largely because of (1) differences in test material used across studies, and (2) uncertainties in the 

influence of oral dosing methods. As discussed in Section 2.1.7 and Appendix C, Section C.1.5, 

differences in test material formulation and particle size may affect RDX absorption and subsequent 

toxicity, which in turn could influence the characterization and integration of toxicity findings 

across studies. The available evidence also suggests that bolus dosing of RDX that results from 

gavage administration induces neurotoxicity at doses lower than administration in the diet, 

although a rigorous examination of these differences cannot be performed with the available 

database. To the extent that dietary administration is more representative of potential human 

exposures to RDX, the use of toxicity data from a gavage (bolus dosing) study introduces 

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1 uncertainty in the RfD. Reflecting high confidence in the principal study and medium confidence in 

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the database, overall confidence in the RfD is medium. 

2.1.9. Previous IRIS Assessment

The previous RfD for RDX, posted to the Integrated Risk Information System (IRIS) database 

in 1988, was based on a 2-year rat feeding study by Levine et al. (1983b). The no-observed-effect 

level (NOEL) of 0.3 mg/kg-day based on suppurative inflammation of the prostate in male F344 rats 

from this study was identified as the POD. An RfD of 3 × 10−3 mg/kg-day was derived following 

application of an overall UF of 100 (UFA = 10, UFH = 10).

2.2. INHALATION REFERENCE CONCENTRATION FOR EFFECTS OTHER 

THAN CANCER

The inhalation reference concentration (RfC, expressed in units of mg/m3) is defined as an 

estimate (with uncertainty spanning perhaps an order of magnitude) of a continuous inhalation 

exposure to the human population (including sensitive subgroups) that is likely to be without an 

appreciable risk of deleterious effects during a lifetime. It can be derived from a NOAEL, LOAEL, or 

the 95% lower bound on the benchmark concentration (BMCL), with UFs generally applied to 

reflect limitations of the data used. 

As discussed in Section 1.3.1, the available inhalation literature does not support 

characterization of the health hazards specifically associated with chronic inhalation exposure to 

RDX, nor do the studies support quantitative dose-response analysis. Of the available human 

epidemiological studies of RDX (West and Stafford, 1997; Ma and Li, 1993; Hathaway and Buck, 

1977), none provided data that could be used for dose-response analysis. The studies by Ma and Li 

(1993) of neurobehavioral effects in Chinese workers and West and Stafford (1997) of 

hematological abnormalities in ordnance factory workers had numerous methodological 

limitations that preclude their use for quantitative analysis (see Literature Search Strategy | Study 

Selection and Evaluation). The study by Hathaway and Buck (1977) found no evidence of adverse 

health effects in munition plant workers (based on evaluation of liver function, renal function, and 

hematology), and therefore does not identify a POD at which there would be an effect from which to

derive an RfC. Multiple case reports provide some evidence of effects in humans associated with 

acute exposure to RDX; however, while case reports can support the identification of hazards 

associated with RDX exposure, data from case reports are inadequate for dose-response analysis 

and subsequent derivation of a chronic reference value because of short exposure durations and 

incomplete or missing quantitative exposure information.

As discussed in Literature Search Strategy | Study Selection and Evaluation, a single 

experimental animal study involving inhalation exposure was identified in the Defense Technical 

Information Center (DTIC) database; the study is not publicly available. However, the study would 

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1 not have provided useful data on responses to inhaled RDX, as it was limited by small numbers of 

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animals tested, lack of controls, and incomplete reporting of exposure levels. 

Therefore, the available health effects literature does not support the derivation of an RfC 

for RDX. While inhalation absorption of RDX particulates is a plausible route of exposure, there are 

no toxicokinetic studies of RDX inhalation absorption to support an inhalation model. Therefore, a 

PBPK model for inhaled RDX was not developed to support route-to-route extrapolation from the 

RfD.

2.2.1. Previous IRIS Assessment

An RfC for RDX was not previously derived under the IRIS Program. 

2.3. ORAL SLOPE FACTOR FOR CANCER

The oral slope factor (OSF) is a plausible upper bound on the estimate of risk per 

mg/kg-day of oral exposure. The OSF can be multiplied by an estimate of lifetime exposure (in 

mg/kg-day) to estimate the lifetime cancer risk.

2.3.1. Analysis of Carcinogenicity Data

As noted in Section 1.3.2, there is “suggestive evidence of carcinogenic potential” for RDX. 

The Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a) state: 

When there is suggestive evidence, the Agency generally would not attempt a doseresponse assessment, as the nature of the data generally would not support one;

however, when the evidence includes a well-conducted study, quantitative analyses 

may be useful for some purposes, for example, providing a sense of the magnitude 

and uncertainty of potential risks, ranking potential hazards, or setting research 

priorities. 

In the case of RDX, there are well-conducted studies that tested large numbers of animals at 

multiple dose levels (Lish et al., 1984; Levine et al., 1983b), making the cancer response suitable for 

dose-response analysis. Considering the data from these studies, along with the uncertainty 

associated with the suggestive nature of the weight of evidence, quantitative analysis of the tumor 

data may be useful for providing a sense of the magnitude of potential carcinogenic risk.

The incidences of liver and lung tumors in female mice from the study by Lish et al. (1984)

were selected for quantitative dose-response analysis. The study by Lish et al. (1984): (1) included 

comprehensive histopathological examination of major organs; (2) contained four dose groups and 

a control; (3) used adequate numbers of animals per dose group (85/sex/group, with interim 

sacrifice groups of 10/sex/group at 6 and 12 months) and a sufficient overall exposure duration 

(2 years); and (4) adequately reported methods and results (including individual animal data). 

Female mouse liver tissues from the original unpublished study by Lish et al. (1984) were 

reevaluated by a pathology working group (PWG) (Parker et al., 2006) in order to apply more 

up-to-date histopathological criteria established by Harada et al. (1999). The updated liver tumor 

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1 incidences from the PWG reanalysis of Lish et al. (1984) were used for quantitative dose-response 

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analysis. 

In the case of both liver and lung tumors, benign and malignant tumors (i.e., adenomas and 

carcinomas) were combined for dose-response analysis because benign and malignant tumors in 

both organs develop from the same cell line and there is evidence for progression from benign to 

the malignant stage (U.S. EPA, 2005a; McConnell et al., 1986). Female mouse liver and lung tumor 

incidences from the Lish et al. (1984) study are summarized in Appendix D, Table D-23. 

The incidence of hepatocellular carcinomas in male F344 rats from the study by Levine et al. 

(1983b) and the incidence of alveolar/bronchiolar carcinomas in male B6C3F1 mice from the study 

by Lish et al. (1984) were also considered for quantitative dose-response analysis. Both studies 

were well conducted, using similar study designs (described above). In both instances, the 

response was less robust than the response observed in female mice from the Lish et al. (1984)

study. The hepatocellular carcinoma result in male F344 rats is based on a small number of tumors

(two each at 8 and 40 mg/kg-day, and one in controls). The alveolar/bronchiolar carcinomas in 

male B6C3F1 mice showed a positive trend; however, a positive trend was not observed when the 

incidence of adenomas and carcinomas was combined. Modeling results are provided in 

Appendix D, Section D.2.3 for comparison.

2.3.2. Dose-Response Analysis—Adjustments and Extrapolation Methods

The EPA Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a) recommend that the 

method used to characterize and quantify cancer risk from a chemical be determined by what is 

known about the mode of action (MOA) of the carcinogen and the shape of the cancer 

dose-response curve. The linear approach is recommended when there are MOA data to indicate 

that the dose-response curve is expected to have a linear component below the POD or when the 

weight-of-evidence evaluation of all available data are insufficient to establish the MOA for a tumor 

site (U.S. EPA, 2005a). In the case of RDX, the mode of carcinogenic action for hepatocellular and 

alveolar/bronchiolar tumors is unknown (see discussion of Mechanistic Evidence in Section 1.2.5). 

Therefore, a linear low-dose extrapolation approach was used to estimate human carcinogenic risk 

associated with RDX exposure.

The survival curves were compared across dose groups in each study to determine whether 

time of death should be incorporated in the dose-response analysis of tumors. For female mice in 

Lish et al. (1984), the survival curves were determined to be similar across dose groups after the 

dose was reduced in the high-dose group to 100 mg/kg-day (log-rank test, p-value ≥0.10); 

therefore, a time-to-tumor analysis was not necessary for this study. Tumor incidence was

modeled using the multistage-cancer models in BMDS (versions 2.4 and 2.5). A standard BMR of 

10% ER was applied to both tumor sites in the mouse. 

Given the finding of an association between RDX exposure in the female mouse and 

increased tumor incidence at two tumor sites, basing the OSF on only one tumor site could 

potentially underestimate the carcinogenic potential of RDX. Therefore, an analysis that combines 

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1 the results from the mouse liver and lung tumor incidence is preferred. The MS-COMBO procedure

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(BMDS, version 2.5) extends the multistage-cancer models to the case with multiple tumors 

assuming independence between tumor types. There is no known biological relationship between 

liver and lung tumors in RDX-exposed mice, and therefore, as noted by the National Research 

Council (NRC, 1994), this assumption of independence is considered not likely to produce 

substantial error in risk estimates. MS-COMBO analyzes tumor incidence as present if either organ 

(or both) has a tumor and as absent otherwise. The procedure derives a maximum likelihood 

estimate of the combined risk at a 95% confidence level based on the parameter values obtained for 

the individual tumor multistage model fits. 

EPA’s preferred approach for extrapolating results from animal studies to humans is 

toxicokinetic modeling. As described in Appendix C, Section C.1.5, PBPK models for RDX in mice 

and humans published by Sweeney et al. (2012b) were evaluated and further developed by EPA. 

Consideration was given to whether the available toxicokinetic information supported using an 

internal dose metric derived by PBPK modeling. The available mechanistic data (Section 1.2.5) 

point to some evidence, although not conclusive, that RDX-generated metabolites may be 

implicated in the observed tumorigenicity in the female mouse. However, there are no data on the 

toxicokinetics of RDX metabolites, and metabolism in the liver is the only route of elimination of 

RDX in the PBPK model. In this case, as is to be expected from mass balance principles, the PBPK 

modeling provides no further information; the HED obtained from the model-estimated amount of 

total RDX metabolites scaled by BW3/4 was equal to that calculated using administered dose scaled 

by BW3/4. In addition to the lack of data on metabolism, other major uncertainties were identified 

in the mouse PBPK modeling; EPA’s evaluation of these uncertainties is discussed in Summary of 

Confidence in PBPK Models for RDX in Appendix C, Section C.1.5. Therefore, the PBPK model 

developed for the mouse was not used, and consistent with the EPA’s Guidelines for Carcinogen Risk 

Assessment (U.S. EPA, 2005a), the preferred approach for calculating an HED from the mouse 

tumors is adjustment of the administered dose by allometric scaling to achieve toxicological 

equivalence across species. 

As discussed in Section 2.1.1, the administered dose in animals was converted to an HED on 

the basis of (body weight)3/4 (U.S. EPA, 1992). This was accomplished by multiplying administered 

dose by (animal body weight in kg/human body weight in kg)1/4 (U.S. EPA, 1992), where the body 

weight for the mouse is 0.035 kg and the reference body weight for humans is 70 kg (U.S. EPA, 

1988). Details of the BMD modeling can be found in Appendix D, Section D.2. 

2.3.3. Derivation of the Oral Slope Factor

The lifetime cancer OSF for humans is defined as the slope of the line from the BMR (10% 

ER) at the BMDL to the estimated control response at zero (OSF = 0.1/BMDL10-HED). This slope, a 

95% upper confidence limit on the true slope, represents a plausible upper bound on the true slope 

or risk per unit dose. The PODs estimated for each mouse tumor site are summarized in Table 2-7. 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 Using linear extrapolation from the BMDL10-HED, human equivalent OSFs were derived for each 

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tumor site individually and both sites combined and are listed in Table 2-7. 

Table 2-7. Model predictions and OSFs for hepatocellular and alveolar/

bronchiolar adenomas or carcinomas in female B6C3F1 mice administered 

RDX in the diet for 2 years (Lish et al., 1984) 

Tumor type

Selected 

model BMR

BMD

(mg/kg-d)

BMDL

(mg/kg-d)

BMD10-HEDa

(mg/kg-d)

POD = 

BMDL10-HEDb

(mg/kg-d)

OSFc

(mg/kg-d)−1

Hepatocellular 

adenomas or 

carcinomasd

Multistage 1° 10% ER 64.2e 32.6e 9.56 4.89 0.020

Alveolar/bronchiolar 

adenomas or 

carcinomas

Multistage 1° 10% ER 52.8 27.7 7.92 4.16 0.024

Liver + lung tumors Multistage 1°

(MS-COMBO)

10% ER 29.0e 17.7e 4.35 2.66 0.038

a

BMD10-HED = BMD10 × (BWa

1/4/BWh1/4), where BWa = 0.035 kg, and BWh = 70 kg. b

BMDL10-HED = BMDL10 × (BWa

1/4/BWh1/4), where BWa = 0.035 kg, and BWh = 70 kg. c

OSF = BMR/BMDL10-HED, where BMR = 0.1 (10% ER). d

Incidences of female mouse liver tumors from Lish et al. (1984) are those reported in the PWG reevaluation 

(Parker et al., 2006). e

Data for hepatocellular adenomas and carcinomas and for liver and lung tumors combined were remodeled using 

the original sample sizes provided in Lish et al. (1984), which were slightly different for two groups than those 

reported in Parker et al. (2006). The resulting BMDs and BMDLs from the remodeling were 64.8 and 

32.8 mg/kg-day, respectively, for hepatocellular adenomas and carcinomas and 29.1 and 17.7 mg/kg-day, 

respectively, for liver and lung tumors combined. See Table D-23 and the subsequent MS-COMBO results for 

details. 

An OSF was derived from the BMDL10-HED based on a significantly increased trend in the 

incidence of hepatocellular and alveolar/bronchiolar adenomas or carcinomas in female B6C3F1

mice (i.e., the Liver + Lung BMDL10-HED from MS-COMBO). The OSF of 0.04 (mg/kg-day)−1 is 

calculated by dividing the BMR (10% ER) by the Liver + Lung BMDL10-HED and represents an upper 

bound on cancer risk per unit dose associated with a continuous lifetime exposure:

OSF = 0.10 ÷ (Liver + Lung) BMDL10-HED = 0.10 ÷ 2.66 mg/kg-day 

 = 3.8 × 10−2 (mg/kg-day)−1 

 = 4 × 10−2 (mg/kg-day)−1, rounded to one significant figure

The slope of the linear extrapolation from the central estimate of exposure associated with 

10% extra cancer risk (BMD10-HED) from the same data sets is given by: 

This document is a draft for review purposes only and does not constitute Agency policy.

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Slope of the linear extrapolation from the central estimate 

 = 0.10 ÷ (Liver + Lung) BMD10-HED = 0.10 ÷ 4.35 mg/kg-day 

 = 2.3 × 10−2 (mg/kg-day)−1

 = 2 × 10−2 (mg/kg-day)−1 (rounded to one significant figure)

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because above this level, the fitted dose-response model better characterizes what is known about 

the carcinogenicity of RDX. 

2.3.4. Uncertainties in the Derivation of the Oral Slope Factor 

A number of uncertainties underlie the cancer unit risk for RDX. Table 2-8 summarizes the 

impact on the assessment of issues such as the use of models and extrapolation approaches 

particularly those underlying the Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), the 

effect of reasonable alternatives, the approach selected, and its justification. 

Table 2-8. Summary of uncertainty in the derivation of the cancer risk value 

for RDX

Consideration and impact on 

cancer risk value Decision Justification

Selection of study

The cancer bioassay in the rat 

(Levine et al., 1983b) would provide 

a lower estimate of the OSF 

Lish et al. (1984) as 

principal oral study to 

derive the human 

cancer risk estimate

Lish et al. (1984) was a well-conducted study; 

five dose levels (including control) used, with a 

sufficient number of animals per dose group (at

terminal sacrifice, n = 62−65 female mice/dose 

group except highest dose where n = 31). 

Tumor data from the mouse provided a 

stronger basis for estimating the OSF than rat 

data. Confidence in the OSF based on rat data 

was low because of the small numbers of 

tumors.

Species/sex

Use of data sets from the male 

mouse or male rat would provide a 

lower OSF

OSF based on tumors 

in female B6C3F1

mouse 

It is assumed that a positive tumor response in 

animal cancer studies indicates that the agent 

can have carcinogenic potential in humans in 

the absence of data indicating that animal 

tumors are not relevant to humans (U.S. EPA, 

2005a). As there are no data to inform 

whether the response in any given 

experimental animal species or sex would be 

most relevant for extrapolating to humans, 

tumor data from the most sensitive species and 

sex were selected as the basis for the OSF. 

Other data sets would provide smaller OSF 

values, and are not considered any more or less 

relevant to humans than data from the female 

mouse (i.e., 0.017 per mg/kg-day based on 

hepatocellular carcinomas in male F344 rats, 

This document is a draft for review purposes only and does not constitute Agency policy.

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Consideration and impact on 

cancer risk value Decision Justification

and 0.018 per mg/kg-day based on 

alveolar/bronchiolor carcinomas in male 

B6C3F1 mice; see Appendix D, Section D.2).

Combined tumor types

Human risk would ↓ if OSF was 

based on analysis using only a 

single tumor type

OSF based on liver and 

lung tumors in female 

B6C3F1 mouse 

Basing the OSF on one tumor site could 

potentially underestimate the carcinogenic 

potential of RDX, so an analysis that included 

data from the two tumor sites was chosen to 

calculate the combined risk. Because there is 

no known biological dependence between the 

liver and lung tumors, independence between 

the two tumor sites was assumed. This is not 

likely to produce substantial error in the risk 

estimates (NRC, 1994).

Selection of dose metric

PBPK models are available for the 

rat, mouse, and human, and using 

an appropriate internal metric can 

↑ accuracy in human extrapolation 

Mouse liver and lung 

tumors: administered 

dose used

EPA evaluated a published PBPK model in the 

mouse (Sweeney et al., 2012b); major 

uncertainties associated with limited 

toxicokinetic data in the mouse and unknown 

differences in metabolism across species were 

identified. Althought EPA’s preferred approach 

for extrapolating results from animal studies to 

humans is toxicokinetic modeling, the 

uncertainties associated with use of the mouse 

PBPK model for RDX were considered higher 

than use of administered dose. 

Cross-species scaling

Alternatives could ↓ or ↑ OSF 

(e.g., 3.5-fold ↓ [scaling by body 

weight] or ↑ 2-fold [scaling by 

BW2/3])

BW3/4 scaling (default 

approach)

There are no data to support alternatives. 

Because the dose metric was not an AUC, BW3/4

scaling was used to calculate equivalent 

cumulative exposures for estimating equivalent 

human risks. While the true human 

correspondence is unknown, this overall 

approach is not expected over- or 

underestimate human equivalent risks. 

BMD model uncertainty 

Alternative models could ↓ or ↑ 

OSF

Use multistage model 

to derive a BMD and 

BMDL for combined 

tumor incidence

No biologically based models for RDX are 

available, and there is no a priori basis for 

selecting a model other than the multistage. 

The multistage model has biological support 

and is the model most consistently used in EPA 

cancer assessments (Gehlhaus et al., 2011). 

Low-dose extrapolation approach 

↓ cancer risk would be expected 

with the application of nonlinear 

extrapolation

Linear extrapolation 

from the POD

Where the available information is insufficient 

to establish the MOA for tumors at a given site, 

linear extrapolation is recommended because 

this extrapolation approach is generally 

considered to be health-protective (U.S. EPA, 

2005a). Because the MOA for RDX-induced 

liver and lung tumors has not been established, 

linear low-dose extrapolation was applied,

consistent with EPA guidance. 

Statistical uncertainty at the POD

↓ OSF by 1.6-fold if BMD used as 

the POD rather than the BMDL

BMDL (default 

approach for 

Lower bound is 95% confidence interval (CI) on 

administered exposure at 10% ER of liver and 

lung tumors.

This document is a draft for review purposes only and does not constitute Agency policy.

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Consideration and impact on 

cancer risk value Decision Justification

calculating plausible 

upper bound OSF)

Sensitive subpopulations 

↑ OSF to an unknown extent

Considered

qualitatively

There is little information on whether some 

subpopulations may be more or less sensitive 

to the potential carcinogenicity of RDX (i.e., 

because of variability in toxicokinetics or 

toxicodynamics for RDX). The mode of 

carcinogenic action for liver and lung tumors in 

experimental animals is unknown, and little 

information is available on RDX metabolites or 

variation in metabolic rates that could be used 

to evaluate human variability in cancer 

response to RDX.

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2.3.5. Previous IRIS Assessment: Oral Slope Factor 

The previous cancer assessment for RDX was posted to the IRIS database in 1990. The OSF

in the previous cancer assessment was based on the bioassay by Lish et al. (1984) and analysis of 

data for hepatocellular adenomas or carcinomas in female mice. An OSF of 1.1 × 10−1 (mg/kg-day)−1 

was derived using a linearized multistage procedure (extra risk) and scaling by body weight to the 

2/3 power for cross-species extrapolation. In addition, the previous assessment dropped the highdose group because the dose was reduced at week 11 to address high mortality. 

2.4. INHALATION UNIT RISK FOR CANCER

The carcinogenicity assessment provides information on the carcinogenic hazard potential 

of the substance in question, and quantitative estimates of risk from oral and inhalation exposure 

may be derived. Quantitative risk estimates may be derived from the application of a low-dose 

extrapolation procedure. If derived, the inhalation unit risk (IUR) is a plausible upper bound on the 

estimate of risk per μg/m3 air breathed.

An IUR value was not calculated because inhalation carcinogenicity data for RDX are not 

available. While inhalation absorption of RDX particulates is a plausible route of exposure, there 

are no toxicokinetic studies of RDX inhalation absorption to support an inhalation model. 

Therefore, a PBPK model for inhaled RDX was not developed to support route-to-route 

extrapolation of an IUR from the OSF.

2.5. APPLICATION OF AGE-DEPENDENT ADJUSTMENT FACTORS

As discussed in the Supplemental Guidance for Assessing Susceptibility from Early-Life 

Exposure to Carcinogens (U.S. EPA, 2005b), either default or chemical-specific age-dependent 

adjustment factors (ADAFs) are recommended to account for early-life exposure to carcinogens 

that act through a mutagenic MOA. Because no chemical-specific data on lifestage susceptibility for 

This document is a draft for review purposes only and does not constitute Agency policy.

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1 RDX carcinogenicity are available, and because the MOA for RDX carcinogenicity is not known (see 

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Section 1.2.5), application of ADAFs is not recommended. 

This document is a draft for review purposes only and does not constitute Agency policy.

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This document is a draft for review purposes only and does not constitute Agency policy.

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This document is a draft for review purposes only and does not constitute Agency policy.

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1 West, RR; Stafford, DA. (1997). Occupational exposures and haematological abnormalities among 

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This document is a draft for review purposes only and does not constitute Agency policy.

R-13 DRAFT—DO NOT CITE OR QUOTE