DRAFT DELIBERATIVE: FOR INTERAGENCY REVIEW ONLY. 

DO NOT DISTRIBUTE OUTSIDE YOUR AGENCY. 

 EPA/635/R-14/302

Interagency Review 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)

September 2014

NOTICE

This document is an Interagency Review draft. This information is distributed solely for the 

purpose of pre-dissemination peer review under applicable information quality guidelines. It has 

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

represent any Agency determination or policy. It is being 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
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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.
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CONTENTS

AUTHORS | CONTRIBUTORS | REVIEWERS................................................................................................. viii

PREFACE........................................................................................................................................................ x

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

1.1.1. Nervous System Effects..................................................................................................... 1-1

1.1.2. Kidney and Other Urogenital System Effects.................................................................... 1-3

1.1.3. Reproductive and Developmental Effects....................................................................... 1-19

1.1.4. Liver Effects..................................................................................................................... 1-30

1.1.5. Carcinogenicity ................................................................................................................ 1-41

1.1.6. Other Toxicological Effects.............................................................................................. 1-49

1.2.INTEGRATION AND EVALUATION ............................................................................................... 1-68

1.2.1. Effects Other Than Cancer............................................................................................... 1-68

1.2.2. Carcinogenicity ................................................................................................................ 1-70

1.2.3. Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes.............. 1-71

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

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

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

2.1.2. Methods of Analysis.......................................................................................................... 2-3

2.1.3. Derivation of Candidate Values......................................................................................... 2-7

2.1.4. Derivation of Organ/System-Specific Reference Doses.................................................. 2-11

2.1.5. Selection of the Proposed Overall Reference Dose ........................................................ 2-12

2.1.6. Uncertainties in the Derivation of Reference Dose......................................................... 2-12

2.1.7. Confidence Statement..................................................................................................... 2-13

2.1.8. Previous IRIS Assessment................................................................................................ 2-13

2.2.INHALATION REFERENCE CONCENTRATION FOR EFFECTS OTHER THAN CANCER..................... 2-13

2.2.1. Previous IRIS Assessment................................................................................................ 2-14
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2.3.ORAL SLOPE FACTOR FOR CANCER............................................................................................. 2-14

2.3.1. Analysis of Carcinogenicity Data ..................................................................................... 2-14

2.3.2. Dose-Response Analysis—Adjustments and Extrapolations Methods........................... 2-16

2.3.3. Derivation of the Oral Slope Factor................................................................................. 2-18

2.3.4. Uncertainties in the Derivation of the Oral Slope Factor................................................ 2-19

2.3.5. Previous IRIS Assessment: Oral Slope Factor ................................................................. 2-20

2.4.INHALATION UNIT RISK FOR CANCER ......................................................................................... 2-21

2.5.APPLICATION OF AGE-DEPENDENT ADJUSTMENT FACTORS...................................................... 2-21

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

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

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

Table LS-1. Overview of the search strategy employed for RDX ............................................................ LS-2

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

conduct, or reporting..................................................................................................... LS-5

Table LS-3. Experimental animal studies considered less informative because of certain study 

design, conduct, or reporting limitations.................................................................... LS-11

Table 1-1. Evidence pertaining to nervous system effects in humans .....................................................1-5

Table 1-2. Evidence pertaining to nervous system effects in animals......................................................1-6

Table 1-3. Evidence pertaining to kidney effects in humans....................................................................1-6

Table 1-4. Evidence pertaining to kidney and other urogenital system effects in animals......................1-7

Table 1-5. Six-, 12-, and 24-month incidence of kidney endpoints in male F344 rats reported for 

statistical evaluation in Levine et al. (1983) ..................................................................1-12

Table 1-6. Six-, 12-, and 24-month incidence of urinary bladder endpoints in male F344 rats 

reported for statistical evaluation in Levine et al. (1983)..............................................1-13

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

for statistical evaluation in Levine et al. (1983).............................................................1-14

Table 1-8. Evidence pertaining to reproductive and developmental effects in animals........................1-21

Table 1-9. Evidence pertaining to male reproductive effects in animals ...............................................1-26

Table 1-10. Evidence pertaining to liver effects in humans....................................................................1-33

Table 1-11. Evidence pertaining to liver effects in animals....................................................................1-34

Table 1-12. Liver tumors observed in chronic animal bioassays ............................................................1-43

Table 1-13. Lung tumors observed in chronic animal bioassays.............................................................1-46

Table 1-14. Evidence pertaining to systemic effects (hematological) in humans ..................................1-52

Table 1-15. Evidence pertaining to systemic effects in animals.............................................................1-54

Table 2-1. Summary of derivation of PODs following oral exposure to RDX............................................2-5

Table 2-2. Effects and corresponding derivation of candidate values......................................................2-9

Table 2-3. Organ/system-specific RfDs and proposed overall RfD for RDX............................................2-11

Table 2-4. Incidence of hepatocellular and alveolar/bronchiolar tumors in female B6C3F1 mice 

administered RDX for 2 years in diet.............................................................................2-16

Table 2-5. Model predictions and oral slope factors for hepatocellular and alveolar/bronchiolar 

adenomas or carcinomas in female B6C3F1 mice administered RDX in the diet 

for 2 years (Lish et al., 1984a)........................................................................................2-18

Table 2-6. Summary of uncertainty in the derivation of the cancer risk value for RDX .........................2-19

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. ....................1-12

Figure 1-2. Exposure-response array of kidney and urogenital system effects......................................1-16

Figure 1-3. Exposure response array of reproductive and developmental effects following oral 

exposure. .......................................................................................................................1-25

Figure 1-4. Exposure response array of male reproductive effects following oral exposure.................1-29
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Figure 1-5. Exposure response array of liver effects following oral exposure........................................ 1-40

Figure 2-1. Approach for dose-response analysis..................................................................................... 2-3

Figure 2-2. Candidate values with corresponding POD and composite UF ............................................ 2-10
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ABBREVIATIONS

AAP Army ammunition plants

ACGIH American Conference of Governmental 

Industrial Hygienists 

AChE acetylcholinesterase

ADAF age-dependent adjustment factor

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

BMC benchmark concentration

BMCL benchmark concentration lower 

confidence limit

BMD benchmark dose 

BMDL benchmark dose lower confidence limit 

BMDS Benchmark Dose Software 

BMR benchmark response 

BUN blood urea nitrogen 

BW body weight 

CASRN Chemical Abstracts Service Registry 

Number 

CCL Contaminant Candidate List

CI confidence interval

CNS central nervous system 

CYP450 cytochrome P450 

DAF dosimetric adjustment factor 

DMSO dimethylsulfoxide 

DNA deoxyribonucleic acid 

DTIC Defense Technical Information Center

EPA Environmental Protection Agency 

ER extra risk

FDA Food and Drug Administration 

FOB functional observational battery

GABA gamma amino butyric acid

GD gestational day

GLP good laboratory practices

HEC human equivalent concentration 

HED human equivalent dose 

HERO Health and Environmental Research 

Online

IARC International Agency for Research on 

Cancer

IOM Institute of Medicine

IRIS Integrated Risk Information System 

LDH lactate dehydrogenase

LOAEL lowest-observed-adverse-effect level 

LOD limit of detection

miRNA microRNA

MOA mode of action

NCEA National Center for Environmental 

Assessment 

NHANES National Health and Nutrition 

Examination Survey

NICNAS National Industrial Chemicals 

Notification and Assessment Scheme

NIOSH National Institute for Occupational 

Safety and Health

NOAEL no-observed-adverse-effect level 

NPL National Priorities List

NRC Nuclear Regulatory Commission

NTP National Toxicology Program 

OR odds ratio

ORD Office of Research and Development 

OSF oral slope factor

OSHA Occupational Safety and Health 

Administration

PBPK physiologically based pharmacokinetic 

POD point of departure 

POD[ADJ] duration-adjusted POD 

PWG Pathology Working Group

RBC red blood cell

RfC inhalation reference concentration 

RfD oral reference dose 

RNA ribonucleic acid 

SD Sprague-Dawley

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

TNT trinitrotoluene

TSCATS Toxic Substances Control Act Test 

Submissions

TWA time-weighted average 

U.S. United States of America

UCL upper confidence limit

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 cells

WOS Web of Science
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Assessment Team

Louis D’Amico, Ph.D. (Assessment Manager) 

Todd Blessinger, Ph.D.

Ravi Subramaniam, Ph.D.

Christopher Brinkerhoff, Ph.D.

AUTHORS | CONTRIBUTORS | REVIEWERS

U.S. EPA/ORD/NCEA

Washington, DC

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.

Karen Hogan, MS

Anne Loccisano, Ph.D.

Jordan Trecki, Ph.D.

Belinda Hawkins, Ph.D.

Scott Wesselkamper, Ph.D.

U.S. EPA/ORD/NCEA

Washington, DC

U.S. EPA/ORD/NCEA

Cincinnati, OH

Production Team

Maureen Johnson U.S. EPA/ORD/NCEA

Washington, DC

Contractor Support

Heather Carlson-Lynch, S.M., DABT

Julie Melia, Ph.D., DABT

Megan Riccardi, M.S. 

Pam Ross, M.S.

Robin Blain, Ph.D.

SRC, Inc, North Syracuse, NY

ICF International, Fairfax, VA

Executive Direction

Kenneth Olden, Ph.D., Sc.D., L.H.D. (Center Director)

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

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

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

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

Susan Rieth, MPH (Quantitative Modeling Branch Chief)

U.S. EPA/ORD/NCEA

Washington, DC 

Internal Review Team

General Toxicity/Cancer/Immunotoxicity Workgroup

Neurotoxicity Workgroup

Pharmacokinetics Workgroup

Reproductive and Developmental Toxicity Workgroup

Statistics Workgroup

U.S. EPA/ORD/NCEA

Washington, DC
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Toxicity Pathways Workgroup

Reviewers

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

Comments were submitted by:

Office of Children’s Health Protection, Washington, DC

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

Washington, DC

Office of Solid Waste and Emergency Response, Washington, DC

Office of Water, Washington, DC

Region 2, New York, NY

Region 8, Denver, CO

Federal Facilities Forum

1 
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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) in order to identify its adverse health effects and to characterize 

exposure-response relationships. It was prepared under the auspices of EPA’s 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, as well as derivation of an oral slope factor 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. 

Organ/system-specific RfDs are calculated based on data for applicable hazards, e.g., nervous

system toxicity. 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 cited and 

summarized in the Preamble to IRIS Toxicological Reviews. The findings of this assessment and 

related documents produced during its development are available on the IRIS web site 

(http://www.epa.gov/iris). Appendices for assessments by other health agencies, chemical and 

physical properties, toxicokinetic information, and summaries of supporting toxicity information

are provided as Supplemental Information to this assessment (See Appendices A to C).

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).

In April 2011, the National Research Council (NRC) released its Review of the Environmental 

Protection Agency’s Draft IRIS Assessment of Formaldehyde. In addition to offering comments 

specifically about EPA’s draft formaldehyde assessment, the NRC made several recommendations 

to EPA for improving the development of IRIS assessments. EPA agreed with the recommendations 

and is implementing them consistent with the Panel’s “Roadmap for Revision,” which viewed the 

full implementation of their recommendations by the IRIS Program as a multi-year process. 

In response to the NRC’s 2011 recommendations, the IRIS Program has made changes to 

streamline the assessment development process, improve transparency, and create efficiencies in 

the Program. The NRC has acknowledged EPA’s successes in this area. In May 2014, the NRC 

released their report Review of EPA’s Integrated Risk Information System Process reviewing the IRIS 
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1 assessment development process and found that EPA has made substantial improvements to the 

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IRIS Program in a short amount of time. 

The draft RDX assessment represents a significant advancement in implementing the NRC 

recommendations. This assessment is streamlined, and uses tables, figures, and appendices to 

increase transparency and clarity. It is structured to have distinct sections for the literature search 

and screening strategy, study selection and evaluation, hazard identification, and dose-response 

assessment. The assessment includes a comprehensive, systematic, and documented literature 

search and screening approach, provides the database search strategy in a table (databases, 

keywords), visually represents the inclusion and exclusion of studies in a flow diagram, and all of 

the references are integrated within the Health and Environmental Research Online (HERO) 

database. A study evaluation section provides a systematic review of methodological aspects of 

epidemiology and experimental animal studies, including study design, conduct, and reporting, that 

was subsequently taken into consideration in the evaluation and synthesis of data from these 

studies. The evidence is presented in standardized evidence tables, and exposure-response arrays. 

The hazard identification and dose-response sections include subsections based on organ/systemspecific effects in which the evidence is synthesized within and integrated across all evidence for 

each target organ/systems. 

In the draft RDX assessment, the IRIS Program has attempted to transparently and 

uniformly identify strengths and limitations that would affect interpretation of results. All human 

and animal studies of RDX that were considered to be of acceptable quality, whether yielding 

positive, negative, or null results, were considered in assessing the evidence for health effects 

associated with chronic exposure to RDX. These studies were evaluated for aspects of design, 

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

the synthesis of evidence for determination of human hazard potential using the study quality 

considerations outlined in the Preamble. A brief summary of the evaluation is included in the 

section on methods for study selection and evaluation. Information on study features related to this 

evaluation is reported in evidence tables and documented 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. 

In this assessment, the IRIS Program is using existing guidelines to systematically approach 

the integration of noncancer human, animal, and mechanistic evidence. In conducting this analysis 

and developing the synthesis, the IRIS Program evaluates the data for the: strength of the 

relationship between the exposure and response and the presence of a dose-response relationship; 

specificity of the response to chemical exposure and whether the exposure precedes the effect; 

consistency of the association between the chemical exposure and response; and biological 

plausibility of the response or effect and its relevance to humans. The IRIS Program uses this 

weight-of-evidence approach to identify the potential human hazards associated with chemical 

exposure.
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1 The IRIS RDX assessment provides a streamlined presentation of information, integrated 

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hazard identification of all toxic effects, and derivation of organ/system-specific reference values. 

Additionally, consistent with the goal that assessments should provide a scientifically sound and 

transparent evaluation of the relevant scientific literature and presentation of the analyses 

performed, this assessment contains an expanded discussion of study selection and evaluation, as 

well as increased documentation of key assessment decisions.

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

please contact EPA’s IRIS Hotline at 202-566-1676 (phone), 202-566-1749 (fax), or 

hotline.iris@epa.gov. 

Chemical Properties

RDX is a white, crystalline solid member of the nitramine class of organic nitrate explosives

(Boileau et al., 2005; Bingham et al., 2001). It is a synthetic chemical not found naturally in the 

environment. The solubility of RDX in water is poor, having been reported as 59.7 mg/L at 25°C

(Yalkowsky and He, 2003). The Henry’s law constant for RDX is approximately 2 × 10−11 atmm3/mole at 25°C, suggesting slow volatilization from water or moist soil (ATSDR, 2012). The 

normalized soil organic carbon/water partition coefficient (KOC) values for RDX range from 42 to 

167, indicating a potential for RDX to be mobile in soil (Spanggord et al., 1980). The vapor pressure 

of 4.10 × 10−9 mm Hg at 20°C suggests that it will exist as particulate matter in air and be removed 

by both wet and dry deposition. RDX degrades in the environment, and can be subject to both 

photolysis (Sikka et al., 1980; Spanggord et al., 1980) and biodegradation (Funk et al., 1993; 

Mccormick et al., 1981). Further information on the physical and chemical properties of RDX are 

provided in Appendix C, Section C.1. 

Uses and Environmental Occurrence

RDX is used primarily as a military explosive. In the United States, RDX is produced at Army 

ammunition plants (AAP) 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). According to the U.S. EPA Inventory Update 

Reporting program, the aggregated national production volume in 2006 was between 1 and 

10 million pounds. 

RDX releases have been reported into the air, water, or soil (ATSDR, 2012, 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., 1980). 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, 2012b). An extensive discussion of RDX properties and fate and transport 

is available in U.S. EPA (2012b). 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., 
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1 1991). RDX has also been detected in indoor air samples from military facilities where RDX is 

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produced (Bishop et al., 1988). 

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

or around military facilities where RDX is or was produced, stored, or used. Oral, inhalation, and 

dermal routes of exposure may be relevant. 

RDX has been 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 releases or 

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

States and its territories. The NPL list 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, although the total number of sites where RDX is present is unknown. 

RDX is 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 the Office of Water’s Drinking Water Contaminant 

Candidate Lists (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 values for RDX have been established by the Agency for Toxic Substances and 

Disease Registry (ATSDR), the American Conference of Governmental Industrial Hygienists 

(ACGIH), the Australian National Industrial Chemicals Notification and Assessment Scheme 

(NICNAS), the National Institute for Occupational Safety and Health (NIOSH), and the Occupational 

Safety and Health Administration (OSHA). These toxicity values and their basis are presented in 

Appendix A. It is important to recognize that the assessments performed by other health agencies 

may have been prepared for different purposes and may utilize different methods, and that newer 

studies may be included in the IRIS assessment. 
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PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS

1. Scope of the IRIS Program

1 Soon after the EPA was established in 

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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 

toxicologically related chemicals, or a 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 

assessment will meet their needs. Then a 

public meeting on problem formulation 

invites discussion of the key issues and the 

studies and analytical approaches that might 

contribute to their resolution.

Step 1. Development of a draft 

Toxicological Review. The draft 

assessment considers all pertinent 

publicly available studies and applies 
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1 consistent criteria to evaluate study 

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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.

Step 2. Internal review by scientists in EPA 

programs and regions. The draft 

assessment is revised to address the

comments from within the EPA.

Step 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. 

Step 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. 

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The peer review draft, written public 

comments, and peer review report 

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.
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3. Identifying and selecting 

pertinent studies

3.1. Identifying studies

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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 web site 

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 

peer-reviewed, the EPA will have it peerreviewed.

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)]: 

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– Studies of a sufficiently similar 

mixture. In evaluating similarity, the 

assessment considers the alteration 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. 
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3.3. Selecting pertinent experimental 

studies

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consideration for selecting pertinent 

experimental animal studies or human 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

40 After the subsets of pertinent 

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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 

documentation of each study, but not whether 

the results are positive, negative, or null. The 

objective is to identify the stronger, more 

informative studies based on a uniform 

evaluation of quality characteristics across 

studies of similar design.

4.1. Evaluating the quality of 

epidemiologic studies

The assessment evaluates design and 

methodological aspects that can increase or 

decrease the weight given to each 

epidemiologic 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 
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1 outcomes, and other factors) and other 

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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 outcome and is sufficiently 

prevalent 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.

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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 historical 

control data from the same 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 
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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.

As a check on the selection and evaluation 

of pertinent studies, the EPA asks peer 

reviewers to identify studies that were not 

adequately considered.

5. Evaluating the overall evidence of 

each effect

5.1. Concepts of causal inference

For each health effect, the assessment 

evaluates the evidence as a whole to 

determine whether it is reasonable to infer a 

causal association between exposure to the 

agent and the occurrence of the effect. This 

inference is based on information from 

pertinent human studies, animal studies, and 

mechanistic studies of adequate quality. 

Positive, negative, and null results are given 

weight according to study quality.

Causal inference involves scientific 

judgment, and the considerations are nuanced 

and complex. Several health agencies have 

developed frameworks for causal inference, 

among them the U.S. Surgeon General (CDC, 

2004; HEW, 1964), the International Agency 

for Research on Cancer (IARC, 2006), the 

Institute of Medicine (IOM, 2008), and the EPA 

(2010, §1.6; 2005a, §2.5). Although developed 

for different purposes, the frameworks are 

similar in nature and provide an established 

structure and language for causal inference. 

Each considers aspects of an association that 

suggest causation, discussed by Hill (1965) 

and elaborated by Rothman and Greenland

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

1994, Appendix C). 

Strength 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 

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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.

Consistency of association: An inference of 

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 
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1 be found in other lines of evidence, such as 

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changing disease patterns in the 

population.

“Natural experiments”:A change in exposure 

that brings about a change in disease 

frequency provides strong 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.

47 Sufficient epidemiologic evidence of an 

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association consistent with causation:

The evidence establishes a causal 

association for which alternative 

explanations such as chance, bias, and 

confounding can be ruled out with 

reasonable confidence.

Suggestive 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.

Inadequate epidemiologic evidence to infer 

a causal association: The available 

studies do not permit a conclusion 

regarding the presence or absence of an 

association.

Epidemiologic 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.

5.3. Evaluating evidence in animals

For each effect, the assessment evaluates 

the evidence from the animal experiments as a 

whole to determine the extent to which they 

indicate a potential for effects in humans. 

Consistent results across various species and 

strains increase confidence that similar results 

would occur in humans. Several concepts 

discussed by Hill (1965) are pertinent to the 

weight of experimental results: consistency of 

response, dose-response relationships, 

strength of response, biologic plausibility, and 

coherence (U.S. EPA, 2005a, §2.2.1.7; 

1994, Appendix C). 

In weighing evidence from multiple 

experiments, U.S. EPA (2005a, §2.5) 

distinguishes: 

Conflicting evidence (that is, mixed positive 

and negative results in the same sex and 

strain using a similar study protocol) from 
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1 Differing results (that is, positive results and 

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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 

possible, mechanistic pathways that lead to a 

health 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 processes that lead to a 

health effect at this or another site 

(also known as a mode of action).

For each effect, the assessment discusses 

the available information on its modes of 

action and associated key events (key events

being empirically observable, necessary 

precursor steps or biologic markers of such 

steps; mode of action being a series of key 

events involving interaction with cells, 

operational and anatomic changes, and 

resulting in disease). Pertinent information 

may also come from studies of metabolites or 

of compounds that are structurally similar or 

that act through similar mechanisms. 

Information on mode of action is not required 

for a conclusion that the agent is causally 

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

The assessment addresses several 

questions about each hypothesized mode of 

action (U.S. EPA, 2005a, §2.4.3.4). 

1) 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 
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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 

similar properties and structure-activity 

49 relationships to mutagenic carcinogens (U.S. 

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EPA, 2005a ,§2.3.5). 

5.5. Characterizing the overall weight 

of the evidence

After evaluating the human, animal, and 

mechanistic evidence pertinent to an effect, 

the assessment answers the question: Does 

the agent cause the adverse effect? (NRC, 

2009, 1983). In doing this, the assessment 

develops a narrative that integrates the 

evidence pertinent to causation. To provide 

clarity and consistency, the narrative includes 

a standard hazard descriptor. For example, 

the following standard descriptors combine 

epidemiologic, experimental, and mechanistic 

evidence of carcinogenicity (U.S. EPA, 2005a, 

§2.5). 

Carcinogenic 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.

Likely 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.

Suggestive 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.
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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 

is 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, §1.6).

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 co-exposures are 

difficult to address or other lines of 

evidence are limited or inconsistent; or 

multiple animal studies from different 

laboratories demonstrate effects and 

there are limited or no human data.

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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 

toxicokinetic model can be used to 

extrapolate across exposure routes.
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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 

exposure level near the lower end of the 

observed range, without significant 

43 extrapolation to lower doses) (Sections 7.2-

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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 

standard approaches that can be applied to 

typical data sets. These standard approaches 

also facilitate comparison across exposure 

patterns and species.
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1 – Intermittent study exposures are 

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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 

require sufficient data to ascertain a mode of 

action and to quantitatively support model 

parameters associated with its key events. 

Because different models may provide 

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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, 2012a). 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, 2012a; 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.

– For dichotomous data, a response level 

of 10% extra risk is generally used for 
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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 

bound on the dose associated with the 

selected response level.

7.4. Extrapolating to lower doses and 

response levels

The purpose of extrapolating to lower 

doses is to estimate responses at exposures 

below the observed data. Low-dose 

extrapolation, typically used for cancer data, 

considers what is known about modes of 

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

1) 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. 

The result of linear extrapolation is 

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). 

1) If an epidemiologic or experimental study 

reports quantitative results for a 

susceptible population or lifestage, these 

data are analyzed to derive separate 

toxicity values for susceptible individuals.
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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 developmental toxicity, 

reproductive toxicity, and neurotoxicity there 

is guidance on adverse effects and their 

biologic markers (U.S. EPA, 1998, 1996, 1991). 

To account for uncertainty and variability 

in the derivation of a lifetime human exposure 

where adverse effects are not anticipated to 

occur, reference values are calculated by 

applying a series of uncertainty factors to the 

point of departure. If a point of departure 

cannot be derived by modeling, a noobserved-adverse-effect level or a lowestobserved-adverse-effect level is used instead. 

The assessment discusses scientific 

considerations involving several areas of 

variability or uncertainty.

Human 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 non-susceptible 

individuals) (U.S. EPA, 2002, §4.4.5; 

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

1991, §3.4). 

Animal-to-human extrapolation. If animal 

results are used to make inferences about 

humans, the assessment adjusts for crossspecies 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; 

2002, §4.4.5; 1998, §4.2; 1996, §4; 

1994, §4.3.9.1; 1991, §3.4). 

Adverse-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 
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xxviii DRAFT—DO NOT CITE OR QUOTE

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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). 

Subchronic-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). 

Incomplete 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 twogeneration reproduction study are 

missing and a factor of 101/2 if either is 

missing (U.S. 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 

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

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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/systemspecific 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, 

except for mechanistic studies that might 

affect the interpretation of prior test 

results. 

Medium confidence: This is a matter of 

judgment, between high and low 

confidence.
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1 Low confidence: The reference value is 

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

EXECUTIVE SUMMARY

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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 ingestion of crops irrigated with 

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

settings. 

Epidemiological studies provide only limited information on occupational 

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

nervous system following acute exposure to RDX. Animal studies 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

EPA identified nervous system effects as a human hazard of RDX exposure. Several human 

case reports and animal studies provide consistent evidence of associations between RDX exposure

and effects on the nervous system, including seizures or convulsions. Increased mortality was 

generally observed at RDX doses that induced nervous system effects, and several studies 

documented that deaths in most cases were preceded by tremors and convulsions. Although 

mechanistic data are insufficient to establish a mode of action (MOA) for RDX-induced convulsions, 

the available information suggests that nervous system effects are mediated by RDX binding to the 

picrotoxin convulsant site of the GABAA channel, resulting in disinhibition that leads to the onset of 

seizures. 

EPA identified kidney and other urogenital effects as a potential human hazard of RDX 

exposure based on observations in 2-year 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 

RDX-related urogenital effects. There is no established MOA for RDX-related effects on the 

urogenital system.

Based on the finding of testicular degeneration in male mice exposed to RDX in diet for 2 

years, in the only mouse study conducted of that duration, EPA identified suggestive evidence of 
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1 male reproductive effects as a potential human hazard of RDX exposure. There is no known MOA

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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. EPA concluded that the evidence does 

not support effects on other organs/systems, including liver and developmental effects, as a 

potential human hazard of RDX exposure. 

Oral Reference Dose (RfD) for Effects Other Than Cancer 

Organ-specific RfDs were derived for hazards associated with RDX exposure (see 

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 proposed overall RfD for RDX

Effect Basis RfD (mg/kg-day)

Study exposure 

description Confidence

Nervous system Convulsions 9 × 10−4 Subchronic Medium

Kidney/urogenital Suppurative prostatitis 2 × 10−3 Chronic Low

Male reproductive Testicular degeneration 2 × 10−2 Chronic Low

Proposed overall RfD Nervous system effects 9 × 10−4 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 it represents the most sensitive human hazard of RDX 

exposure. Incidence of seizures or convulsions as observed in a subchronic gavage study (Crouse et 

al., 2006) was selected for derivation of the overall RfD as the study was well-conducted, utilized a 

more pure form of test material 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; this is supported by the 

observation in Crouse et al. (2006) that for all the dose groups where unscheduled deaths were 

recorded, mortality was strongly associated with convulsions. A physiologically-based 

pharmacokinetic (PBPK) model was used to extrapolate the BMDL01 to a human equivalent dose 

(HED) based on RDX arterial blood concentration, which was then used for RfD derivation.

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

effects by a composite uncertainty factor of 300 to account for extrapolation from animals to 
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1 humans (3), interindividual differences in human susceptibility (10), extrapolation of results from a 

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subchronic study to a chronic study (3), and deficiencies in the toxicity database (3).

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

Critical effect Point of departure* UF Chronic RfD

Nervous system effects (convulsions)

90-d F344 rat study

Crouse et al. (2006)

BMDL01-HED: 1.3 mg/kg-d 300 9 × 10−4 mg/kg-d

*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 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. Additionally, 

a PBPK model for inhaled RDX is not available to support route-to-route extrapolation from the RfD. 

Evidence for Human Carcinogenicity

Under EPA’s Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), the database for 

RDX provides “suggestive evidence of carcinogenic potential” based on the finding of statistically 

significant trends for hepatocellular adenomas or carcinomas and alveolar/bronchiolar adenomas 

or carcinomas in female, but not male, B6C3F1 mice (Lish et al., 1984). This is further supported by 

the finding of a statistically significant trend for hepatocellular carcinomas in male, but not female, 

F344 rats (Levine et al., 1983) exposed to RDX in the diet for two years. On the other hand, there 

was no evidence of carcinogenicity in Sprague-Dawley rats in a 2-year dietary study of RDX (Hart, 

1976). No human studies are available to assess the carcinogenic potential of RDX. The MOA for 

liver and lung tumors in experimental animals is not known. Available in vitro and in vivo 

genotoxicity assays were largely negative for RDX, suggesting that parent RDX does not interact 
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1 directly with DNA. N-nitroso metabolites of RDX generated anaerobically have tested positive in 

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some genotoxicity assays; their contribution to the overall carcinogenic potential of RDX is not 

known. 

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 two-year dietary study was generally well conducted, with four dose groups and

adequate numbers of animals per dose group (85/sex/group, with interim sacrifices of 

10/sex/group at 6 and 12 months), and included detailed reporting of methods and results 

(including individual animal data). The initial high dose (175 mg/kg-day) was reduced to 

100 mg/kg-day at week 11 due to high mortality. 

Although EPA concluded that there is “suggestive evidence of carcinogenic potential” for 

RDX, the Agency determined that quantitative analysis of the mouse tumor data may be useful for 

providing a sense of the magnitude of potential carcinogenic risk. 

EPA calculated a single oral slope factor (OSF) that considered the combination of tumors. 

Point of departure (i.e., BMD and BMDL) estimates that corresponded to a specific risk of incidence 

of either of the tumors (liver or lung) were calculated. The single BMDL10 so derived from the 

mouse tumors was extrapolated to the HED using BW3/4 scaling, and an OSF was derived by linear 

extrapolation from the BMDL10-HED. The OSF is 4 × 10−2 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

The carcinogenic potential of RDX by inhalation has not been investigated. A PBPK model to 

support route-to-route extrapolation of an inhalation unit risk based on oral carcinogenicity data

was not available. 

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. Data to suggest 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 than in females. Some evidence 

suggests 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, 
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1 there is no information to indicate how genetic polymorphisms may affect susceptibility to RDX. 

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Key Issues Addressed in Assessment

In most instances, the spectrum of effects associated with chemical exposure will range in 

severity, with relatively less severe effects generally occurring at doses lower than those associated 

with more severe or “frank” 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 (2012a) emphasizes that when modeling a dose-response relationship from a given set of data, 

statistical and biological characteristics of the dataset must be considered, including consideration

of the severity of the effect. For convulsions, because of the severity of the effect itself and the 

strong association with mortality, a benchmark response (BMR) level of 1% was selected for 

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

of the effect. Use of a BMR of 1% extra risk of convulsions resulted in extrapolation below the range 

of experimental data and could potentially increase uncertainty in the BMD and BMDL values. 

The candidate RfD for kidney and other urogenital effects is based on suppurative 

prostatitis. This organ/system-specific RfD is based on a dose-related increase in suppurative 

prostatitis as reported in a 2-year feeding study in male F344 rats (Levine et al., 1983), the only 

2-year study in rats that examined the prostate. Some reports have hypothesized that the observed 

suppurative prostatitis was secondary to a bacterial infection unrelated to RDX toxicity (ATSDR, 

2012; Sweeney et al., 2012a; Crouse et al., 2006). In reviewing the findings in Levine et al. (1983), 

EPA concluded that while an opportunistic bacterial infection may have been the proximal cause of 

the suppurative prostatitis, the infection was 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. (1983). 
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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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 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 health effects of RDX, and remaining references were sorted into categories for further 

evaluation. 

The literature search for RDX was conducted through January 2014 using the databases and 

general keywords listed in Table LS-1 (see Appendix B for further details of the literature search 

strategy). More specifically, the literature search for RDX was conducted in four online scientific 

databases—Pubmed, Toxline, Toxcenter, and TSCATS. The detailed search approach for these 

databases, including the search strings and number of citations identified per database, is provided 

in Appendix B, Table B-1. Given the military applications of RDX, 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. This search of the five online 

databases identified 995 citations (after electronically eliminating duplicates). The computerized 

database searches were supplemented by review of online regulatory sources, “forward” and 

“backward” searches of Web of Science (Appendix B, Table B-3), as well as additional references 

added during development of the toxicological review (including guidance documents and other 

references that provide context for evaluating RDX health effects); 113 citations were obtained 

using these additional search strategies. In total, 1108 citations were identified using online 

scientific databases and additional search strategies.

EPA requested public submissions of additional information in 2010 (75 FR 76982;

December 10, 2010); no submissions were received in response to this call for data. Additionally, 

EPA 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. 
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Table LS-1. Overview of the search strategy employed for RDX

Database Keywords

Pubmed

Toxline

TSCATS1

Toxcenter

DTIC

WOS (forward and backward 

search only)

Chemical CASRN: 121-82-4

Synonyms: Cyclonite OR RDX OR Cyclotrimethylenetrinitramine OR 

"cyclotrimethylene trinitramine" OR "Hexahydro-1,3,5-trinitro-1,3,5-triazine" 

OR "Hexahydro-1,3,5-trinitro-s-triazine" OR Hexogen OR "1,3,5-trinitro-1,3,5-

triazine" OR "1,3,5-Triaza-1,3,5-trinitrocyclohexane" OR "1,3,5-Trinitro-1,3,5-

triazacyclohexane" OR "1,3,5-Trinitrohexahydro-1,3,5-triazine" OR "1,3,5-

Trinitrohexahydro-s-triazine" OR "1,3,5-Trinitroperhydro-1,3,5-triazine" OR 

"Esaidro-1,3,5-trinitro-1,3,5-triazina" OR "Hexahydro-1,3,5-trinitro-1,3,5-

triazin" OR "Perhydro-1,3,5-trinitro-1,3,5-triazine" OR 

Cyclotrimethylenenitramine OR Trimethylenetrinitramine OR "Trimethylene 

trinitramine" OR Trimethyleentrinitramine OR "Trinitrocyclotrimethylene 

triamine" OR Trinitrotrimethylenetriamine OR "CX 84A" OR Cyklonit OR 

Geksogen OR Heksogen OR Hexogeen OR Hexolite OR "KHP 281" OR "PBX (af) 

108" OR "PBXW 108(E)" OR "Pbx(AF) 108"

Synonym and CASRN search for all databases; Toxcenter, Pubmed, and WOS 

limited using toxicity-related keywords

Toxicity-related terms (see Appendix B for specific keywords) 

Toxicity (including duration, effects to children and occupational exposure); 

development; reproduction; teratogenicity; exposure routes; 

pharmacokinetics; toxicokinetics; metabolism; body fluids; endocrinology; 

carcinogenicity; genotoxicity; antagonists; inhibitors

ChemID 

TSCATS 2 & 8e submissions

Searched by CASRN 

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 

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

in Figure LS-1.1

• 21 references were identified as potential sources of health effects data and were 

considered for data extraction to evidence tables and exposure-response arrays. 

• 65 references were identified as supporting studies; these included 16 studies describing 

physiologically-based pharmacokinetic (PBPK) models and other toxicokinetic information, 

25 studies providing genotoxicity and other mechanistic information, 7 acute toxicity 

studies, and 17 human case reports. Studies investigating the effects of acute exposures and 

 

1 Studies were assigned (or “tagged”) to a given category in HERO that best reflected the primary content of 

the study. 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 Supporting 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. 
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case reports are generally less pertinent for characterizing health hazards associated with 

chronic oral and inhalation exposure. Therefore, information from these studies was not 

considered for extraction into evidence tables. Nevertheless, these studies were still 

evaluated as possible sources of supporting health effects information. 

• 277 references were identified as secondary sources of health effects information (e.g., 

reviews and other agency assessments) or as studies providing potentially useful contextual 

information (e.g., studies providing information on exposure levels); these references were 

kept as additional resources for development of the Toxicological Review. 

• 745 references were identified as not being pertinent to an evaluation of the health effects 

of RDX and were excluded from further consideration (see Figure LS-1 for exclusion 

categories).

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

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Health and Environmental Research Online (HERO) website 

(http://hero.epa.gov/index.cfm/project/page/project_id/2216).
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Figure LS-1. Summary of literature search and screening process for RDX. 
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Selection of Critical Studies for Inclusion in Evidence Tables

Selection of Critical Studies 

1 The 21 studies retained after the literature search and screen (Figure LS-1) were evaluated 

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for aspects of its design or conduct that could affect the interpretation of results and the overall 

contribution to the synthesis of evidence for determination of hazard potential. Much of the key 

information for conducting this evaluation can generally be found in the study’s methods section 

and in how the study results are reported. Importantly, the evaluation at this stage does not 

consider study results, or more specifically, the direction or magnitude of any reported effects.

To facilitate this evaluation, evidence tables were constructed that systematically 

summarize the important information from each study in a standardized tabular format as 

recommended by the NRC (2011). The studies selected for inclusion in evidence tables are critical 

for assessing the health effects of RDX. The evidence tables include all studies that inform the 

overall synthesis of evidence for hazard potential; in general, the goal in developing evidence tables 

is to be inclusive. 

Studies were excluded from evidence tables if flaws in its design, conduct, or reporting are 

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

essential historical 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 removed from further 

consideration in the assessment because of fundamental issues with study design, conduct, or 

reporting. The specific studies and basis for exclusion are summarized in Table LS-2. 

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

issues with design, conduct, or reporting

Reference Rationale for exclusion

Haskell Laboratories (1942);

14-wk study in dogs

Incomplete information on exposure levels; low numbers of 

animals; breed of dog not reported; inadequate reporting of 

results; sections of document 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. 
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Reference Rationale for exclusion

Unpublished report 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, low numbers of animals 

tested, incomplete information on exposure levels, and inadequate 

reporting of results.

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studies of RDX involving repeated exposure that were not identified as uninformative because of 

fundamental issues with study design, conduct, or reporting were considered in assessing the 

evidence for health effects associated with chronic exposure to RDX. These studies are considered 

the “critical” studies for which study methods and results are presented in evidence tables and 

exposure-response arrays.

Other health effect studies of RDX, including human case reports and experimental animal 

studies involving exposures of 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. 

Study Evaluation

In evaluating the evidence to determine whether RDX exposure may pose a hazard for each 

of the health effects considered in this assessment, methodological aspects of a study’s design, 

conduct, and reporting were considered in the overall evaluation and synthesis of the pertinent 

data. In general the relevance and informativeness of the available studies were evaluated as 

outlined in the Preamble and in EPA guidance (e.g., A Review of the Reference Dose and Reference 

Concentration Processes (U.S. EPA, 2002) and Methods for Derivation of Inhalation Reference 

Concentrations and Application of Inhaled Dosimetry (U.S. EPA, 1994)). In addition, in 2012, EPA 

obtained external peer reviews of two 2-year bioassays (Lish et al., 1984; Levine et al., 1983) and 

one 90-day study (Crouse et al., 2006) that were available only as laboratory reports. The report of 

the peer reviews is available at www.epa.gov/hero (search for HERO ID 2519581). 

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 1.1. 

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; Ma and Li, 

1992; Hathaway and Buck, 1977). 

To varying degrees, these epidemiology studies of RDX are limited by their study design, 

uncertainty in estimates of exposure, inadequate reporting, and/or failure to account for potential 

confounding exposures. All three studies were based on a relatively small number of participants 
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1 (60‒69 exposed workers in the cross-sectional studies and 32 cases in the case-control study). The 

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study by Ma and Li (1992) of Chinese industrial workers provided limited information on 

participant recruitment, selection, and participation rate; information was not adequate to evaluate 

the potential for selection bias. 

Of the three epidemiological studies, more detailed exposure information was collected by

Hathaway and Buck (1977). Atmospheric and paired breathing zone sampling was performed; 

however, the paper included limited reporting of RDX concentrations in workplace air. Ma and Li 

(1992) reported mean RDX exposure concentrations (with standard deviations) for two exposure 

groups, but provided no information on the source of these concentrations or how monitoring was 

performed. In the case-control study by West and Stafford (1997), semi-quantitative exposure 

estimates (low, moderate, or high) were based on interviews with employees.

Ma and Li (1992) did not adjust for any potential risk factors, e.g., alcohol consumption. In 

the study by Hathaway and Buck (1977) that included evaluations of liver, renal, and hematology 

endpoints, workers with trinitrotoluene (TNT) exposure were appropriately excluded from the 

exposed groups, since TNT is another explosive that is associated with liver and hematological 

system toxicity. The case-control study by West and Stafford (1997), which examined hematology 

outcomes, did not perform statistical analyses to adjust for other risk factors or occupational 

exposures (including TNT). Further, the impact of age or gender could not be assessed as the cases 

and controls were not matched. 

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 1.1). 

In addition to these three occupational epidemiology studies, the human health effects 

literature includes 16 case reports that describe effects following acute exposure to RDX. Case 

reports are often anecdotal and typically describe unusual or extreme exposure situations, 

providing little information that would be useful for characterizing chronic health effects. 

Therefore, RDX case reports were only briefly reviewed; a critical evaluation of these studies was 

not undertaken. A summary of these case reports is provided in Appendix C, Section C.3. 

Experimental Animal Studies

The oral toxicity database for RDX 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). Only one inhalation study of RDX was identified. As discussed in Appendix B 

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

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

(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. 
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Test animal

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

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(F344, Sprague-Dawley, CD), mice (B6C3F1), dog (beagle), and monkey. 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 is 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 effect 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. For example, the informativeness of Hart 

(1974) and Martin and Hart (1974) was reduced in light of the small sample sizes in each study 

(3/sex/group), but the studies would still inform the consistency of effects observed for a specific 

endpoint across species (dog and monkey). 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 

RDX hazard are the particle size and purity of the test material. The purity of RDX used in health 

effects studies varied from 84-99.99%. The major contaminants were octahydro-1,3,5,7-tetranitro1,3,5,7-tetrazocine (HMX) and water, which are the primary contaminants of RDX produced 

through the Bachmann process. 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 reviewed by the IRIS Program in 1988 (www.epa.gov/iris); histopathological changes in the 
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1 liver in male F344 rats and in the kidney in female rats were reported in a 13-week feeding study. 

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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 of ≥20 mg/kg-day. It should be noted that the test 

materials employed in these studies are considered representative 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 relatively fine particle size 

(majority of particles were <66 μm in size) while others used 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 

some of the toxicities 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. 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 1.1. 

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 number animals with seizures 

anecdotally. 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. 

Outcomes and data reporting

In evaluating studies, consideration was given to whether data were reported for all prespecified endpoints and 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 individual fetus). Data from studies have been extracted and presented in evidence tables.
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Notable features of the RDX database

1 Three two-year toxicity bioassays of RDX are available as unpublished laboratory studies. 

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The bioassays by Levine et al. (1983) 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. 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-3). Because they were available only as 

laboratory reports, peer reviews of the Levine et al. (1983) and Lish et al. (1984) studies were 

conducted by EPA in 2012. The peer reviewers generally concluded that the reports provided 

useful information on the toxicity of RDX, noting that there were limitations in interpretation due to 

the histopathological analysis and the statistical approaches employed in the reports. An earlier 

two-year study in the rat by Hart (1976) used a dose range that was lower than the subsequent 

studies (high dose of 10 mg/kg-day), and that may not have been sufficient to elicit some effects in 

treated animals. Histopathology findings were limited by the lack of pathology examinations in the 

mid-dose groups and the 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 on study (see Table LS-3).

Short-term and subchronic toxicity studies of RDX were published or reported between the 

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

that range. All but one of the eight short-term and subchronic toxicity studies of RDX are available 

as unpublished laboratory studies; only Von Oettingen et al. (1949) was published. 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 EPA in 

2012. The peer reviewers determined that the report provided useful information on the toxicity of 

RDX, including an array of endpoints for neurotoxicity and immunotoxicity. Limitations in the 

study were based on an incomplete understanding of the neurotoxicity that may have been 

resolved with more histological evaluation as well 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-3. Limitations of these 

studies were taken into consideration in evaluating and synthesizing the evidence for each of the 

health effects in Section 1.1.
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Table LS-3. 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); Levine 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. There were still more 

than 80 rats/sex/group after the overheating incident, and ≥50 

rats/sex/group at termination.

Histopathology findings were limited by the lack of pathology 

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

death, which impacts the ability to interpret the histopathology data.

Cholakis et al. (1980)

13-wk mouse study (Experiment 1)

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, 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.

Von Oettingen et al. (1949)

12-wk rat study

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

autopsy.

Von Oettingen et al. (1949)

6-wk dog study

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

autopsy.

Martin and Hart (1974)

90-d monkey study

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).
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1.HAZARD IDENTIFICATION

1.1. PRESENTATION AND SYNTHESIS OF EVIDENCE BY ORGAN/SYSTEM

1.1.1. Nervous System Effects

1 Nervous system effects following RDX exposure have been observed in multiple case

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reports, and the association between RDX exposure and neurobehavioral effects has been examined 

in a single 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 1-1 and 

1-2 and Figure 1-1.

In a cross-sectional study by Ma and Li (1992), neurobehavioral effects were evaluated in 

Chinese workers occupationally exposed to RDX. Memory retention and block design scores2 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, block design, and letter cancellation test) (Table 1-1). This study 

did not consider potential confounders such as alcohol consumption or co-exposure to TNT. 

Case reports support an association between RDX exposure and neurological effects (see 

Appendix C, Section C.3). Severe neurological disturbances include tonic-clonic seizures (formerly

known as grand mal seizures) in factory workers (Testud et al., 1996b; Testud et al., 1996a; 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 non-wartime/nonoccupational exposures (Kasuske et al., 2009; Davies et al., 2007; Küçükardali et al., 2003; Hett and 

Fichtner, 2002; Harrell-Bruder 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). 

Nervous system effects in experimental animals, including seizures and convulsions (used 

interchangeably by study authors), tremors, behavioral changes, irritability, and hyperactivity, have 

been observed in the majority of chronic, subchronic, and developmental studies following oral 

exposure to RDX (see Table 1-2 and Figure 1-1). In a 2-year dietary study in F344 rats, 

 

2The 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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1 administration of 40 mg/kg-day RDX resulted in convulsions (Levine et al., 1983); convulsions were

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not observed at lower doses in the same study (≤8 mg/kg-day) (Levine et al., 1983) or in SpragueDawley rats following chronic dietary administration of 10 mg/kg-day, the highest dose tested

(Hart, 1976). Convulsions were observed in B6C3F1 mice exposed to RDX for 2 years at doses 

similar to or higher than those inducing convulsions in the rat (Levine et al., 1984; Lish et al., 1984). 

Subchronic dietary exposure 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 contrast, Levine et al. (1990) reported convulsions in rats 

following subchronic exposure only at a dose of 600 mg/kg-day; however, the unpublished 

technical report of this study (Levine et al., 1981a) inconsistently reported convulsions at 600 

mg/kg-day and ≥30 mg/kg-day, thereby reducing confidence in the identification of the dose level 

at which nervous system effects are observed in this study. No evidence of seizures, convulsions or 

tremors was reported in three subchronic rat studies that used relatively lower doses of RDX 

(highest administered doses: 10−50 mg/kg-day) (MacPhail et al., 1985; Cholakis et al., 1980). 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)—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 (Von Oettingen et al., 1949), but not 10 mg/kg-day (Hart, 1974), and in two of 

three monkeys of both sexes following a gavage dose of 10 mg/kg-day (Martin and Hart, 1974). 

In the only study addressing susceptibility to seizures, Burdette et al. (1988) found that 

seizure occurrence was greater 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 post-dosing test periods even 

though RDX plasma concentrations were elevated throughout the testing 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 these experiments were 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) 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 1-2). 

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 (Crouse et al., 
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1 2006; Cholakis et al., 1980) induced convulsions at lower doses than did dietary administration, 

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possibly due to the bolus dosing resulting from gavage administration and the comparatively faster 

peak absorption of RDX. 

Several experimental animal studies documented that unscheduled deaths were frequently 

preceded by convulsions or seizures. Crouse et al. (2006) stated that nearly all observed pre-term 

deaths in rats exposed to RDX for 90 days were preceded by neurotoxic signs such as tremors and 

convulsions. In a 2-year study in rats, Levine et al. (1983) observed that tremors and/or 

convulsions were often seen in high-dose animals prior to their death. Further, in a rat 

developmental 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 round the mouth and nose) in other dams that died during the

study. A few studies reported mortality that was not specifically or directly associated with 

neurological effects (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.

Additional neurobehavioral effects associated with RDX exposure in rats included increased 

hyperactivity, hyper-reactivity, fighting, and irritability at doses similar to those that induced 

tremors, convulsions, and seizures (10−100 mg/kg-day) (Levine et al., 1990; Angerhofer et al., 

1986; Levine et al., 1983; Levine et al., 1981a; Levine et al., 1981b; Cholakis et al., 1980; Von 

Oettingen et al., 1949). Hyperactivity and nervousness were also reported in male mice that 

received a subchronic exposure to 320 mg/kg-day 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, 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)

concluded that stained haircoats and increased barbering in female F344 rats receiving 15 mg/kgday may have been caused by the oral dosing procedure (gavage) alone.

Observations of changes in absolute and relative brain weight were mixed across studies. 

Among chronic 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 (Levine et al., 1984; Lish et al., 1984). 

Conversely, an increase in absolute brain weight of 2% relative to control was observed in F344 

rats at 40 mg/kg-day in another two-year oral bioassay (Levine et al., 1983; Thompson, 1983). 

Similarly 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; Levine et al., 1981b; 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., 1983; Thompson, 1983; 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 
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1 of effects of RDX on brain weights. Based on an evaluation of the relationship between organ 

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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 brain weights are not modeled well by any of the choices, and that 

alternative analysis methods should be utilized. 

Across the studies summarized in Table 1-2, nervous system responses to RDX did not show 

a predicted relationship with duration of exposure. For example, seizures or convulsions were 

observed in F344 rats in some subchronic studies at doses lower than in studies of chronic

duration, and at even lower doses in dams exposed for approximately 2 weeks during gestation. In 

some studies, seizures appeared soon after dosing, suggesting that seizure induction was more 

strongly correlated with dose level rather than with duration of exposure. Williams et al. (2011)

demonstrated that RDX is rapidly absorbed and crosses the blood-brain barrier following oral 

administration in rats, and that distribution of low levels of RDX (8 μg/g ww) to the brain 

correlated with seizure onset. 

Similarly, nervous system effects across studies did not show a consistent relationship with 

dose. This lack of consistency may, at least in part, be attributed to differences in the purity or 

particle size of the test material across studies. Assuming that increased particle size 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. (1983)

that used a smaller (<66 μm) particle size. This could contribute to why the Cholakis et al. (1980)

subchronic dietary study in the mouse (doses up to 320 mg/kg-day RDX) and rat (doses up 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 number of daily observations for clinical signs may not have 

been sufficiently frequent to provide an accurate measure of the incidence of seizures or other 

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

Reference and study design Results

Ma and Li (1992) (China)

Cross-sectional study, 60 workers 

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; exposed workers were 

divided into two groups based on RDX 

concentration in the air:

Concentration (mg/m3

) 

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).

1 

2 aSymptom data were not included in evidence table because of incomplete reporting.
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Table 1-2. Evidence pertaining to nervous system effects in animals

Reference and study design Results

Convulsions and neurobehavioral effects

Lish et al. (1984); Levine 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

24 mo

One male mouse in the 35 mg/kg-d dose group and one female 

mouse 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 neurological effects, as evidenced by clinical signs or changes in 

appearance or behavior, were reported.

Levine et al. (1983); Thompson (1983)

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

24 mo

Tremors, convulsions, and hyper-responsiveness to stimuli were 

noted 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)b

Diet

13 wks

Hyperactivity and/or nervousness observed in 50% of the high-dose 

males; no signs observed in femalesa; no incidence data were 

reported.
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Reference and study design Results

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 neurological effects, as evidenced by clinical signs or changes in 

appearance or behavior, were reported. 

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

13 wks

No neurological 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

90 d

Doses 0 4 8a 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. (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

Hyper-reactivity to approach was observed in groups receiving 

≥100 mg/kg-d; no incidence data were reported.

Tremors and convulsions were observed prior to death in some 

animals receiving 600 mg/kg-d; no incidence data were reported.c

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

3 mo

Hyperirritability and convulsions were observed in the 25 and 

50 mg/kg-d groupsa; no incidence data were reported.
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1-8 DRAFT—DO NOT CITE OR QUOTE

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

0, 0.1, 1, or 10 mg/kg-d

Diet

90 d

No neurological effects, as evidenced by clinical signs or changes in 

appearance or behavior, were reported.

Martin and Hart (1974)

Monkeys, Cynomolgus or Rhesus, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d

Gavage

90 d

Doses 0 0.1 1 10a

CNS effects characterized as trembling, shaking, jerking, or 

convulsions (incidence)

M 0/3 0/3 0/3 2/3

F 0/3 0/3 0/3 2/3

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 reflexesa; 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

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.
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1-9 DRAFT—DO NOT CITE OR QUOTE

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 hyperactivitya were observed at 20 mg/kg-day; 

no incidence data were reported.

Brain weight

Lish et al. (1984); Levine 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

24 mo

Doses 0 1.5 7 35 175/100

Absolute brain weight

M 0% −0.2% 0.61% 0.81% −1%

F 0% −2% −2% −4%* −3%*

Relative brain weight

M 0% 4% 2% 2% 5%

F 0% −4% −1% −3% 18%*

Levine et al. (1983); Thompson (1983)

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

24 mo

Doses 0 0.3 1.5 8 40

Absolute brain weight

M 0% 2% −1% 2% 2%

F 0% −0.3% −0.4% 1% 2%*

Relative brain weight

M 0% 0% 8% 2% 22%*

F 0% −1% 3% 4% 20%*

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

M 0% − − − 2% 2%

F 0% − − − 4% 2%

Relative brain weight

M 0% − − − 6% 2%

F 0% − − − 0% 3%
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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)b

Diet

13 wks

Doses 0 80 160 320

Absolute brain weight

M 0% 0% 2% 10%

F 0% 0% 4% 2%

Relative brain weight

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

M 0% − − − 3% 0%

F 0% − − − 0% 0%

Relative brain weight

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

90 d

Doses 0 4 8 10 12 15

Absolute brain weight

M 0% −1% −0.3% 2% 5%* 7%*

F 0% −2% 6% 1% 4% 6%

Relative brain weight

M 0% 6% 10% 5% 3% 4%

F 0% −2% −2% −12%* −12%* −15%*

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

Absolute brain weight

M 0% 1% 0.53% −6% − −

F 0% −1% 1% 2% − −

Relative brain weight

M 0% 4% 7% 14% − −

F 0% 0.3% 2% 5% − −

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a

Mortality was reported in some RDX-treated groups in this study. b

Doses were calculated by the study authors.

c

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 
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1-11 DRAFT—DO NOT CITE OR QUOTE

1 rats receiving 100 mg/kg-d and that hyperactivity to approach, tremors, and convulsions were observed in 

2 

3 

4 

animals exposed to lethal doses (discussion). 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.
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1-12 DRAFT—DO NOT CITE OR QUOTE

 

M M

M

M

M

M

M

M

M

M

M

M

M M

M M

M

M

0.1

1

10

100

1000

Lish et al., 1984 (mouse, male)

Lish et al., 1984 (mouse, female)

Hart, 1976 (rat)

Levine et al., 1983 (rat)

Cholakis et al., 1980 (rat)Cholakis et al., 1980 (rat)

Crouse et al., 2006 (rat, male)

von Oettingen et al., 1949 (rat)

Hart, 1974a (dog)

Martin and Hart, 1974b (monkey)

von Oettingen et al., 1949 (dog)

MacPhail et al., 1985 (rat)

Cholakis et al., 1980 (rat, male)

Cholakis et al., 1980 (rat, female)

Angerhofer et al., 1986 (rat)Angerhofer et al., 1986 (rat)

Lish et al., 1984 (mouse, male)

Lish et al., 1984 (mouse, female)

Levine et al., 1983 (rat, male)

Levine et al., 1983 (rat,female)

Cholakis et al., 1980 (mouse)Cholakis et al., 1980 (mouse)

Cholakis et al., 1980 (rat, male)

Crouse et al., 2006 (rat, male)

Levine et al., 1990, 1981a,b (rat)

Chronic Subchronic Gestational (dams)

Convulsions and/or Seizures Absolute Brain Weight

Dose (mg/kg-day)

significantly changed

not significantly changed

M- Mortality observed at this dose and above

Figure 1-1. Exposure response array of nervous system effects following oral exposure.3

3Due to the severity of the endpoint for convulsions and/or seizures, a response in treated groups was determined to be significant (filled circles) in the 

exposure-response array where there was an observation of convulsions and/or seizures reported in the study. 
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1-1 DRAFT—DO NOT CITE OR QUOTE

Mechanistic Evidence

1 The few studies that have explored the MOA of RDX on the central nervous system have 

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focused on potential impacts on neurotransmission. These studies suggest that the MOA for RDXinduced seizures and convulsions involves distribution to the brain (across the blood-brain barrier) 

and subsequent effects on neurotransmitters, including gamma amino butyric acid (GABA) and 

glutamate. The strongest mechanistic information for RDX neurotoxicity comes from documented 

interactions 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). It 

is also a target of many anticonvulsant therapies (e.g., benzodiazepines, propofol, barbiturates)

(Meldrum and Rogawski, 2007; Möhler, 2006). The affinity of RDX for the GABAA receptor provides

biological plausibility for the association of seizures with exposure to RDX in both human case 

reports and experimental animal studies. 

In research conducted by the U.S. Army Center for Health Promotion and Preventative 

Medicine, Williams et al. (2011) and Bannon et al. (2009) 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., 2009). 

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) common AChE-induced symptoms such as salivation and lacrimation have not always been 

observed (Williams et al., 2011); 2) blood and brain levels of AChE are unaffected by RDX (Williams 

et al., 2011; Williams and Bannon, 2009); and 3) in vitro neurotransmitter receptor binding studies 

do not reveal any affinity of RDX for acetylcholine receptors (Williams et al., 2011; Williams and 

Bannon, 2009). 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). 

As noted above, in receptor binding assays RDX 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. The authors demonstrated that RDX 

treatment in brain slices from the basolateral amygdala inhibit GABAA−mediated inhibitory 

postsynaptic currents and initiated seizure-like neuronal discharges. RDX exposure may reduce the 
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1-2 DRAFT—DO NOT CITE OR QUOTE

1 inhibitory effects of GABAergic neurons, resulting in enhanced excitability that could lead to 

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seizures (Williams et al., 2011; Williams and Bannon, 2009), although additional studies are 

necessary to substantiate this observation and to clarify the potential cellular and regional targets 

of RDX-induced neurotoxicity. 

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 (e.g., kindling) (Jefferys et al., 2012; Gilbert, 1994). Burdette et al. (1988) hypothesized that 

the limbic system was involved in seizures caused by RDX exposure, given than rats exhibited proconvulsant activity in response to amygdaloid kindling at a dose that was approximately half the 

dose necessary for RDX to induce spontaneous 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, and irritability 

noted across several studies of RDX toxicity (Levine et al., 1990; Levine et al., 1983; Thompson,

1983; Levine et al., 1981a; Levine et al., 1981b; Cholakis et al., 1980; Von Oettingen et al., 1949). 

In a microarray experiment, Bannon et al. (2009) 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 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/kg-day for 28 days. One miRNA, miR206, was upregulated 26-fold in 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. Effects of RDX on BDNF expression may play a role in RDX 

neurotoxicity, but the utility of miRNAs as predictors of toxicity has not been established, and the 

contribution, if any, of aberrant expression of a suite of miRNAs to the MOA for RDX neurotoxicity is 

unknown. 

Information from a small number of studies suggests that inhibition of GABAergic signaling 

in the limbic system could represent a likely mechanism for RDX-induced hyperactivity and 

seizures. However, the available data are insufficient to identify any specific mode(s) of action for 

the nervous system effects observed following RDX exposure.

Summary 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 
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1-3 DRAFT—DO NOT CITE OR QUOTE

1 in certain neurobehavioral tests in workers exposed to RDX compared to controls (Ma and Li, 

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1992), and human case reports provide other evidence of an association between acute RDX 

exposure and neurological effects. Eleven of 16 repeat-dose animal studies reported neurological 

effects, 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., 

1983; 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 particle size and purity of the RDX administered could possibly account for the 

lack of effect. Although the specific mode(s) of action for RDX-induced nervous system effects 

remains unknown, evidence that RDX exposures may lead to seizures through binding to the GABAA

receptor provides biological support for this association. EPA identified nervous system effects as a 

human hazard of RDX exposure. 

1.1.2. Kidney and Other Urogenital System Effects

The association between RDX exposure and effects on clinical measures of kidney function

was examined in a single occupational epidemiology study. Case reports involving accidental 

exposure to ingested or inhaled RDX provide some information on the potential for acute exposures 

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 kidney and other urogenital effects associated with 

RDX exposure is presented in Tables 1-3 to 1-7 and Figure 1-2.

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 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 to RDX and HMX; average exposure of up to 

1.5 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 1-3).
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-4 DRAFT—DO NOT CITE OR QUOTE

1 Studies in experimental animals provide some evidence that RDX exposure is associated 

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with kidney and other urogenital effects (Table 1-4 and Figure 1-2). Dose-related increases in 

absolute and relative kidney weights (19−27% compared to control) were observed in male B6C3F1

mice exposed to RDX in the diet for 2 years (Lish et al., 1984) and a dose-related increase in relative 

kidney weights (up to 19%) was observed in female mice. Relative, but not absolute, kidney 

weights were increased (20−21% compared to control) in male and female F344 rats exposed to 

40 mg/kg-day RDX in the diet for 2 years (Levine et al., 1983). Changes in kidney weights in other 

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 (Crouse et al., 

2008; Levine et al., 1990; Levine et al., 1981a; Levine et al., 1981b; Cholakis et al., 1980; Hart, 1974; 

Martin and Hart, 1974). Based 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 kidney weights are not modeled well by any of the choices, and that 

alternative analysis methods should be utilized. 

Histopathological changes in the urogenital system associated with exposure to RDX were 

observed in a 2-year bioassay in which 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., 1983). Similar kidney lesions were 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; Levine et al., 1981b; 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 165 mg/kg RDX in DMSO for 4 weeks

(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; Levine et al., 1981b; 

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. 

Exposure to the major contaminant in many of the available RDX studies, HMX, was 

associated with histopathological changes in the kidney and alterations in renal function in female 
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-5 DRAFT—DO NOT CITE OR QUOTE

1 rats fed doses ≥450 mg/kg-day HMX for 13 weeks. No effects were observed at doses ≤115 mg/kg2 

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day. Because the percentage of HMX as an impurity ranged from 3−10% resulting in HMX 

exposures of ≤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. 

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 two years (Levine et al., 

1983). The Levine et al. (1983) report is the only 2-year study that reported examination of the 

prostate in rats. 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; Levine et al., 1981b). Similarly, prostate effects were not observed in a 2-year dietary study 

in mice (Lish et al., 1984). Some reports have hypothesized that the observation of prostate 

inflammation in Levine et al. (1983) is secondary to a bacterial infection unrelated to RDX toxicity 

(ATSDR, 2012; Sweeney et al., 2012a; Crouse et al., 2006). For example, Crouse et al. (2006)

concluded 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. (1983) distinguish 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 are combined. Additionally, the doserelated nature of the increased incidence suggests that the primary cause (potentially leading to 

bacterial infection) was treatment-related since a more uniform distribution of rats with 

suppurative prostatitis would be expected with a spontaneous or age-related lesion. The doseresponsiveness could be explained if the infections were secondary to treatment-related 

immunotoxicity, but there is no evidence from Levine et al. (1983) 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). 

Levine et al. (1983) document 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 1-5 to 1-7). The absence of these observations in the 

6-month interim pathology report suggests that an exposure duration of greater than 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. 
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-6 DRAFT—DO NOT CITE OR QUOTE

1 Considered as a group, treatment-related kidney and urogenital lesions may have led to a blockage 

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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, 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 the suppurative prostatitis (ATSDR, 2012; Sweeney 

et al., 2012a; Crouse et al., 2006). 

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 the observations of 

suppurative prostatitis would arise 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. 

(1983) 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, since 

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 dose groups (Levine et al., 1983; Thompson, 1983). 

Although the pathogenesis of kidney and urogenital 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 effects observed by Levine et al. (1983). 

Table 1-3. 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 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 in men: mean (standard deviation not 

reported)

Test

Referent

(n = 237)

RDX exposed

Undetected

(n = 22)

>0.01 mg/m3

(n = 45)

BUN 15.5 15.6 16.4

Total protein 7.2 7.2 7.3

No differences were statistically significant. Similar results in 

women.
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1-7 DRAFT—DO NOT CITE OR QUOTE

Table 1-4. Evidence pertaining to kidney and other urogenital system effects 

in animals 

Reference and study design Results

Kidney weight

Lish et al. (1984); Levine 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

24 mo

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. (1983); Thompson (1983)

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

24 mo

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%
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1-8 DRAFT—DO NOT CITE OR QUOTE

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, 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, 50 

mg/kg-d

Diet

13 wks

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

90 d

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%*

Levine et al. (1981a);Levine et al. (1990); 

Levine et al. (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

Data were not reported for rats in the 300 or 600 mg/kg-d groups 

because all of the rats died before the 13-wk necropsy.

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

1-9 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

Diet

13 wks

Relative kidney weight (percent change compared to control)

M 0% 5% 7% 10% − −

F 0% 3% 5% 2% − −

Hart (1974)

Dogs, Beagle, 3/sex/group

Pre-mix with ground dog chow containing 

20 mg RDX/g-chow, 60 grams of dog food

0, 0.1, 1, or 10 mg/kg-d

Diet

90 d

Numerical values given only for control and 10 mg/kg-d groups.

Doses 0 0.1 1 10

Absolute kidney weight (percent change compared to control)

M 0% − − 38%

F 0% − − −18%

Martin and Hart (1974)

Monkeys, Cynomolgus or Rhesus, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d

Gavage

90 d

Doses 0 0.1 1 10

Absolute kidney weight (percent change compared to control)

M + F 0% −2% −3% 4%

Histopathological lesions

Lish et al. (1984); Levine 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

24 mo

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 control animals as animals 

treated with RDX. 

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. (1983); Thompson (1983)

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

24 mo

Note: More detailed histopathological 

results, including interim sacrifice data at 

6 and 12 mo, are provided in Tables 1-5 to 

1-7.

Data were analyzed separately for animals sacrificed on schedule (SS) 

and those that died spontaneously or were sacrificed moribund 

(SDMS); 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)
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1-10 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

(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

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

Incidence data reported only for controls and the 320 mg/kg-d group.

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. 
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1-11 DRAFT—DO NOT CITE OR QUOTE

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

13 wks

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

90 d

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. (1981a);Levine et al. (1990); 

Levine et al. (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.

Hart (1974)

Dogs, Beagle, 3/sex/group

Pre-mix with ground dog chow containing 

20 mg RDX/g-chow, 60 grams of dog food

0, 0.1, 1, or 10 mg/kg-d 

Diet

90 d

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 Rhesus, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d 

Gavage

90 d

Doses 0 0.1 1 10

Medulla; mineralization, minimal to mild (incidence)

M + F 0/6 1/6 0/6 4/6

1 

2 

3 

4 

5 

*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.
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1-12 DRAFT—DO NOT CITE OR QUOTE

Table 1-5. Six-, 12-, and 24-month incidence of kidney endpoints in male F344 

rats reported for statistical evaluation in Levine et al. (1983) 

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 *Statistically significant (p < 0.05) based on analysis by study authors. 

2 

3 

4 

Source: Levine et al. (1983). 

Table 1-6. Six-, 12-, and 24-month incidence of urinary bladder endpoints in 

male F344 rats reported for statistical evaluation in Levine et al. (1983)

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

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 - 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 

*Statistically significant (p < 0.05) based on analysis by study authors. 

Source: Levine et al. (1983).

Table 1-7. Six-, 12-, and 24-month incidence of prostate endpoints in male 

F344 rats reported for statistical evaluation in Levine et al. (1983)

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

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 - 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 

*Statistically significant (p < 0.05) based on analysis by study authors. 

Source: Levine et al. (1983). 
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M

M

M

M M

M

M M M M M

M

M M M

0.1

1

10

100

1000

 Lish et al., 1984 (mouse, male)

 Lish et al., 1984 (mouse, female)

Hart, 1976 (rat)

Levine et al., 1983 (rat)

Cholakis et al., 1980 (mouse)Cholakis et al., 1980 (mouse)

Cholakis et al., 1980 (rat)

 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., 1983 (rat)

 papillary necrosis, pyelitis; Levine et al., 1983 (rat) 1

luminal distention; Levine et al., 1983 (rat)

cystitis; Levine et al., 1983 (rat)2

suppurative prostitis: Levine et al., 1983 (rat)

tubular nephrosis

Cholakis et al., 1980 (mouse)

Cholakis et al., 1980 (rat)

Cortical cysts

Cholakis et al., 1980 (rat)

subacute prostate inflammation

Crouse et al., 2006 (rat)

Levine et al., 1990, 1981a,b (rat)

Hart, 1974a (dog)

medulla mineralization

Martin and Hart, 1974b (monkey)

Chronic Subchronic Chronic Subchronic

↑ Relative kidney 

weight

↓Relative kidney weight Histopathological lesions

Dose (mg/kg-day)

significantly changed

not significantly changed

The following studies were excluded from array because absolute kidney weight was reported: Cholakis, 1980 (2-gen rat); Hart, 1974; Martin and Hart, 1974

M - Mortality observed at this dose and above 1 statistical significance determined from incidence at time of of scheduled sacrfice

2 statistical significance determined from incidence at spontaneous death. 

Figure 1-2. Exposure-response array of kidney and urogenital system effects.
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Mechanistic Evidence

1 No MOA information is available for RDX-induced kidney and other urogenital effects, 

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including suppurative prostatitis. 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 modes 

of action for the effects reported in Levine et al. (1983). 

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, 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. (1983) could have been

produced by disruption of pituitary prolactin or other hormonal signal via interference with normal 

regulatory GABA-related hormonal control. However, no direct evidence for this hypothesized 

MOA is available. Levine et al. (1983) did not evaluate serum endocrine measures or pituitary

weights, and pituitary adenomas that could account for higher prolactin levels were not observed. 

A MOA based on pituitary-mediated alterations in endocrine signaling also does not explain the 

other urogenital lesions observed by Levine et al. (1983). 

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 murine 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, since they bind 

to the same site) was able to reduce this effect. However, picrotoxin on its own did not significantly 

alter cytokine production, suggesting the effects are limited to reversal of agonist-induced 

GABAergic activity. 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 it is assumed that the kidney and other urogenital effects are mediated through localized 

interaction with GABAA receptors, another possibility is 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 
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1 maintained in a non-inflamed state, mediated by GABA; RDX binding to GABAA receptor-convulsant 

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sites on the prostate would result in a reduction of the inhibitory effects of the GABA receptor 

leading to increased inflammation. No evidence was found to support this potential pathway 

leading to prostate inflammation. 

Another hypothesis is that the kidney and other urogenital effects of RDX are caused by

interactions with GABAA receptors mediating inputs to the urogenital system. 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. (1983); however, no evidence to support this hypothesized MOA is 

available.

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, accounting for the observed 

kidney and urogenital effects. Among the mechanistic information presented above, modes of 

action that require direct action on the prostate are considered less likely, because the available 

information suggests the prostatitis is a secondary effect. However, the ways 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. 

Summary of Kidney and Other Urogenital System Effects

Evidence for kidney effects resulting from RDX exposure consists of human case reports and 

some findings of increased kidney weight and 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 accidental exposure (ingestion and inhalation) to unknown 

amounts of RDX. No RDX-related changes in kidney parameters were found in a small crosssectional study of RDX-exposed workers (Hathaway and Buck, 1977). Treatment-related increases 

in relative kidney weight were consistently observed in rats and mice of both sexes in two chronic 

oral toxicity studies (Lish et al., 1984; Levine et al., 1983); however, kidney weights across studies 

of subchronic duration generally failed to show a consistent pattern of change. Measurement of 
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1 serum chemistry parameters in multiple animal species did not provide consistent evidence of 

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dose-related changes associated with RDX exposure. 

Histopathological changes in a two-year study in F344 rats, including a dose-related 

increase in the incidence of suppurative prostatitis in male rats (Levine et al., 1983; Thompson, 

1983), provides the strongest evidence of RDX-associated kidney and other urogenital 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. A second 2-year study in 

Sprague-Dawley 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. (1983). In light of 

the dose-related increase in suppurative prostatitis and lack of support for an alternative (i.e., nonRDX-related) basis for this effect, EPA identified kidney and other urogenital effects as a potential 

human hazard of RDX exposure. 

1.1.3. 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 study in rats and 

studies in rats and rabbits involving gestational exposure to ingested RDX. In addition, oral 

subchronic and chronic studies in experimental animals provide information useful for examining 

the association between RDX exposure and effects on the male reproductive system. A summary of 

the developmental and reproductive effects associated with RDX exposure is presented in Tables 

1-8 and 1-9 and Figures 1-3 and 1-4. 

Developmental Effects

Animal studies report effects of RDX on offspring survival. Pup survival rates in the F0 and 

F1 generations 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), but only at the highest dose tested (50 mg/kg-day) that also produced toxicity in adults

(neurotoxicity, mortality, and reduced body weights and food consumption). Decreased fetal 

viability was observed at 20 mg/kg-day in F344 rats (Cholakis et al., 1980), although no effect on 

live fetuses was observed in Sprague-Dawley rats at the same dose (Angerhofer et al., 1986); both 

of these studies reported significant mortality in dams at 20 mg/kg-day. Increased resorptions 

were similarly limited to the highest dose tested (20 mg/kg-day), i.e., a dose associated with 

maternal toxicity (Cholakis et al., 1980). There was no evidence of maternal toxicity, 

embryotoxicity or decreased fetal viability in a teratology study of pregnant rabbits exposed to 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 exposed to RDX by gavage from GD 6 to 15 (Angerhofer et al., 
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1986).4 1 Maximum decreases in fetal body weight (9%) and body length (5%) were observed at 

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20 mg/kg-day, a dose that produced significant 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 maternal mortality (Cholakis et al., 1980). The larger SpragueDawley litter sizes and number of fetuses, compared to F344 rats, may account for the greater 

statistical power to observe treatment-related effects. Dose-related reductions in fetal body weight 

were not observed in rabbits at doses up to 20 mg/kg-day (Cholakis et al., 1980). 

No treatment-related teratogenic effects have been reported in rats exposed to a dose as 

high as 20 mg/kg-day RDX, a dose that resulted in approximately 30% 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 7-29 also provided little evidence of teratogenicity (Cholakis et al., 

1980). Increased incidences of enlarged front fontanel and unossified sternebrae were observed in 

all groups of rabbits exposed to RDX (Cholakis et al., 1980); however, these developmental 

anomalies did not exhibit a dose-related increase. Gestational exposure to RDX did not result in any 

other skeletal abnormalities. 

Reproductive Effects

Evidence of male reproductive toxicity is provided by the finding of testicular degeneration 

in male mice (Table 1-9 and Figure 1-4). An increased incidence of testicular degeneration was 

observed in male B6C3F1 mice exposed to ≥35 mg/kg-day RDX for 2 years in the diet (10−11%)

compared to concurrent (0%) and historical (1.5%) controls (Lish et al., 1984). Reductions in 

absolute testicular weight were observed, but the magnitude of the effect was small (≤6% 

compared to controls) and not dose-related. 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 all male rats (including controls) had testicular masses and no instances of germ cell 

degeneration were identified in control or RDX-treated groups (Levine et al., 1983). No doserelated histopathological changes in the testes were identified in other studies in rats (Crouse et al., 

2006; Levine et al., 1990; Levine et al., 1981a; Levine et al., 1981b; 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; Levine et al., 1981b; 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. Based 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 

 

4 The 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.
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1 toxicity, Bailey et al. (2004) concluded that testes weights are not modeled well by any of the 

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choices, and that alternative analysis methods should be utilized. 

Reproductive function was assessed in two separate studies reported by Cholakis et al. 

(1980). In the dominant lethal mutation study, no effects on fertility were observed in male rats 

exposed to ≤16 mg/kg-day RDX. Pregnancy rates were lower in females mated to males exposed to 

50 mg/kg-day RDX for 15 weeks prior to mating, although this effect was attributed to decreased 

well-being of the males in this high-dose group (Cholakis et al., 1980). No specific effects on 

reproductive function were observed in F0 and F1 rats exposed to ≤16 mg/kg-day RDX in a twogeneration 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/kgday dose, a dose associated with reduced body weight and feed consumption and increased 

mortality, suggests that effects on reproductive function were likely due to the general toxicity of 

RDX rather than a direct effect of RDX on reproduction.

Table 1-8. Evidence pertaining to reproductive and developmental effects in 

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

13 wks

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%*

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 six died 

during subsequent treatment.

Note: results on a per litter basis were not provided.
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Reference and study design Results

Cholakis et al. (1980)

Rabbits, New Zealand White, 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%

Complete litter resorptions (number of litters)

0 0 0 2

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.

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.
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Reference and study design Results

Offspring growth

Cholakis et al. (1980)

Rabbits, New Zealand White, 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%b

Significant maternal mortality (16/51) occurred at 20 mg/kg-d.

Morphological development

Cholakis et al. (1980)

Rabbits, New Zealand White, 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
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1-24 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

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.
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1-25 DRAFT—DO NOT CITE OR QUOTE

M

M M M M M M

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

M - Maternal mortality observed at the highest dose 1 Statistically signficant dose-related trend (p <= 0.05) by linear trend test, performed for this assessment. 

Figure 1-3. Exposure response array of reproductive and developmental effects following oral exposure. 
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1-26 DRAFT—DO NOT CITE OR QUOTE

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

Reference and Study Design Results

Lish et al. (1984); Levine 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

24 mo

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. (1983); Thompson (1983)

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

24 mo

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.

SDMS = spontaneous death or moribund sacrifice; 

SS = scheduled sacrifice

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 

Doses 0 80 160 320

Absolute testes weight (percent change compared to control)
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1-27 DRAFT—DO NOT CITE OR QUOTE

Reference and Study Design Results

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)d

Diet

13 wks

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 day

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

13 wks

In F2 offspring of 0, 5, and 16 mg/kg-d groups. No high-dose F2 

animals available.

Doses 0 5 16 50

Absolute testes weight (percent change compared to control)

0% 3% −31% −

Testes were examined microscopically in all F2 groups; no effects 

observed.

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d

Gavage

90 d

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)

Dogs, Beagle, 3/sex/dose

Pre-mix with ground dog chow containing 

20 mg RDX/g-chow, 60 grams of dog food

0, 0.1, 1, or 10 mg/kg-d

Doses 0 0.1 1 10

Absolute testes (with epididymis) weight (percent change compared 

to control)

0% − − 51%
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1-28 DRAFT—DO NOT CITE OR QUOTE

Reference and Study Design Results

Diet

90 d

Testes were not examined microscopically.

1 

2 

3 

4 

5 

6 

7 

8 

9 

10

11

12

*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. d

Testicular atrophy was observed at 12 months along with a statistically reduced mean testes weight (compared 

with controls). By 24 months, all male rats (including controls) had testicular masses; 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.
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1-29 DRAFT—DO NOT CITE OR QUOTE

M

M

M

M

M

M

M

0.1

1

10

100

1000

Lish et al., 1984 (mouse)

Hart, 1976 (rat) 1

Cholakis et al., 1980 (mouse)Cholakis et al., 1980 (mouse)

Cholakis et al., 1980 (rat)

Cholakis et al., 1980 (rat, F2)

Crouse et al., 2006 (rat)

Levine et al., 1990; 1981a; b (rat)

Hart, 1974a (dog) 1

Lish et al., 1984 (mouse)2

Hart, 1976 (rat)

Levine et al., 1983 (rat)

Cholakis et al., 1980 (mouse)

Cholakis et al., 1980 (rat)

Cholakis et al., 1980 (rat, F2)

Levine et al., 1990; 1981a; b (rat)

↓ Absolute testes weight Testicular degeneration

Dose (mg/kg-day)

signifcantly changed not significantly changed

1 Increased absolute weight of testes and epididymis 2Although 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.

Figure 1-4. Exposure response array of male reproductive effects following oral exposure.
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Summary of Reproductive and Developmental Effects

1 Developmental studies in rats (Angerhofer et al., 1986; Cholakis et al., 1980) and rabbits

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

(Cholakis et al., 1980) suggest that developmental effects related to offspring survival, growth, and 

morphological development were likely associated with severe maternal toxicity. Developmental 

effects were observed only at doses that caused maternal mortality. 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, EPA concluded that the evidence does not support developmental effects 

as a potential human hazard of RDX exposure.

Testicular effects were reported in male B6C3F1 mice chronically exposed to RDX in the diet 

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 (Levine et al., 1984). Other testicular effects 

were inconsistent across rat studies. Based on the evidence reported by Lish et al. (1984), EPA 

identified suggestive evidence of male reproductive effects as a potential human hazard of RDX 

exposure. 

1.1.4. Liver Effects

 The association between RDX exposure and changes in serum liver enzymes was examined 

in a single occupational epidemiology study. 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 1-10 and 1-11 and Figure 1-5.

Reports in humans provide limited evidence of liver toxicity associated with acute exposure 

to RDX. Elevated serum levels of aspartate aminotransferase (AST) and 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.3). 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., 1996b; 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; mean average exposure concentration was 0.28 mg/m3) 

(Hathaway and Buck, 1977), serum chemistry analysis (including the serum liver enzymes AST, 
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-31 DRAFT—DO NOT CITE OR QUOTE

1 ALT, and alkaline phosphatase (ALP)) revealed no statistically significant differences between 

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

35

36

37

38

exposed and unexposed workers (Table 1-10). 

In experimental animals, the most consistent noncancer liver effect associated with RDX 

exposure is elevated liver weight in studies of subchronic exposure (Table 1-11 and Figure 1-5). 

Dose-related increases in absolute and relative liver weight 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; Levine et al., 

1981b; Cholakis et al., 1980). In another 90-day study, only absolute liver weights were increased 

in female F344 rats exposed to RDX by gavage (Crouse et al., 2006). The magnitude of liver weight 

increases in B6C3F1 mice and female F344 rats across these studies ranged from 4−29% in the 

high-dose groups. Male F344 rats did not exhibit similar increases in liver weight in other 

subchronic studies (Crouse et al., 2006; Levine et al., 1990; Levine et al., 1981a; Levine et al., 1981b; 

Cholakis et al., 1980). In male and female monkeys exposed subchronically to RDX, absolute liver 

weights were increased (6−16% relative to control at 1 and 10 mg/kg-day) (Martin and Hart, 1974) 

and similarly 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 2-year studies 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, absolute liver weight showed no dose-related changes; however,

relative liver weights were increased in high-dose (40 mg/kg-day) males and females (by 11 and 

18% compared to controls, respectively) (Levine et al., 1983). The changes in relative liver weight 

likely reflected the depressed weight gain in the high-dose rats (2-30% in males and 10-15% in 

females). Based 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 

relative liver weights (expressed as organ to body weight ratios) were better modeled for 

quantitative analysis than organ weight alone, or organ-to-brain weight ratios. 

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., 1983; Levine et al., 1981a; Levine et al., 1981b; Hart, 1974; Martin 

and Hart, 1974; Von Oettingen et al., 1949), 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., 1983). The few 

findings of liver lesions were reported in studies with more limited histopathological analyses, and 

were not confirmed in the studies with more complete histopathologic examination and longer 

exposure durations (Levine et al., 1984; Lish et al., 1984; Levine et al., 1983; Thompson, 1983; Von 

Oettingen et al., 1949). For example, the incidence of liver portal inflammation was increased in 
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1-32 DRAFT—DO NOT CITE OR QUOTE

1 female but not male rats exposed to 40 mg/kg-day in the diet for 90 days (Cholakis et al., 1980). 

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

35

36

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). In both the rat and mouse studies by Cholakis et al. 

(1980), groups sizes were relatively small (n = 10/sex/group) and histopathologic findings were 

reported for the control and high-dose groups only. 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. 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; Levine et 

al., 1981b). In a chronic toxicity study by the same investigators (Levine et al., 1983), 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., 1983), in rats exposed to RDX by gavage for 90 days at doses up to 

15 mg/kg-day (Crouse et al., 2006), and in monkeys exposed to RDX in the diet for 90 days (Martin 

and Hart, 1974) were relatively small (within 38% of control mean) and were not dose related.
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1-33 DRAFT—DO NOT CITE OR QUOTE

Table 1-10. 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: less than 

the limit of detection (LOD) or 

≥ 0.01 mg/m3 (mean 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). 

Liver function tests in men; mean (standard deviation not reported)

Test

Referent

(n = 237)

RDX exposed

Undetected

(n = 22)

>0.01 mg/m3

(n = 45)

LDH 173 191 174

Alkaline 

phosphatase

82 78 80

ALA (SGOT) 22 25 21

AST (SGPT) 21 26 18

Bilirubin 0.5 0.4 0.4

No differences were statistically significant as reported by study 

authors. Similar results in women.

Liver function tests in men: prevalence of abnormal values

Test

(abnormal 

range) Referent

RDX exposed

Undetected >0.01 mg/m3

LDH (>250) 2/237 1/22 0/45

Alkaline 

phosphatase 

(>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

No differences were statistically significant as reported by study 

authors. Similar results in women.
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1-34 DRAFT—DO NOT CITE OR QUOTE

Table 1-11. Evidence pertaining to liver effects in animals

Reference and study design Results

Liver weight

Lish et al. (1984); Levine 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

24 mo

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.

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. (1983); Thompson (1983)

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

24 mo

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%

Doses 0 80 160 320
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1-35 DRAFT—DO NOT CITE OR QUOTE

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

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

13 wks

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

90 d

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%

Levine et al. (1981a); Levine et al. 

(1990); Levine et al. (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 

Data were not reported for rats in the 300 or 600 mg/kg-d dose groups 

because all of the rats died before the 13-wk necropsy.

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%* − −
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Reference and study design Results

Diet

13 wks

Relative liver weight (percent change compared to control)

M 0% 9% 6% 20% − −

F 0% 3% 5% 19%* − −

Hart (1974)

Dogs, Beagle, 3/sex/group

Pre-mix with ground dog chow 

containing 20 mg RDX/g-chow, 

60 grams of dog food

0, 0.1, 1, or 10 mg/kg-d 

Diet

90 d

Doses 0 0.1 1 10

Absolute liver weight (percent change compared to control)

M 0% − − 53%

F 0% − − 3%

Martin and Hart (1974)

Monkeys, Cynomolgus or Rhesus, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d 

Gavage

90 d

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); Levine 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

24 mo

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. 

Levine et al. (1983); Thompson (1983)

Rats, F344, 3−4 wks old; 75/sex/group; 

interim sacrifices (10/sex/group) at 

6 and 12 mo

Doses 0 0.3 1.5 8.0 40

Microgranulomas (incidence)

M 0/38 0/36 0/25 0/29 0/4
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1-37 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

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

24 mo

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

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

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

90 d

Histopathology examination of the 15 mg/kg-d group showed one male 

rat with mild liver congestion and one female rat with a moderate-sized 

focus of basophilic cytoplasmic alteration; neither finding was attributed 

by study authors to RDX treatment.

Levine et al. (1981a);Levine et al. 

(1990); Levine et al. (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. 
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1-38 DRAFT—DO NOT CITE OR QUOTE

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 

grams of dog food

0, 0.1, 1, or 10 mg/kg-d 

Diet

90 d

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 Rhesus, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d 

Gavage

90 d

An increase in the amount of iron-positive material in liver cord 

cytoplasm was reported in monkeys treated with 10 mg/kg-d RDX; 

however, the study authors considered the toxicological significance to 

be uncertain.

Serum chemistry

Lish et al. (1984); Levine 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

24 mo

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. (1983); Thompson (1983)

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

24 mo

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

90 d

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%

F 0% −16% −21% 7% −37% 18%
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1-39 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

Levine et al. (1981a);Levine et al. 

(1990); Levine et al. (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

Data not reported for 300 and 600 mg/kg-d dose groups because all of 

the animals died before the 13-wk blood sampling.

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 Rhesus, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d 

Gavage

90 d

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 

*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.
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-40 DRAFT—DO NOT CITE OR QUOTE

M

M M

M

M M

M

M

M

M

M M

M

M

M

M

M

M

M

M

0.1

1

10

100

1000

Lish et al., 1984 (mouse)

Hart, 1976 (rat)

Levine et al., 1983 (rat, male)

 Levine et al., 1983 (rat, female)

Cholakis et al., 1980 (mouse)Cholakis et al., 1980 (mouse)

Cholakis et al., 1980 (rat)

 Crouse et al., 2006 (rat)

 Levine et al., 1990, 1981a,b (rat, male)

 Levine et al., 1990, 1981a,b (rat, female)

Lish et al., 1984 (mouse)

Hart, 1976 (rat)

Levine et al., 1983 (rat)

Cholakis et al., 1980 (mouse)Cholakis et al., 1980 (mouse)

Crouse et al., 2006 (rat)

Levine et al., 1990, 1981a,b (rat)

Hart, 1974a (dog)

Martin and Hart, 1974b (monkey)

Lish et al., 1984 (mouse)

Levine et al., 1983 (rat)

 Crouse et al., 2006 (rat, male)

 Crouse et al., 2006 (rat, female)

Martin and Hart, 1974b (monkey)

Lish et al., 1984 (mouse)

 Levine et al., 1983 (rat, male)

 Levine et al., 1983 (rat, female)

Crouse et al., 2006 (rat)

Levine et al., 1990, 1981a,b (rat)

Chronic Subchronic Chronic Subchronic Chronic Subchronic Chronic Subchronic

↓ Cholesterol ↓ Trigylcerides

↑ Relative liver weight Histopathology Serum biochemistry changes

Dose (mg/kg-day)

significantly changed not significantly changed not determined

X- not determined due to confounding caused by presence of tumors

These studies were excluded from array because only absolute liver weight was reported: Cholakis, 1980 (2-gen rat); Hart, 1974; Martin and Hart, 1974

M - Mortality observed at this dose and above

Figure 1-5. Exposure response array of liver effects following oral exposure.
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1-41 DRAFT—DO NOT CITE OR QUOTE

Summary of Liver Effects

1 There is limited evidence from reports of human exposure and from studies in experimental 

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

35

animals that RDX may affect the liver. Several human case reports of short-term elevations of

serum liver enzymes in individuals who ingested unknown amounts of RDX suggest that RDX might 

target the liver; however, serum liver enzymes were not elevated in a small prevalence study of 

munition plant workers exposed to RDX. In experimental animals, dose-related increases in 

relative or absolute liver weight were observed in multiple studies following subchronic oral 

exposure, in multiple species (mice, rats, dogs, and monkeys), and in both sexes; however, an 

association between RDX exposure and increased liver weight was not similarly supported by 

lifetime studies in mice and rats. Changes in serum liver enzymes 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. EPA concluded that the evidence does not support liver effects as a potential 

human hazard of RDX exposure. 

1.1.5. 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., 1983; Hart, 1976). The 

2-year studies by Lish et al. (1984) and Levine et al. (1983) were performed in accordance with 

FDA Good Laboratory Practice regulations (FDA, 1979) and included comprehensive 

histopathological examination of major organs, multiple dose groups and a control, and more than 

50 animals/dose group (plus additional interim sacrifice groups). The Hart (1976) study is largely 

limited by lack of characterization of the test material and pathology analysis limited to the control 

and high-dose groups. 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 more 

than 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 1-12 and 1-13. 

Liver tumors

Increased incidence of liver tumors was observed in one chronic mouse study and one of 

two chronic rat studies. In the chronic mouse dietary study (Lish et al., 1984), the combined 

incidences of hepatocellular adenomas or carcinomas were increased with increasing RDX doses in 

female B6C3F1 mice as compared to concurrent controls, but not in male B6C3F1 mice similarly 

exposed to RDX for 2 years. In addition, the incidence of hepatocellular carcinomas showed a cant 

positive trend with dose in male, but not female, F344 rats exposed to RDX in the diet for 2 years 

(Levine et al., 1983) (Cochran-Armitage trend test performed for this review, p = 0.032). On the 
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1-42 DRAFT—DO NOT CITE OR QUOTE

1 other hand, there were no increased incidences of hepatocellular adenomas or carcinomas in 

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

Sprague-Dawley rats of either sex exposed to RDX via diet for two years at doses up to 10 mg/kgday (Hart, 1976). Incidences of hepatocellular neoplasms are presented in Table 1-12. The tumor 

responses are discussed in further detail below.

In the female B6C3F1 mouse study by Lish et al. (1984), the finding of a statistically 

significant increase in hepatocellular tumors may have been influenced by the incidence of

hepatocellular adenomas/carcinomas in the concurrent female control mice, which the study 

authors noted was relatively low (1/65). However, as noted by the authors, the incidence of 

hepatocellular adenomas or carcinomas at RDX doses ≥35 mg/kg-day (19% at both doses) was also 

statistically significantly elevated when compared to the mean historical control incidence for 

female B6C3F1 mice in National Toxicology Program (NTP) studies (147/1781 or 8%; range: 

0−20%) (Haseman et al., 1985).5

A Pathology Working Group (PWG) review of the slides of female mouse liver lesions from 

the Lish et al. (1984) study resulted in some changes in lesion diagnosis (Parker et al., 2006; Parker, 

2001). Some malignant tumors were downgraded to benign status and several lesions initially 

characterized as tumors were changed to non-tumors based on more recent diagnostic criteria used 

by the PWG (Harada et al., 1999). There was a statistically significant trend in the combined 

incidence of hepatocellular adenomas or carcinomas (using a Cochran-Armitage one-sided trend 

test performed by EPA), 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. 

As noted above, male F344 rats showed a positive trend with dose in the incidence of 

hepatocellular carcinomas in the Levine et al. (1983) bioassay; however, the association with 

exposure is not strong, in part reflecting the lower magnitude of response. There were only a few 

tumors observed in the exposed groups (0/55, 0/52, 2/55, 2/31) relative to the control (1/55), as 

compared with the mice. There is less confidence that the final incidence in the highest-dose group 

accurately reflects lifetime cancer incidence because of low survival and no time-to-death 

information to estimate mortality-adjusted incidences; the available information may 

underestimate lifetime cancer incidence by overestimating the number of rats truly at risk. Some 

perspective on the magnitude of response is provided by comparing with incidence rates in 

 

5Comparison 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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1-43 DRAFT—DO NOT CITE OR QUOTE

1 historical controls, despite the limitation of this comparison due to the historical data originating 

2 

3 

4 

5 

6 

7 

8 

9 

10

11

12

from a different laboratory. In a paper published concurrently with the Levine et al. (1983) study, 

the 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). The incidence of liver carcinomas in control male 

rats in Levine et al. (1983) was at the upper end of the NTP range, and higher than the NTP range in 

the highest two dose groups. Nonmalignant liver tumors (neoplastic nodules) in F344 male rats in 

the 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. (1983) reported a higher 

incidence of neoplastic nodules, 7.3%, in their control male rats, with a decline in incidence with 

increasing RDX exposure. Although there are several reasons to conclude that the observation of an 

association between RDX exposure and liver tumors in rats is not strong, this suggestive site 

concordance supports the response in female mice. 

Table 1-12. Liver tumors observed in chronic animal bioassays 

Reference and study design Results

Lish et al. (1984); Levine 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

24 mo

Doses 0 1.5 7.0 35 175/100

Hepatocellular adenomas (incidence)a

M 8/63 6/60 1/62* 7/59 7/27

F 1/65 1/62 6/64 6/64 3/31b

Hepatocellular carcinomas (incidence)a

M 13/63 20/60 16/62 18/59 6/27

F 0/65 4/62 3/64 6/64 3/31a

Hepatocellular adenoma or carcinoma combined (incidence) a

M 21/63 26/60 17/62 25/59 13/27

F 1/65 5/62 9/64* 12/64* 6/31*b

Pathology workgroup reanalysis of liver lesion slides from female mice 

(Parker et al., 2006; Parker, 2001)

c

Doses 0 1.5 7.0 35 175

Hepatocellular adenomas (incidence)a

F 1/67 3/62 2/63 8/64 2/31b

Hepatocellular carcinomas (incidence)a

F 0/67 1/62 3/63 2/64 2/31b

Hepatocellular adenoma or carcinoma combined (incidence)a

F 1/67c 4/62 5/63c 10/64 4/31b

Hart (1976)

Rats, Sprague-Dawley, 100/sex/group

Doses 0 1.0 3.1 10

Neoplastic nodules (incidence)a
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Reference and study design Results

Purity and particle size not specified

0, 1.0, 3.1, or 10 mg/kg-d 

Diet

2 yrs

M 0/82 − − 3/77

F 1/72 − − 1/81

Hepatocellular carcinomas (incidence)a

M 1/82 − − 1/77

F 1/72 − − 1/81

Neoplastic nodules or hepatocellular carcinomas combined

(incidence)a

M 1/82 − − 4/77

F 2/72 − − 2/81

Levine et al. (1983); Thompson (1983)

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

24 mo

Doses 0 0.3 1.5 8.0 40

Neoplastic nodules (incidence)a

M 4/55 3/55 0/52 2/55 1/31

F 3/53 1/55 1/54 0/55 4/48

Hepatocellular carcinomas (incidence)a

M 1/55 0/55 0/52 2/55 2/31b

F 0/53 1/55 0/54 0/55 0/48

Neoplastic nodules or hepatocellular carcinomas combined

(incidence)a

M 5/55 3/55 0/52 4/55 3/31

F 3/53 2/55 1/54 0/55 4/48

1 

2 

3 

4 

5 

6 

7 

*Statistically significant difference compared to the control group (p < 0.05), identified by the authors. a

The incidences reflect the animals surviving to month 12. b

Statistically significant trend (p < 0.05) was identified using Cochran-Armitage trend tests performed by EPA. c

It is not clear why the numbers of animals at risk in the control group (n = 67) and 7 mg/kg-day dose group (n = 63) 

differed from the numbers reported in the original study (n = 65 and 64, respectively). 
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1-45 DRAFT—DO NOT CITE OR QUOTE

1 Lung tumors

2 

3 

4 

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9 

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16

Cochran-Armitage trend tests (as performed for this review, p = 0.019) found statistically 

significant positive trends in the incidences of alveolar/bronchiolar adenomas in female B6C3F1

mice, alveolar/bronchiolar carcinomas in male mice, and alveolar/bronchiolar adenomas or 

carcinomas combined in female mice. The combined incidence in male B6C3F1 mice did not show a 

statistically significant trend (see Table 1-13). 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. However, the authors regarded these neoplasms as random and not biologically significant. 

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/kgday (Hart, 1976) or in F344 rats of either sex exposed chronically to RDX at doses up to 40 mg/kgday (Levine et al., 1983). 
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1-46 DRAFT—DO NOT CITE OR QUOTE

Table 1-13. Lung tumors observed in chronic animal bioassays 

Reference and study design Results

Lish et al. (1984); Levine 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

24 mo

Doses 0 1.5 7.0 35 175/100

Alveolar/bronchiolar adenomas (incidence)a

M 

F 

6/63

4/65

5/60

2/62

5/62

5/64

7/59

9/64

1/27

3/31b

Alveolar/bronchiolar carcinomas (incidence)a

M 

F 

3/63

3/65

6/60

1/62

3/62

3/64

7/59

3/64

5/27b

4/31

Alveolar/bronchiolar adenoma or carcinoma combined (incidence)a

M 

F 

9/63

7/65

11/60

3/62

8/62

8/64

14/59

12/64

6/27

7/31b

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 

F 

2/83

0/73

−

−

−

−

1/77

0/82

Alveolar/bronchiolar carcinoma 

None reported by study authors.

(incidence)

Levine et al. (1983); Thompson (1983)

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

24 mo

Doses 0 0.3 1.5 8.0 40

Alveolar/bronchiolar adenomas (incidence)a

M 

F 

1/55

3/53

0/15

0/7

1/17

0/8

0/16

1/10

1/31

0/48

Alveolar/bronchiolar carcinomas (incidence)a

 

M 

F 

−

0/53

−

0/7

−

1/8

−

0/10

−

0/48

Alveolar/bronchiolar adenoma or carcinoma combined (incidence)a

M − − − − −

F 3/53 0/7 1/8 1/10 0/48

1 

2 

3 

a

The incidences reflect the animals surviving to month 12. b

Statistically significant trend (p < 0.05) was identified using Cochran-Armitage trend test performed by EPA.
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1-47 DRAFT—DO NOT CITE OR QUOTE

Mechanistic Evidence

1 There are few mechanistic data to support a MOA determination for either liver or lung 

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tumors induced by exposure to RDX. 

The increase in liver weights observed in subchronic studies of RDX in mice (Cholakis et al., 

1980) and rats (Levine et al., 1990; Levine et al., 1981a; Levine et al., 1981b; Cholakis et al., 1980) 

and chronic studies in female B6C3F1 mice (Lish et al., 1984; Cholakis et al., 1980) raises the 

possibility of RDX-related liver cell proliferation as a precursor to tumorgenicity. Sweeney et al. 

(2012b) reviewed hypothesized MOA’s for carcinogenicity and concluded that a MOA involving a 

proliferative response generated by tissue-derived oxidative metabolites of RDX was the most 

plausible of the MOAs considered, but acknowledged that the overall support for this MOA was 

limited. The following lines of evidence do not support a metabolite-based proliferative response 

as the MOA for RDX carcinogenicity:

• 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; 

• increased liver weight was also observed in rats and male mice where tumors did not occur;

• no studies were available that directly measured RDX-induced cell proliferation rates; and 

• no information was available to rule out non-precancerous causes of liver weight increase.

The available in vitro and in vivo genotoxicity assay results are largely negative for parent 

RDX (see Appendix C, Section C.4), supporting the hypothesis that parent RDX does not interact 

directly with DNA. Sweeney et al. (2012b) proposed that the increased incidence of liver adenomas 

and carcinomas in female mice (Parker et al., 2006; Lish et al., 1984) may result from livergenerated metabolites as the most likely agents responsible for liver tumors. Sweeney et al. 

(2012b) estimated an approximately 30-fold higher metabolic rate for RDX in mice (which 

displayed a more robust liver tumor response to RDX exposure than did rats) compared with rats

based on the results of a PBPK model. These authors hypothesized a non-linear, cell proliferation

MOA in conjunction with the lack of evidence to support a genotoxic/mutagenic MOA for RDX or its 

oxidative metabolites. Sweeney et al. (2012b) suggest that RDX is unlikely to be genotoxic because 

it does not induce tumors at multiple sites and species. This observation is inconsistent with the 

finding in Lish et al. (1984) that showed positive trends in the incidence of both 

alveolar/bronchiolar adenomas or carcinomas and liver tumors.

In contrast to the negative results for RDX oxidative metabolites, there are some positive 

genotoxicity results for the N-nitroso metabolites of RDX, specifically hexahydro-1-nitroso3,5-dinitro-1,3,5-triazine [MNX) and hexahydro-1,3,5-trinitroso-1,3,5-triazine (TNX). MNX and 

TNX have been identified from minipigs; minipigs were chosen as the animal model for the 

metabolism of RDX because the gastrointestinal tract of pigs more closely resembles that of humans
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1-48 DRAFT—DO NOT CITE OR QUOTE

1 (Musick et al., 2010; Major et al., 2007). MNX has tested positive in some in vitro assays, including 

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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 S. typhimurium

(strains TA98, TA100, TA1535, TA1537, and TA1538), with or without the addition of the S9 

metabolic activating mixture (Pan et al., 2007; 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 gastrointestinal tract to various Nnitroso derivatives (Pan et al., 2007). Exposure to potentially mutagenic N-nitroso metabolites of 

RDX generated in the gastrointestinal 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-diazbutanal and 4-nitro-2,4-diazbutanamide. Thus, the contribution of the N-nitroso 

metabolites to the overall carcinogenic potential of RDX is unclear. 

Aberrant expression of microRNAs (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), with several 

oncogenic miRNAs being upregulated, while several tumor-suppressing miRNAs were down 

regulated. 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); however, the utility of miRNAs as predictive of 

carcinogenesis has not been established, and whether or not aberrant expression of a specific 

miRNA (or suite of miRNA’s) plays a role in the MOA of RDX carcinogenicity is unknown. 

Microarray analysis of gene expression in male Sprague-Dawley rats after exposure to a single oral 

(capsule) dose of RDX revealed a general up-regulation in gene expression (predominantly genes 

involved in metabolism) in liver tissues (Bannon et al., 2009); however, the relevance of this finding 

to the carcinogenicity of RDX is unclear. 

In summary, the available evidence indicates that RDX is not mutagenic (see Appendix C,

Section C.4); however, anaerobically-derived N-nitroso metabolites have demonstrated some 

genotoxic potential. While these metabolites have been measured in the mouse (Pan et al., 2007) 

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 events leading to the observed 

liver and lung tumor response in RDX-exposed rodents and lack of toxicokinetic information on 

RDX metabolites, neither a cell proliferative MOA or a mutagenic N-nitroso metabolite MOA is 
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1-49 DRAFT—DO NOT CITE OR QUOTE

1 supported. Thus, the MOA leading to the increased incidence of liver and lungs tumors is not 

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known. 

1.1.6. Other Toxicological Effects

There is limited evidence that RDX can produce systemic effects in several organs/systems, 

including the eyes, and the musculoskeletal, cardiovascular, immune, and gastrointestinal systems. 

However, there is less evidence for these effects compared to organ systems described earlier in 

Section 1.1. A summary of the evidence for toxicological effects in other organ systems is shown in 

Tables 1-14 and 1-15. 

Ocular Effects

There are no reports of ocular effects in human case reports or epidemiological studies. 

The incidence of cataracts was statistically significantly increased in high-dose female rats in one 

chronic oral study; however, this finding was not reproduced in other subchronic and chronic 

studies in rats or mice. 

The incidence of cataracts was 73% in female F344 rats exposed to 40 mg/kg-day RDX in 

the diet for 2 years, compared to 32% in the control group (Levine et al., 1983). After 76 weeks of 

exposure, the incidence of cataracts in female rats at 40 mg/kg-day (23%) was also elevated 

compared to controls (6%). The incidence of cataracts was not increased in RDX-exposed male rats 

in the same study (Levine et al., 1983), and other studies have not observed ocular effects 

associated with RDX exposure. Only 2 rats (dose groups not reported) were observed to have mild 

cataracts in a 90-day study of male and female F344 rats exposed to RDX at doses up to 15 mg/kgday by gavage; however, the authors noted that these observations are common in F344 rats at

4 months of age and should not be attributed to treatment (Crouse et al., 2006). Furthermore, 

cataracts were not observed in male or female F344 rats exposed to 40 mg/kg-day RDX by diet for 

90 days (Cholakis et al., 1980), or in male or female B6C3F1 mice exposed to RDX in the diet for 

2 years at doses up to approximately 100 mg/kg-day (Lish et al., 1984). A statistically significant 

increase in the incidence of cataracts in male mice was initially noted by Lish et al. (1984), but was 

not confirmed when mice used for orbital bleedings were excluded from the analysis, suggesting 

the effect was not treatment related. 

Cardiovascular Effects

Human evidence for cardiovascular effects is limited to case reports that include 

observations of transient arterial hypertension in male Italian workers following inhalation of RDX 

during manufacturing (Barsotti and Crotti, 1949), sinus tachycardia, and in one instance premature 

ventricular beats in 5 men following accidental ingestion of RDX at 37−250 mg/kg body weight 

(Küçükardali et al., 2003) (see Appendix C, Section C.3). 

Inconsistent observations of cardiovascular effects have been reported in animal studies. 

An increase in the relative heart-to-body weight ratio was observed at the highest dose tested in 
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1-50 DRAFT—DO NOT CITE OR QUOTE

1 B6C3F1 mice (male: 13%; female 17%) and F344 rats (male: 22%; female 15%) following chronic 

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dietary administration of RDX (Lish et al., 1984; Levine et al., 1983); however, this dose also 

resulted in reductions in body weight in both males and females. Dose-related decreases in 

absolute heart weight were reported following subchronic exposures to RDX in the diet (Levine et 

al., 1990; Levine et al., 1981a; Levine et al., 1981b; Cholakis et al., 1980), while a subchronic study in 

male dogs reported a 31% increase in absolute heart weight at the highest dose tested 

(10 mg/kg-day) (Hart, 1974). 

Evidence for histopathologic changes associated with RDX exposure is limited to findings of

an increased incidence of focal myocardial degeneration in female rats compared to controls (60 vs. 

20%, respectively) and male mice (50 vs. 0%, respectively) following exposure to RDX in the diet

for 90 days (Cholakis et al., 1980). In each study, the finding of myocardial degeneration was

limited to one sex and to the high-dose group only. Other studies in monkeys (Martin and Hart, 

1974) and rats (Von Oettingen et al., 1949) reported no observable cardiovascular effects. 

Musculoskeletal Effects

Evidence of musculoskeletal effects in humans consists of case reports that include 

observations of muscle twitching, myalgia/muscle soreness, and muscle injury as indicated by 

elevated levels of AST or myoglobinuria (Küçükardali et al., 2003; Hett and Fichtner, 2002; 

Hollander and Colbach, 1969; Stone et al., 1969; Merrill, 1968) (see Appendix C, Section C.3).

Histological evaluations of musculature or skeletal tissue did not reveal any alterations in 

mice (Lish et al., 1984) or rats (Levine et al., 1983; Hart, 1976) following chronic oral exposure to 

RDX, in mice and rats following subchronic exposure (Cholakis et al., 1980), or in dogs following a 

90-day dietary exposure (Hart, 1974). 

Immune Effects

RDX is structurally similar to various drugs known to induce the autoimmune disorder 

systemic lupus erythematosus (SLE). Three cases of SLE were initially identified among workers at 

one U.S. Army munitions plant; however, upon further investigation of 69 employees at five U.S. 

Army munitions plants with potential exposure to RDX, no additional cases of SLE were identified

(Hathaway and Buck, 1977). Increased white blood cell (WBC) counts have been reported in some 

case reports of individuals who ingested RDX or C-4 (91% RDX) (Knepshield and Stone, 1972; 

Hollander and Colbach, 1969; Stone et al., 1969; Merrill, 1968). 

In animal studies, increased WBC count in female rats following subchronic dietary 

exposure to RDX was the only dose-related immune effect reported (Levine et al., 1990; Levine et 

al., 1981a; Levine et al., 1981b); WBC counts in male rats were unaffected. Conversely, decreased 

WBC counts in were reported in male and female rats in a 2-year study (Hart, 1976). Changes in 

spleen weights were observed across studies, but the responses were not consistent and did not

appear to be dose-related. For example, in 90-day studies, Cholakis et al. (1980) identified a 

statistically significant decrease in absolute spleen weight in F344 rats at 40 mg/kg-day, while
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1 Crouse et al. (2006) observed an increase in spleen weight at 15 mg/kg-day (not statistically 

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significant). Across studies, there was no significant or dose-dependent pattern of response to 

suggest that the WBC changes reflect RDX-induced immunotoxicity. No dose-related immune 

effects from oral exposure to RDX were observed in other animal studies, including a 90-day study 

in F344 rats specifically designed to evaluate immunotoxicity (parameters included evaluation of 

red and white blood cell populations, proportion of cell surface markers, cellularity in proportion to 

organ weight, B and T cells in the spleen, and CD4/CD8 antigens of maturing lymphocytes in the 

thymus) (Crouse et al., 2006). Routine clinical and histopathology evaluations of immune-related 

organs in a two-generation study in rats (Cholakis et al., 1980) and chronic studies in rats (Levine et 

al., 1983) and mice (Lish et al., 1984) provide no evidence of immunotoxicity associated with oral 

(dietary) exposure to RDX. 

In summary, evidence for immunotoxicity associated with RDX exposure is limited to

findings from one study of increased WBC counts in female rats (Levine et al., 1981a; Levine et al., 

1981b). Evidence that RDX is not immunotoxic comes from several animal studies, including other 

repeat-dose oral studies in mice and rats (Crouse et al., 2006; Lish et al., 1984; Levine et al., 1983; 

Cholakis et al., 1980). 

Gastrointestinal Effects

Clinical signs of nausea and/or vomiting have been frequently identified in case reports of 

accidental or intentional RDX poisonings, and generally concurrent with severe neurotoxicity

(Kasuske et al., 2009; Davies et al., 2007; Küçükardali et al., 2003; Hett and Fichtner, 2002; Ketel 

and Hughes, 1972; Knepshield and Stone, 1972; Hollander and Colbach, 1969; Stone et al., 1969; 

Merrill, 1968; Kaplan et al., 1965; Barsotti and Crotti, 1949) (see Appendix C, Section C.3). In 

animal studies, nausea and vomiting have also been observed following oral exposure of swine

(Musick et al., 2010), dogs (Hart, 1974), and monkeys (Martin and Hart, 1974). One subchronic oral 

(diet) rat study from the early literature reported congestion of the gastrointestinal tract at doses 

also associated with elevated mortality (Von Oettingen et al., 1949); however, none of the 

subsequent subchronic or chronic animal studies reported histological findings of the 

gastrointestinal tract related to RDX administered via gavage or the diet (Crouse et al., 2006; Lish et 

al., 1984; Levine et al., 1983; Hart, 1974; Martin and Hart, 1974). 

Hematological Effects

Elevated prevalence odds ratios (OR) for hematological abnormalities were observed in a 

case-control study of males (32 exposed, 322 controls) exposed to RDX in an occupational setting

(West and Stafford, 1997) (see Table 1-14). The prevalence OR for an association between RDX 

exposure and hematological abnormalities was 1.7 (95% CI 0.7−4.2) for men with greater than 50 

hours of low intensity exposure, while the prevalence OR was 1.2 (95% CI 0.3−5.3) for men with 

>50 hours of high intensity exposure. The ORs from this study must be interpreted with caution 

given the small sample size and wide confidence intervals. No changes in hematological parameters 
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1-52 DRAFT—DO NOT CITE OR QUOTE

1 (including hemoglobin, hematocrit, and reticulocyte count) were observed in a cross-sectional 

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epidemiologic study of 69 workers exposed to RDX by inhalation (average of 0.28 mg/m3) 

(Hathaway and Buck, 1977). Humans who ingested or inhaled unknown amounts of RDX or C-4 

(~91% RDX) for an acute duration displayed temporary hematological alterations, including 

anemia, decreased hematocrit, hematuria, and methemoglobinemia (Kasuske et al., 2009; 

Küçükardali et al., 2003; Knepshield and Stone, 1972; Hollander and Colbach, 1969; Stone et al., 

1969; Merrill, 1968). In other case reports, normal blood counts were observed in accidentally 

exposed individuals (Testud et al., 1996b; Goldberg et al., 1992; Woody et al., 1986; Ketel and 

Hughes, 1972; Kaplan et al., 1965) (see Appendix C, Section C.3).

In animals, hematological alterations were observed following oral exposure in chronic and

subchronic studies in both sexes of rats (F344 or SD) and B6C3F1 mice (see Table 1-15). Increases 

in platelet count were observed in male and female mice and rats in some subchronic and chronic 

studies at doses from 0.3 mg/kg-day to 320 mg/kg-day (Lish et al., 1984; Levine et al., 1983; 

Cholakis et al., 1980); however, findings were generally inconsistent across studies and were not 

necessarily dose-dependent. Similarly, decreased hemoglobin levels/anemia were observed in 

some chronic and subchronic studies (Levine et al., 1983; Cholakis et al., 1980; Von Oettingen et al.,

1949), particularly at doses greater than or equal to 15 mg/kg-day, but trends in hemoglobin levels

across studies did not show a consistent relationship with dose. Other hematological parameters, 

including WBC counts, reticulocyte counts, and hematocrit, showed conflicting results between 

studies, marginal responses, or inconsistent changes with increasing dose. Other subchronic 

studies in rats and dogs (Crouse et al., 2006; Hart, 1974; Von Oettingen et al., 1949) did not identify

any changes in hematological parameters. 

In summary, evidence for hematological effects associated with RDX exposure in humans 

comes from several case reports that found transient fluctuations in hematological endpoints after 

acute exposures. Hematological findings from two epidemiological studies were inconsistent and 

difficult to interpret because of small sample sizes (Table 1-14). In general, animal studies of 

chronic and subchronic durations showed no consistent, dose-related pattern of increase or 

decrease in hematological parameters. 

Table 1-14. Evidence pertaining to systemic effects (hematological) in 

humans

Reference and study design Results

Hematological effects

West and Stafford (1997) (United Kingdom)

Case-control study, 32 cases with abnormal 

and 322 controls with normal hematology test 

drawn from 1991 study of 404 workers at 

ammunitions plant; participation rate 97% of 

Odds ratio (95% CI) [number of exposed cases] of blood 

disorder and RDX

Low intensity, 50 hr-duration 1.7 (0.7,4.2) [22]

Medium intensity, 50-hr 

duration

1.6 (not reported) [5]
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1-53 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

Hematological effects

cases, 93% of controls. Analysis limited to men 

(29 cases, 282 controls).

Exposure measures: Exposure determination 

based on employee interviews and job title 

analysis; data included frequency (hrs/d, d/yr), 

duration (yrs), and intensity (low [1−10 ppm], 

moderate [10−100 ppm], and high 

[100−1,000 ppm], based on ventilation 

considerations).

Effect measures: Hematology tests; blood 

disorder defined as neutropenia (2.0 × 109

/L), 

low platelet count (<150 × 109

/L), or

macrocytosis (mean corpuscular volume = 99 fl 

or >6% macrocytes).

Analysis: Unadjusted odds ratio.

High intensity, 50-hr duration 1.2 (0.3, 5.3) [2]

Hathaway and Buck (1977) (United States)

Cross-sectional study, 2,022 workers, 

1,491 participated (74% response rate). 

Analysis 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 

0.28 mg/m3

).

Effect measures: Hematology 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). 

Hematology tests in men; mean (standard deviation not 

reported)

Test

Referent

(n = 237)

RDX exposed

Undetected

(n = 22)

>0.01 mg/m3

(n = 45)

Hemoglobin 15.2 14.7 15.2

Hematocrit 42 45.6 47

Reticulocyte 

count 

0.7 0.9 0.7

No differences were statistically significant. Similar results in 

women.

Hematology tests in men: prevalence of abnormal values

Test

(abnormal 

range) Referent

RDX exposed

Undetected >0.01 mg/m3

Hemoglobin 

(<14)

15/237 3/22 4/45

Hematocrit 

(<40)

1/237 1/22 1/45

Reticulocyte 

count (>1.5)

18/237 3/22 2/45

No differences were statistically significant. Similar results in 

women.
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1-54 DRAFT—DO NOT CITE OR QUOTE

Table 1-15. Evidence pertaining to systemic effects in animals 

Reference and study design Results

Ocular effects

Lish et al. (1984); Levine 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

24 mo

Doses 0 1.5 7.0 35 175/100

Cataracts; 103 wks (incidence)a

M 2/47 2/41 0/41 2/37 2/16

F 2/50 1/37 6/52 0/46 1/26

Levine et al. (1983); Thompson (1983)

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

24 mo

Doses 0 0.3 1.5 8.0 40

Cataracts; 103 wks (incidence)

M 8/40 6/39 6/31 8/35 2/6

F 14/44 4/48 11/44 8/43 22/30*

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 ocular effects were observed (gross examination of eye was 

performed in all animals, and microscopic examination in control and 

40 mg/kg-d animals).

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d 

Gavage

90 d

No ocular effects were observed (ophthalmic examinations were 

performed in all animals within 1 wk of sacrifice, and microscopic 

examination of the eye was performed in control and 15 mg/kgdanimals).

Martin and Hart (1974)

Monkeys, Cynomolgus or Rhesus, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d 

Gavage

90 d

No ocular effects were observed (ophthalmoscopic examination was 

performed at the end of exposure).

Cardiovascular effects

Lish et al. (1984); Levine et al. (1984) Doses 0 1.5 7.0 35 175/100
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-55 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

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

24 mo

Absolute heart weight; 104 wks (percent change compared to control)

M 0% 4% 4% 5% 7%

F 0% 1% 5% 2% −5%

Relative heart-to-body weight; 104 wks (percent change compared to 

control)

M 0% 7% 5% 5% 13%*

F 0% 0% 6% 4% 17%*

Body weight was significantly lower at termination in males and 

females exposed to 175/100 mg/kg-d (−5 and −19%, respectively).

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

Myocardial fibrosis (percent incidence; number not reported)

M 20% − − 5%

F 5% − − 1%

Endocardial disease (percent incidence; number not reported)

M 1% − − 3%

F 0% − − 0%

Absolute heart weight; 104 wks (percent change compared to control)

M 0% −6% −2% −5%

F 0% 13% 3% 15%

Relative heart-to-body weight; 104 wks (percent change compared to 

control)

M 0% −2% 4% 1%

F 0% 23% 13% 27%

Levine et al. (1983); Thompson (1983)

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

24 mo

Doses 0 0.3 1.5 8.0 40

Absolute heart weight; 104 wks (percent change compared to control)

M 0% 3% −2% −2% 1%

F 0% −1% 0% −4% −3%

Relative heart-to-body weight; 104 wks (percent change compared to 

control)

M 0% 2% 6% 0% 22%

F 0% −2% 3% −1% 15%

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

Doses 0 10 14 20 28 40

Absolute heart weight (percent change compared to control)

M 0% − − − 7% 7%

F 0% − − − 0% 0%
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-56 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

Experiment 1: 0, 10, 14, 20, 28, or 

40 mg/kg-d 

Diet

13 wks

Relative heart weight (percent change compared to control)

M 0% − − − 6% 0%

F 0% − − − −4% 0%

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)b

Diet

13 wks

Doses 0 80 160 320

Focal myocardial degeneration (incidence)

M** 0/10 − − 5/10*

F*** 0/11 − − 2/11

Absolute heart weight (percent change compared to control)

M 0% 0% 0% 8%

F 0% 0% 0% 8%

Relative heart-to-body weight (percent change compared to control)

M 0% 0% −2% 6%

F 0% 0% −2% 2%

**Includes one affected and three unaffected animals that died 

prematurely.

***Includes one unaffected animal that died prematurely.

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

Focal myocardial degeneration (incidence)

M 3/10 − − − − 1/10

F 2/10 − − − − 6/10

Absolute heart weight (percent change compared to control)

M 0% − − − 0% −8%*

F 0% − − − −6% −11%*

Relative heart-to-body weight (percent change compared to control)

M 0% − − − 3% 0%

F 0% − − − −3% −8%

Relative heart-to-brain weight (percent change compared to control)

M 0% − − − −4% −10%*

F 0% − − − −5% −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

No cardiac effects were observed (microscopic examination of heart 

was performed in randomly selected F2 animals). 

Heart weight data were reported only for F2 generation controls, 5 and 

16 mg/kg-day groups.

Doses 0 5 16 50

Absolute heart weight (percent change compared to control)
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-57 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

F0 and F1 parental animals: 0, 5, 16, or 

50 mg/kg-d 

Diet

13 wks

F2 M 0% 3.2% −6.5% −

F2 F 0% 15% −3.7% −

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d 

Gavage

90 d

Doses 0 4 8 10 12 15

Cardiomyopathy (incidence)

M 2/10 − − − − 3/8

F 0/10 − − − − 1/6

Absolute heart weight (percent change compared to control)

M 0% −2% −7% −1% 1% 11%

F 0% −2% 0% 8% 7% 6%

Relative heart-to-body weight (percent change compared to control)

M 0% 4% 2% 1% −1% 8%

F 0% −2% −7% −6% −9% −16%*

Levine et al. (1981a);Levine et al. (1990); 

Levine et al. (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

All animals in the 300 and 600 mg/kg-d groups died prior to study 

termination.

Doses 0 10 30 100 300 600

Chronic focal myocarditis (incidence)

M 8/30 8/10 6/10 1/10 1/10 0/10

F 8/30 3/10 1/10 1/10 1/10 1/9

Absolute heart weight (percent change compared to control)

M 0% −2% −10% −15% − −

F 0% −3% 0% −5% − −

Relative heart-to-body weight (percent change compared to control)

M 0% 2% −4% 3% − −

F 0% −2% 0% −3% − −

Von Oettingen et al. (1949)

Rats (sex/strain not specified); 20/group

Purity and particle size not specified

0, 15, 25, or 50 mg/kg-d 

Diet

3 mo

The study authors reported that there were no cardiac effects 

(microscopic examination of the heart was performed in all rats; data 

were not shown).

Hart (1974)

Dogs, Beagle, 3/sex/group

Pre-mix with ground dog chow 

containing 20 mg RDX/g-chow, 60 grams 

of dog food

Doses 0 0.1 1 10

Focal hyalinization of the heart (incidence)

M 0/3 − − 0/3

F 0/3 − − 1/3
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-58 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

0, 0.1, 1, or 10 mg/kg-d

Diet

90 d

Absolute heart weight (percent change compared to control)

M 0% − − 31%

F 0% − − 5.7%

Martin and Hart (1974)

Monkeys, Cynomolgus or Rhesus, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d 

Gavage

90 d 

Doses 0 0.1 1 10

Myocarditis (percent change compared to control)

M 1/3 − − 1/3

F 0/3 − − 0/3

Absolute heart weight (percent change compared to control)

M 0% 7% −1% 5%

F 0% 10% 12% −12%

Immune effects

Lish et al. (1984); Levine 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

24 mo

No immune effects were observed with routine hematology, clinical 

chemistry, or histopathology evaluations.

Doses 0 1.5 7.0 35 175/100

WBC count; 105 wks (percent change compared to control)

M 0% −13% −8% −16% −30%

F 0% 12% 39%* 28% 0%

Absolute spleen weight; 105 wks (percent change compared to 

control)

M 0% 24% 31% −10% −28%

F 0% 4% 15% −17% 16%

Relative spleen weight; 105 wks (percent change compared to control)

M 0% 26% 32% −11% −21%

F 0% 4% 15% −17% 44%

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

WBC count; 104 wks (percent change compared to control)

M 0% −13% −22%* −34%*

F 0% 5% −32%* −12%

Absolute spleen weight; 104 wks (percent change compared to 

control)

M 0% −11% −16% −4%

F 0% 58% 8% 37%

Relative spleen weight; 104 wks (percent change compared to control)

M 0% −11% −14% 1%
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-59 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

F 0% 77% 19% 55%

Levine et al. (1983); Thompson (1983)

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

24 mo

No immune effects were observed with routine hematology, clinical 

chemistry and histopathology evaluations.

Doses 0 0.3 1.5 8.0 40

WBC count; 105 wks (percent change compared to control)

M 0% −11% 103% d 184%d 15%

F 0% 7% 12% 354%d 251%d

Absolute spleen weight; 105 wks (percent change compared to 

control)

M 0% 5% −10% −32% −49%

F 0% −28% −44% −35% 17%

Relative spleen weight; 105 wks (percent change compared to control)

M 0% 9% 4% −29% −38%

F 0% −34% −45% −36% 9%

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 spleen weight (percent change compared to control)

M 0% − − − 18% 13%

F 0% − − − −2% −8%

Relative spleen weight (percent change compared to control)

M 0% − − − 24% 14%

F 0% − − − −3% −5%

Experiment 2: 0, 40, 60, 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)b

Diet

13 wks

Doses 0 80 160 320

WBC count (percent change compared to control)

M 0% −27% −12% 30%

F 0% −17% 3% −3%

Absolute spleen weight (percent change compared to control)

M 0% 17% 0% −17%

F 0% −22% 0% 0%

Relative spleen weight (percent change compared to control)

M 0% 25% 5% 0%

F 0% −12% 0% −3%

Cholakis et al. (1980)

Rats, F344, 10/sex/group

Doses 0 10 14 20 28 40

WBC count (percent change compared to control)

M 0% − − − −12% 7%
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-60 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

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

F 0% − − − 17% 30%

Absolute spleen weight (percent change compared to control)

M 0% − − − 2% −4%

F 0% − − − −10% −12%*

Relative spleen weight (percent change compared to control)

M 0% − − − 5% 5%

F 0% − − − −8% −8%

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

13 wks

No immune effects were observed upon routine histopathology 

evaluation.

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d 

Gavage

90 d

No effects were observed on thymus or spleen histology, red and white 

blood cell populations, or lymphocyte populations.

Doses 0 4 8 10 12 15

WBC count (percent change compared to control)

M 0% −5% −12% −7% 1% −3%

F 0% 22% 45% 12% 52% 29%

Absolute spleen weight (percent change compared to control)

M 0% −3% −6% 3% 1% 5%

F 0% 1% 8% 23%* 17%* 24%*

Relative spleen weight (percent change compared to control)

M 0% 3% 4% 7% −1% 2%

F 0% 1% 0% 6% −1% −2%

Absolute thymus weight (percent change compared to control)

M 0% −1% 3% −10% −12% −25%

F 0% −7% 12% 19% 32% 19%

Relative thymus weight (percent change compared to control)

M 0% −1% 3% −10% −12% −25%

F 0% −7% 4% 4% 12% −6%

Levine et al. (1981a);Levine et al. (1990);

Levine et al. (1981b)c

Data were not reported for rats in the 300 or 600 mg/kg dose groups 

because all of the rats died before the 13-wk necropsy.
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-61 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

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

WBC count (percent change compared to control)

M 0% 4% 7% 15% − −

F 0% 23%* 24%* 62%* − −

Absolute spleen weight (percent change compared to control)

M 0% −11% −16% −34% − −

F 0% 2% 12% 0% − −

Relative spleen weight (percent change compared to control)

M 0% −9% −12% −21% − −

F 0% 2% 12% 3% − −

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

3 mo

Doses 0 15 25 50

WBC count (percent change compared to control)

M 0% −30% 7% −6%

Hart (1974)

Dogs, Beagle, 3/sex/group

Pre-mix with ground dog chow 

containing 20 mg RDX/g-chow, 60 grams 

of dog food

0, 0.1, 1, or 10 mg/kg-d 

Diet

90 d

Doses 0 0.1 1 10

WBC count (percent change compared to control)

M 0% 5% 2% −19%

F 0% −2% 24% 6%

Absolute spleen weight (percent change compared to control)

M 0% − − 123%

F 0% − − −11%

Martin and Hart (1974)

Monkeys, Cynomolgus or Rhesus, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d 

Gavage

90 d

Doses 0 0.1 1 10

WBC count (percent change compared to control)

M 0% −32% 0% −3%

F 0% −38% −1% −41%

Gastrointestinal effects
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-62 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

Lish et al. (1984); Levine 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

24 mo

No gastrointestinal tract effects were observed as clinical signs or on 

gross pathology or histopathology examination.

Levine et al. (1983); Thompson (1983)

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

24 mo

No gastrointestinal tract effects were observed as clinical signs or on 

gross pathology or histopathology examination.

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d 

Gavage

90 d

No gastrointestinal tract effects were observed on gross pathology or 

histopathology examination. Increased salivation and blood stains 

around the mouth were noted (affected doses and incidences were not 

reported); it is not clear whether these effects occurred in animals also 

experiencing convulsions.

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

3 mo

Congestion of the gastrointestinal tract was observed in 50 and 100 

mg/kg-d rats that also exhibited mortality (40%) and severe 

neurotoxicity. 

Martin and Hart (1974)

Monkeys (Cynomolgus or Rhesus); 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d 

Gavage

90 d

Vomiting was observed more frequently in the 1 and 10 mg/kg-d

groups compared to the control or 0.1 mg/kg-d groups, although some 

episodes occurred during the intubation procedure.

Hart (1974)

Dogs, Beagle, 3/sex/group

Pre-mix with ground dog chow 

containing 20 mg RDX/g-chow, 60 grams 

of dog food

0, 0.1, 1, or 10 mg/kg-d 

Diet

90 d

Some nausea and vomiting were reported (incidences and affected 

dose groups were not reported).
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-63 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

Hematological effects

Lish et al. (1984); Levine 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

24 mo

Doses 0 1.5 7.0 35 175/100

RBC count; 105 wks (percent change compared to control)

M 0% −4% 3% −3% 14%

F 0% 4% −7% 5% 3%

Hemoglobin; 105 wks (percent change compared to control)

M 0% −6% 3% −5% 9%

F 0% 2% −7% 3% 1%

Hematocrit; 105 wks (percent change compared to control)

M 0% −4% 3% −4% 9%

F 0% 3% −6% 3% 1%

Platelets; 105 wks (percent change compared to control)

M 0% 33% 9% 21% 27%

F 0% −14% −7% 1% 5%

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

RBC count; 104 wks (percent change compared to control)

M 0% 3% 7% −2%

F 0% −14% 7% 2%

Reticulocyte count; 104 wks (percent change compared to control)

M 0% 250%c 500%*c 850%*c

F 0% 180%*c −40% 20%

Hemoglobin; 104 wks (percent change compared to control)

M 0% 3% 4% 0%

F 0% −1% 1% −2%

Levine et al. (1983); Thompson (1983)

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

24 mo

Doses 0 0.3 1.5 8.0 40

Hemoglobin levels; 105 wks (percent change compared to control)

M 0% 6% 6% 3% −13%

F 0% −5% 1% −9% −14%

RBC count; 105 wks (percent change compared to control)

M 0% 5% 2% −1% −9%

F 0% −2% 2% −9% −13%

Platelet count; 105 wks (percent change compared to control)

M 0% 6% −4% −10% −7%
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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

1-64 DRAFT—DO NOT CITE OR QUOTE

Reference and study design Results

F 0% 14% −4% 5% 22%

Hematocrit; 105 wks (percent change compared to control)

M 0% 5% 5% 2% −7%

F 0% −5% 0% −8% −12%

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)b

Diet

13 wks

Doses 0 80 160 320

RBC count (percent change compared to control)

M 0% −5% −12%* −2%

F 0% −10% −1% 1%

Reticulocytes (percent change compared to control)

M 0% −36% −13% 15%

F 0% 21% 25% −19%

Hematocrit (percent change compared to control)

M 0% −1% −6% 0%

F 0% −8% 2% 1%

Hemoglobin (percent change compared to control)

M 0% −2% −7%* −3%

F 0% −5% 4% 1%

Platelets (percent change compared to control)

M 0% 33% 28% 22%

F 0% 3% 9% 39%

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

RBC count (percent change compared to control)

M 0% − − − 3% −1%

F 0% − − − −1% −7%

Hemoglobin (percent change compared to control)

M 0% − − − 2% −1%

F 0% − − − −1% −1%

Platelet (percent change compared to control)

M 0% − − − 11% 16%*

F 0% − − − −23% −13%

Reticulocytes (percent change compared to control)

M 0% − − − 26% 76%*

F 0% − − − −2% 17%
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Reference and study design Results

Hematocrit (percent change compared to control)

M 0% − − − 3% 0%

F 0% − − − 0% −2%

Crouse et al. (2006)

Rats, F344, 10/sex/group

99.99% pure

0, 4, 8, 10, 12, or 15 mg/kg-d 

Gavage

90 d

Doses 0 4 8 10 12 15

RBC count (percent change compared to control)

M 0% 1% −7% −2% −4% −5%

F 0% 3% 3% −1% 2% −2%

Hemoglobin (percent change compared to control)

M 0% −1% −5% 0% −1% −6%

F 0% 2% 4% −1 4% −4%

Platelet count (percent change compared to control)

M 0% 21% 11% 13% −8% 34%

F 0% 6% 40% 47% 34% −36%

Hematocrit (percent change compared to control)

M 0% 2% −5% 0% −1% −4%

F 0% 3% 4% 0% 4% −2%

Levine et al. (1981a);Levine et al. (1990);

Levine et al. (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

Data were not reported for rats in the 300 or 600 mg/kg dose groups 

because all of the rats died before the 13-wk necropsy.

Doses 0 10 30 100 300 600

Hematocrit (percent change compared to control)

M 0% −2% −1% −5% − −

F 0% 0% −4% −7% − −

Hemoglobin (percent change compared to control)

M 0% −3% −1% −6% − −

F 0% 0% −4% −8%* − −

RBC count (percent change compared to control)

M 0% −2% −2% −5% − −

F 0% −1% −4% −5% − −

Reticulocytes (percent change compared to control)

M 0% −4% 10% 28% − −

F 0% 9% 73% 71% − −

Von Oettingen et al. (1949)

Rats, sex/strain not specified, 20/group

90−97% pure, with 3−10% HMX; particle 

size not specified

Doses 0 15 25 50

RBC count (percent change compared to control)

M + F 0% −23% −12% −14%
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Reference and study design Results

0, 15, 25, or 50 mg/kg-d

Diet

3 mo

Hemoglobin (percent change compared to control)

M + F 0% −25% −7% −11%

Hart (1974)

Dogs, Beagle, 3/sex/group

Pre-mix with ground dog chow 

containing 20 mg RDX/g-chow, 60 g dog 

food

0, 0.1, 1, or 10 mg/kg-d 

Diet

90 d

Doses 0 0.1 1 10

RBC count (percent change compared to control)

M 0% −3% 3% 2%

F 0% 13% 7% 11%

Reticulocyte count (percent change compared to control)

M 0% −66% 0% −50%

F 0% −17% −50% 0%

Hematocrit (percent change compared to control)

M 0% −4% 2% 0%

F 0% 6% 1% 7%

Hemoglobin (percent change compared to control)

M 0% 5% −2% 0%

F 0% 8% −2% 8%

Martin and Hart (1974)

Monkeys, Cynomolgus or Rhesus, 

3/sex/group

Purity of test material not specified

0, 0.1, 1, or 10 mg/kg-d 

Gavage

90 ds 

Histopathological examination revealed increased numbers of 

degenerate or necrotic megakaryocytes in all bone marrow sections.

Doses 0 0.1 1 10

RBC count (percent change compared to control)

M 0% −3% 2% −3%

F 0% 0% −1% 2%

Reticulocyte count (percent change compared to control)

M 0% −33% −50% −50%

F 0% −18% −36% 45%

Hematocrit (percent change compared to control)

M 0% −7% −4% −1%

F 0% 10% 7% 3%

Hemoglobin (percent change compared to control)

M 0% −10% −8% −6%

F 0% 6% 6% 3%

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*Statistically significantly different compared to the control, as determined by study authors (p < 0.05). a

Incidence counts exclude individuals from which blood was obtained via the orbital sinus.

b

Doses were calculated by the study authors.
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c 1 Levine et al. (1981a) is a laboratory report of a 13-week study of RDX in F344 rats; two subsequently published 

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Standard deviations accompanying the mean response in a given dose group were high, suggesting uncertainty in 

the accuracy of the reported percent change compared to control.
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1 Summary of Other Toxicity Data

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Effects on the eyes and the musculoskeletal, cardiovascular, immune, and gastrointestinal 

systems have been reported in some studies. EPA concluded that the evidence does not support 

these effects as a potential human hazard of RDX exposure.

1.2. INTEGRATION AND EVALUATION

1.2.1. Effects Other Than Cancer

The majority of evidence for the health effects of RDX comes from oral toxicity studies. The 

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

of exposure. Three epidemiological 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), uncertainty in identifying exposure levels, small 

sample sizes, and inadequate reporting. The single animal inhalation study identified in the 

literature search had deficiencies that precluded its inclusion in this assessment (see Literature 

Search Strategy | Study Selection and Evaluation). 

The strongest evidence for hazards following exposure to RDX is for nervous system effects. 

A human occupational study (Ma and Li, 1992) describes memory impairment and visual-spatial 

decrements, and several case reports provide additional evidence of associations between exposure 

to RDX and seizures and convulsions (Kasuske et al., 2009; Küçükardali et al., 2003; Testud et al., 

1996b; Testud et al., 1996a; Woody et al., 1986 and others, see Appendix C.3). Other nervous 

system effects identified in human case reports include dizziness, headache, confusion, and 

hyperirritability. Evidence from toxicity studies in multiple animal species involving chronic, 

subchronic and gestational exposures is consistent with the effects seen in humans. 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., 1983; Levine et al., 1981a; Levine et al., 1981b; Cholakis et al., 1980; Von 

Oettingen et al., 1949). In a number of these studies, death occurred at RDX doses that induced 

nervous system effects. Crouse et al. (2006), a study designed to more systematically record 

nervous system effects, reported that pre-term deaths occurred earlier in the higher-dose groups 

and in almost all cases, deaths were preceded by neurotoxic signs such as tremors and convulsions. 

The strength of a direct association between mortality and nervous system effects is less clear in 

most of the earlier studies because the frequency of clinical observations may have been 

insufficient to observe seizures prior to death. 

Induction of convulsions and seizures appears to be more strongly correlated with dose 

than with duration of exposure. It is unclear if nervous system effects increased in severity (e.g., 

from behavioral change to seizures and convulsions) with increasing dose because 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. Additional 
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1 support for an association between RDX exposure and nervous system effects comes from

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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). Although the MOA is 

unknown, 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., 2009). 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). EPA identified nervous system effects as a human 

hazard of RDX exposure. 

Evidence for kidney and other urogenital toxicity is more limited than evidence for

neurotoxicity. Increased relative kidney weight was observed in male and female mice (Lish et al., 

1984), and histopathological changes in the urogenital system (including suppurative prostatitis) 

were reported in male rats exposed to RDX in the diet for 2 years (Levine et al., 1983). 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. Among the lesions identified

in the rat, the incidence of suppurative prostatitis is considered a marker for RDX-related 

urogenital effects. The plausibility of a MOA that shares a common molecular initiating event 

(binding to the GABAA receptor convulsant-site) with the neurotoxic effects of RDX increases 

support for an association between RDX exposure and kidney and other urogenital effects. EPA 

identified the urogenital system as a potential human hazard of RDX exposure.

Evidence for male reproductive toxicity 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 study authors at both the 

penultimate and high dose tested in the study. However, studies in different rat strains did not 

consistently report testicular effects. Although the available data are limited, given the dose-related 

findings of mouse testicular degeneration, EPA identified suggestive evidence of male reproductive 

effects as a potential human hazard of RDX exposure. 

Evidence for developmental toxicity and liver toxicity was more limited than that for the 

endpoints discussed above. In animal studies, embryotoxicity and other developmental effects 

were observed only at doses associated with maternal mortality (Angerhofer et al., 1986; Cholakis 

et al., 1980). Evidence for hepatic effects comes from observations of increases (generally doserelated) in liver weight in some chronic and subchronic oral animal studies (Lish et al., 1984; Levine 

et al., 1983; Levine et al., 1981a; Levine et al., 1981b; Cholakis et al., 1980; Hart, 1976). However, 

these elevations in liver weight were not accompanied by RDX-related histopathological changes in 

the liver or increases in serum liver enzymes. In addition, 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. EPA concluded that evidence does not support developmental toxicity or liver 
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1 effects as potential human hazards of RDX exposure. Thus, these effects were not considered 

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further for dose-response analysis and the derivation of reference values. 

1.2.2. Carcinogenicity

Under EPA’s Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), the database for 

RDX provides “suggestive evidence of carcinogenic potential” based on the finding of statistically 

significant trends for hepatocellular adenomas or carcinomas and alveolar/bronchiolar adenomas 

or carcinomas in female, but not male, B6C3F1 mice (Lish et al., 1984). This is further supported by 

the finding of a statistically significant trend for hepatocellular carcinomas in male, but not female, 

F344 rats (Levine et al., 1983) exposed to RDX in the diet for two years. On the other hand, there 

was no evidence of carcinogenicity in Sprague-Dawley rats in a 2-year dietary study of RDX (Hart, 

1976). No human studies are available to assess the carcinogenic potential of RDX. 

EPA’s Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a) emphasizes the 

importance of weighing the totality of evidence in reaching conclusions about the human 

carcinogenic potential of agents under evaluation. Information taken into consideration in 

weighing the evidence for the human carcinogenic potential of RDX includes the magnitude of

response in rats, availability of a PWG reevaluation of tumors, and potential differences in test 

material across studies.

The incidence of male rat liver carcinomas as reported by Levine et al. (1983) showed a 

positive trend with dose (based on statistical analysis conducted for this review), thus supporting 

the positive finding of liver tumors in female mice. However, as discussed in Section 1.1.5, the 

association of liver tumors in rats with RDX exposure was judged not to be strong for several 

reasons, including the small numbers of carcinomas observed across the study and the low survival 

rate in the high-dose group that reduces confidence that the final incidence in that group accurately 

reflects lifetime cancer incidence. A PWG reevaluation of rat liver tumors has not been conducted. 

The weight of evidence of carcinogenicity also took into consideration the lack of 

carcinogenic response in the two-year bioassay in the Sprague-Dawley rat (Hart, 1976). The 

incidence of liver tumors in the Hart (1976) study was not increased relative to controls at a dose of 

10 mg/kg-day, a dose that fell in the range of doses in the Levine et al. (1983) study that showed a 

positive tumor trend. 

A cancer descriptor may be applicable to a variety of potential data sets and represent 

points along a continuum of evidence (U.S. EPA, 2005a). The available evidence for RDX suggests 

that it could be considered a borderline case between two descriptors—“likely to be carcinogenic to 

humans” and “suggestive evidence of carcinogenic potential.” One of the criteria identified in EPA’s 

Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a) that supports the likely descriptor 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” may qualify as a likely 

carcinogen. Bioassay data provide evidence for RDX carcinogenicity in one sex of one species

(female mouse), weaker evidence for carcinogenicity in one sex of a second species (male rat), and
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1 evidence of tumors in two tissues (liver and lung); this evidence could be considered to meet the 

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criteria for the “likely to be carcinogenic to humans” descriptor. 

The “suggestive evidence of carcinogenic potential” descriptor is appropriate when the 

weight of evidence is suggestive of carcinogenicity, and a concern for potential carcinogenic effects 

in humans is raised, but the data are judged not sufficient for a stronger conclusion. This descriptor 

covers a wide spectrum of evidence associated with varying levels of concern for carcinogenicity, 

including a positive cancer result in the only study on an agent, a single positive cancer result in an 

extensive database that includes negative studies in other species, or evidence of a positive 

response in a study whose power, design, or conduct limits the ability to draw a positive conclusion. 

In reviewing the carcinogenicity data for RDX, EPA considered that either descriptor is 

plausible, as the evidence for increased trends in tumor incidence in two tissues and possibly a 

second species raises a concern for carcinogenic effects in humans. However, in light of the 

determination that the association between RDX exposure and liver tumors in rats is not strong, 

and the lack of a carcinogenic response in male B6C3F1 mice, female F344 rats, and Sprague-Dawley 

rats of both sexes, EPA concluded that there is “suggestive evidence of carcinogenic potential” for 

RDX. 

U.S. EPA’s Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a) indicate that for 

tumors occurring at a site other than the initial point of contact, the weight of evidence for 

carcinogenic potential may apply to all routes of exposure that have not been adequately tested at 

sufficient doses. An exception occurs when there is convincing toxicokinetic data that absorption 

does not occur by other routes. Information available on the carcinogenic effects of RDX via the 

oral route demonstrates that tumors occur in tissues remote from the site of absorption. 

Information on the carcinogenic effects of RDX via the inhalation and dermal routes in humans or 

animals is not available. Based on the observation of systemic tumors following oral exposure, and 

in the absence of information to indicate otherwise, it is assumed that an internal dose will be 

achieved regardless of the route of exposure. Therefore, there is “suggestive evidence of 

carcinogenic potential” following exposure to RDX by all routes of exposure.

1.2.3. 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 toxicokinetic and 

toxicodynamic differences between susceptible and other populations. 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. Reproductive and developmental toxicity studies did not identify effects in offspring 

at doses below those that also caused maternal toxicity (Angerhofer et al., 1986; Cholakis et al., 

1980). 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 
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1 et al., 2007). Because RDX is neurotoxic in adult animals, evidence of gestational transfer of RDX to 

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the developing organism, along with the presence of RDX in milk, suggests that the nervous system 

may be a target in the developing organism; however, developmental neurotoxicity studies of RDX 

have not been conducted. Limited data suggest that male laboratory animals may be more 

susceptible to noncancer toxicity associated with RDX exposure. While no sex-based differences in 

neurotoxicity were observed, urogenital effects have been observed in males at lower doses than in 

females (Levine et al., 1983; Levine et al., 1981a; Levine et al., 1981b; Cholakis et al., 1980), 

suggesting a possible sex-based difference in susceptibility to RDX toxicity. 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 no 

information to support predictions of how genetic polymorphisms may affect susceptibility.
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2.DOSE-RESPONSE ANALYSIS

2.1. ORAL REFERENCE DOSE FOR EFFECTS OTHER THAN CANCER

1 The RfD (expressed in units of mg/kg-day) is defined as an estimate (with uncertainty 

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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-observedadverse-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 and Effects for Dose-Response Analysis

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 

to RDX examined the relationship between certain health endpoints and inhalation exposure; no 

epidemiological studies of ingested RDX are available. Therefore, epidemiological studies could not 

be used for oral dose-response analysis and as the basis for the RfD. 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 Section 1.2.1, based on findings from oral studies in experimental animals, 

EPA identified nervous system effects as a human hazard of RDX exposure, and effects on the 

urogenital system (including the kidney) as a potential human hazard of RDX exposure. EPA also 

identified suggestive evidence of male reproductive effects as a potential human hazard of RDX 

exposure. Experimental animal studies within each health effect category were evaluated using 

general study quality considerations discussed in Section 6 of the Preamble and in the section on 

Literature Search Strategy | Study Selection and Evaluation to help inform the selection of studies 

from which to derive oral reference values. Rationales for selecting the studies and effects to 

represent each of these hazards are summarized below.

Nervous System Effects

Nervous system effects following oral exposure to RDX, including convulsions, seizures, and 

hyper-reactivity, were observed in multiple studies in rats, mice, monkeys, and dogs. Only three 
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1 studies reported data on the incidence of nervous system findings—Crouse et al. (2006), Cholakis 

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et al. (1980), and Martin and Hart (1974). Two of these—Crouse et al. (2006) and Cholakis et al. 

(1980)—were selected for dose-response analysis.

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. Additionally, Crouse 

et al. (2006) observed that for all the dose groups where unscheduled deaths were recorded, 

mortality was strongly associated with seizures or convulsions. This study used a test material of 

high purity (99.99% RDX), six dose groups (including the control) that provided good resolution of 

the dose-response curve, and relatively low doses that still provided adequate responses. Cholakis 

et al. (1980) reported a dose-related increase in convulsions in a developmental toxicity study, with 

convulsions observed at a dose as low as 2 mg/kg-day RDX on GDs 6−19. Because evidence of 

nervous system effects was observed in this study at a relatively low dose, this study was also 

selected for dose-response analysis. 

The study in monkeys by Martin and Hart (1974) was not selected for dose-response 

analysis. This study provided supporting evidence of nervous system effects (trembling, shaking, 

ataxia, and hyperactive reflexes) with 66% incidence at the high dose of 10 mg/kg-day; however, 

this study was not selected for dose-response analysis because it used small group sizes (n = 3/sex) 

and the exposures were relatively variable or uncertain (e.g., purity of the test material was not 

specified, and reported emesis in some animals likely influenced the amount of dose received). 

Other chronic and subchronic studies reported nervous system effects as clinical 

observations (Angerhofer et al., 1986; Lish et al., 1984; Levine et al., 1983; Levine et al., 1981a; 

Levine et al., 1981b; Von Oettingen et al., 1949), but without incidence data. As discussed in Section 

1.1.1, these studies did not systematically monitor or evaluate nervous system effects induced by 

RDX, leading to possible underestimates of incidence of such effects. As such, there is some 

uncertainty associated with identification of NOAELs and LOAELs for nervous system effects from 

these studies. Further, these studies reported convulsions and other indications of nervous system 

effects at doses higher than the doses at which effects were observed in Cholakis et al. (1980), i.e., 

≥2 mg/kg-day, and Crouse et al. (2006), i.e., ≥8 mg/kg-day. 

Kidney and Other Urogenital Effects

Effects on kidney and other urogenital system endpoints included changes in kidney weight 

and histopathological findings in the kidney, bladder, and prostate in experimental animals exposed 

orally to RDX. As discussed in Section 1.1.3, kidney weight changes across experimental animal 

studies were not consistent and were difficult to interpret; therefore kidney weight data sets were 

not selected for quantitative analysis.

Histopathological changes in the urogenital system were reported in a 2-year study in F344 

rats by Levine et al. (1983) and in a 13-week study in B6C3F1 mice by Cholakis et al. (1980). 

Histopathological changes of the kidney and bladder (medullary papillary necrosis, suppurative 

pyelitis, uremic mineralization, and luminal distention and cystitis of the urinary bladder) were 
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1 observed by Levine et al. (1983) in high-dose (40 mg/kg-day) males. The incidence of suppurative 

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prostatitis, considered to be a marker for the broader range of urogenital effects in these animals, 

showed a dose-related trend beginning at doses below 40 mg/kg-day (see Section 1.1.3). 

Therefore, suppurative prostatitis was selected for dose-response modeling as a sensitive measure 

of RDX effects on the urogenital system. 

Cholakis et al. (1980) examined the kidney for histopathological changes in control and 

high-dose (320 mg/kg-day) mice only. Because incidence data from only a single high-dose group 

was available, this study was not selected for dose-response analysis.

Male Reproductive Toxicity

Male reproductive effects were identified in mice following chronic administration of RDX 

in the diet. Lish et al. (1984) observed an increased incidence of testicular degeneration in mice 

given RDX in diet for two years compared to controls. The response was shown to be dose-related 

and was selected for dose-response modeling. Changes in other reproductive outcomes were not 

dose-related or consistently observed across studies, and therefore were not considered for doseresponse modeling. 

2.1.2. Methods of Analysis

Benchmark dose (BMD) modeling and physiologically-based pharmacokinetic (PBPK) 

models were used in this assessment to estimate candidate points of departure (PODs) for the 

derivation of an RfD for RDX. The general approach for the estimation of PODs is presented in 

Figure 2-1 and described further below.

Human 

external dose

(HED)

Benchmark 

dose modeling Animal PBPK model Human PBPK model

Animal 

external dose

BMDL or 

NOAEL

Animal 

internal dose

Figure 2-1. Approach for dose-response analysis. 

No biologically based dose-response models are available for RDX. In this situation, 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 95% lower confidence limit on the BMD (BMDL) using a 

benchmark response (BMR) selected for each effect. A summary of BMD modeling, including 
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1 selection of BMRs, for each of the health effect categories is provided below. 

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Nervous System Effects

Incidence data from Crouse et al. (2006) and Cholakis et al. (1980) were amenable to

modeling. For Crouse et al. (2006), statistical analysis (Cochran-Mantel-Haenszel test) conducted 

by EPA indicated no significant difference in convulsion rates of male and female rats; thus, 

combined incidence data from male and female rats were used for modeling convulsion data from 

this study. A BMR of 1% extra risk for convulsions was used to address the relative severity of this 

endpoint; across the experimental animal database for RDX, convulsions and seizures were 

generally associated with mortality. In general, severe endpoints are not 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. Less severe nervous system outcomes that precede 

convulsions and associated mortality would be preferred, but none were identified for RDX. 

Kidney/Urogenital and Male Reproductive Effects

Incidence data on prostate effects as reported by Levine et al. (1983) and testicular 

degeneration as reported by Lish et al. (1984) were amenable to modeling. Cut-offs for the 

biological significance of these effects were not identified, and a BMR of 10% was applied under the 

assumption that it represents a minimally biologically significant degree of effect. Uncertainty in 

this characterization should be taken into account in comparisons with PODs from other effects.

Human Extrapolation

EPA guidance (U.S. EPA, 2011) advocates a hierarchy of approaches for deriving human 

equivalent doses (HEDs) from data in laboratory animals, with the preferred approach being 

physiologically-based toxicokinetic modeling. Other approaches can include using chemicalspecific information in the absence of a complete physiologically-based toxicokinetic 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 for the purpose of deriving an oral RfD.

As described below, HEDs for candidate PODs for RDX were derived using PBPK models for 

endpoints selected from rat and mouse bioassays, and are compared in Table 2-1 to estimates 

derived from administered RDX dose.
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Table 2-1. Summary of derivation of PODs following oral exposure to RDX

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

Nervous system 

Convulsions

Crouse et al. (2006)

(90-d/gavage)

Male and 

female F344 

rat, 

combined

Multistage 

3°

1% ER 1.53 0.54 0.13 0.27

Convulsions 

Cholakis et al. 

(1980)

(GDs 6−19/gavage)

Female F344 

rat

Quantallinear1% ER 0.18 0.12 0.03 0.06

Kidney/urogenital system

Prostate 

suppurative 

inflammation

Levine et al. (1983)

(2-yr/diet)

Male F344 

rat

LogProbit 10% ER 1.67 0.47 0.11 0.23

Male reproductive system

Testicular 

degeneration

Lish et al. (1984)

(2-yr/diet)

Male B6C3F1

mouse

LogProbit 10% ER 56.0 16.3 2.4 0.08

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For modeling details, see Appendix D.

b

POD was convered 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 area under the curve 

[AUC] for RDX concentration in arterial blood) derived using PBPK models.

ER = extra risk

Physiologically-based pharmacokinetic 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.2.5). As concluded in the MOA analyses for the 

various observed noncancer effects associated with RDX exposure, the available data are 

insufficient to establish any specific mode(s) of action for these effects, and there appears to be no 

clear evidence linking health effects with RDX-generated metabolites. 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). 
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1 Therefore, PBPK model-derived arterial blood concentration of RDX is considered a better dose2 

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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 the area under 

the curve (AUC) for RDX concentration in arterial blood, which represents the average blood RDX 

concentration for the exposure duration normalized to 24 hours. 

It appears logical to use RDX concentration levels in the brain as the internal dose metric for 

analyzing convulsions as the health effect. 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. It may also be noted that there is greater confidence in model estimates of 

blood AUC versus peak blood concentrations because, as discussed in Appendix C, Section C.2.5, the 

rate constant for oral absorption (KAS) is uncertain, and peak concentrations are more sensitive to 

variations in this parameter than average values. Furthermore, a more consistent dose-response 

for convulsions is observed in chronic studies than for the higher exposures in subchronic studies.

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 dose. The human PBPK model was then 

applied to derive the corresponding HEDs (see Figure 2-1). 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 (i.e., 

the sequence of this calculation is immaterial for the RDX data).

HEDs were also calculated consistent with EPA guidance (U.S. EPA, 2011) using PODs 

(BMDLs or NOAELs) determined from administered RDX doses and employing 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 a BWa 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 2-1): 

PODHED = Laboratory animal dose (mg/kg-day) × DAF 
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1 Further details of the BMDL modeling, BMDS outputs, and graphical results for the best fit 

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model for each dataset included in Table 2-1 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.2.5. Table 2-1 summarizes the results of the BMD modeling and the PODHED for each data 

set discussed above.

2.1.3. Derivation of Candidate Values

Pharmacokinetic models are useful to examine species differences in pharmacokinetic 

processing. Because of relatively high confidence in the rat and human PBPK modeling, these 

models were used to derive reliable internal dose metrics for extrapolation. For datasets selected 

from the rat bioassays, the candidate RfDs were calculated assuming cross-species toxicological 

equivalence of the AUC of RDX blood concentration derived from the PBPK modeling. However, 

there were major uncertainties identified in the mouse PBPK modeling. Therefore, for endpoints

selected from the mouse bioassay, the preferred approach for determining the candidate RfDs is 

that based on the administered dose of RDX extrapolated to humans using allometric BW3/4 scaling. 

The evaluation of confidence in the PBPK model results is summarized in Summary of confidence in 

PBPK models for RDX in Appendix C, Section C.2.5. 

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. 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. 

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, a 

UFA of 3 was applied because BW3/4 scaling is used to extrapolate oral doses from laboratory 

animals to humans. Although BW3/4 scaling addresses some aspects of cross-species extrapolation 

of toxicokinetic and toxicodynamic processes, some residual uncertainty 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 administered doses in humans. 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, differs depending on the exposure duration. 

An UFs of 1 was applied to the POD values for kidney/urogenital effects and testicular degeneration 

derived from the 2-year bioassays in the rat (Levine et al., 1983) and mouse (Lish et al., 1984). POD
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1 values for nervous system effects were derived from studies of subchronic duration or gestational 

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exposure; a UFS of 3 was applied to these PODs. Typically, a UFS of 10 is applied to extrapolate 

results from a subchronic duration study in the absence of a chronic study based on the assumption 

that effects from a given compound would occur at approximately a 10-fold higher exposure level in 

a subchronic study than in a chronic study, if a chronic study were available (U.S. EPA, 2002). 

However, the available nervous system effects data for RDX support an UFS of less than 10. As 

discussed in Section 1.1.1, seizure induction appears to be more strongly correlated with dose level 

than with duration of exposure. In addition, the available empirical evidence from rodent bioassays

provide support for an UFS no greater than 3. Dose levels associated with convulsions in chronic 

dietary studies of RDX are ≥35 mg/kg-day and are higher than doses that induced convulsions in 

the 14- and 90-day (gavage) studies that were used to derive candidate PODs for nervous system 

effects (i.e., 2 mg/kg-day in Cholakis et al. (1980) and 8 mg/kg-day in Crouse et al. (2006)) (also see 

Table 1-2 and Figure 1-1). Thus, the available RDX data for nervous system effects is consistent 

with the application of a UFS that is less than the default of 10. 

A LOAEL to NOAEL uncertainty factor, UFL, of 1 was applied to all POD values because the 

PDO was a BMDL. When the POD is a BMDL, the current approach is to address this factor as one of 

the considerations in selecting a BMR for benchmark dose 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. Deficiencies 

in the database related to neurobehavioral and neurodevelopmental testing were identified. The 

database for neurotoxicity is characterized primarily by observations of frank effects (convulsions). 

Additional observations of neurobehavioral effects were reported (Levine et al., 1990; Angerhofer 

et al., 1986; Levine et al., 1983; Levine et al., 1981a; Levine et al., 1981b; Cholakis et al., 1980; Von 

Oettingen et al., 1949); however, a FOB conducted by Crouse et al. (2006) did not report any

consistent, treatment-related behavioral effects. Further, Crouse et al. (2006) noted that the ability 

of the FOB to identify neurobehavioral effects at doses ≥8 mg/kg-day was limited due to the timing 

of the dosing procedure and timing of the FOB screenings. Given the reports of neurobehavioral 

effects in several studies, additional systematic evaluation of neurobehavioral effects would be 

informative. 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. Given the potential for exposure during gestation and lactation and the 

neurotoxic potential of RDX, the lack of a developmental neurotoxicity study was identified as a 

data gap. A UFD of 3 was applied to all PODs to account for limitations in neurobehavioral and 

neurodevelopmental testing. 
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1 Table 2-2 is a continuation of Table 2-1 and summarizes the application of UFs to each 

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PODHED to derive a candidate value for each data set. The candidate values presented in the table 

below 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.

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

Endpoint and reference PODHEDa

POD 

type UFA UFH UFL UFS UFD

Composite 

UF

Candidate

value 

(mg/kg-d)

Nervous system (rats)

Convulsions

Crouse et al. (2006)

0.27 BMDL01 3 10 1 3 3 300 8.8 × 10−4

Convulsions 

Cholakis et al. (1980)

0.06 BMDL01 3 10 1 3 3 300 2.0 × 10−4

Kidney/urogenital system (rats)

Prostate suppurative 

inflammation

Levine et al. (1983)

0.23 BMDL10 3 10 1 1 3 100 2.3 × 10−3

Male reproductive system (mice)

Testicular degeneration

Lish et al. (1984)

2.4 BMDL10 3 10 1 1 3 100 2.5 × 10−2

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PODHED values based on data from the rat were derived using PBPK modeling; the HED POD based on data from 

the mouse was derived using BW3/4 adjustment (see Section 2.1.3 and discussion of the PBPK models in 

Appendix C, Section C.2.5).

Figure 2-2 presents graphically the candidate values, UFs, and PODHEDs, with each bar 

corresponding to one data set described in Tables 2-1 and 2-2. 
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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

1 Table 2-3 distills the candidate values from Table 2-2 into a single value for each organ or 

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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 2-3. Organ/system-specific RfDs and proposed overall RfD for RDX

Effect Basis RfD (mg/kg-day)

Study exposure 

description Confidence

Nervous system Convulsions 9 × 10−4 Subchronic Medium

Kidney/urogenital

system

Suppurative prostatitis 2 × 10−3 Chronic Low

Male reproductive

system

Testicular degeneration 2 × 10−2 Chronic Low

Proposed overall RfD Nervous system 9 × 10−4 Subchronic Medium

Nervous System Effects

The organ/system-specific RfD for nervous system effects was based on the incidence of 

convulsions in 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 other studies. Although the candidate value derived from Cholakis et al. (1980) is lower 

(by approximately fourfold), there is greater certainty in the value derived from Crouse et al. 

(2006) because of the longer exposure duration (90 versus 14 days), more systematic evaluation of 

neurobehavioral endpoints, and higher test compound purity. 

Kidney/Urogenital Effects

A single data set for incidence of suppurative prostatitis in male B6C3F1 mice as reported by

Lish et al. (1984) was brought forward for quantitative analysis as a sensitive marker for the 

broader array of RDX-associated effects observed in the urogenital system. As previously 

discussed, the data supporting RDX-related kidney and other urogenital effects are largely limited 

to this 2-year study in the mouse. Accordingly, the candidate value for kidney and other urogenital 

effects is based on the incidence of suppurative prostatitis in male mice (Lish et al., 1984). 

Male Reproductive Effects

A single dataset for male reproductive effects was brought forward for quantitative 

analysis: the incidence of testicular degeneration as reported in male B6C3F1 mice exposed to RDX 
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1 in diet for 24 months (Lish et al., 1984). The candidate value for male reproductive effects is based

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on this dataset. 

2.1.5. Selection of the Proposed 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 

(Levine et al., 1983), and evidence for male reproductive effects comes primarily from a single 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 9 × 10−4 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 should provide 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. Uncertainties in the Derivation of Reference Dose

The following discussion identifies uncertainties associated with the RfD for RDX. To derive 

the RfD, the UF approach (U.S. EPA, 2000a, 1994) was applied to a PODHED based on nervous system 

effects in rats exposed to RDX for a subchronic duration. UFs were applied to the PODHEDs to 

account for uncertainties in extrapolating from an animal bioassay to human exposure, the likely 

existence of a diverse population of varying susceptibilities, subchronic to chronic duration, and 

database deficiencies. These extrapolations are carried out with default approaches given the lack 

of data to inform individual steps.

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 
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1 purity, formulation, and particle size. There is evidence that differences in test material 

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formulation and particle size can affect absorption of RDX. 

Nervous system effects have been documented in multiple studies and animal species and 

strains; however, there is some uncertainty 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.

2.1.7. 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. 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 is reduced largely because of limited 

examination of the potential for RDX to induce neurobehavioral and neurodevelopmental effects

and the incomplete understanding of a MOA for convulsions. Reflecting high confidence in the 

principal study and medium confidence in the database, overall confidence in the RfD is medium. 

2.1.8. Previous IRIS Assessment

The previous RfD for RDX, posted to the IRIS database in 1993, was based on a two-year rat 

feeding study by Levine et al. (1983). The no observed effect level (NOEL) of 0.3 mg/kg-day 

(LOAEL = 1.5 mg/kg-day) based on suppurative prostate inflammation 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 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. 
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basis for a reference value when quantitative measures of exposure are reported and the reported 

effects are determined to be associated with exposure. Of the available human epidemiological 

studies of RDX (West and Stafford, 1997; Ma and Li, 1992; Hathaway and Buck, 1977), none 

provided data that could be used for dose-response analysis. The studies by Ma and Li (1992) 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, and therefore does not provide a basis for derivation of 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 the Literature Search Strategy | Study Selection and Evaluation, a single 

experimental animal study involving inhalation exposure was identified in the DTIC database; the 

study is not publicly available. However, the study would not have provided useful data on 

responses to inhaled RDX, as the study was limited by small numbers of animals tested, a 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. Further, a PBPK model for inhaled 

RDX is not available to support route-to-route extrapolation from the RfD. 

2.2.1. Previous IRIS Assessment

An RfC for RDX was not derived in the previous assessment posted to the IRIS database in 

1990. 

2.3. ORAL SLOPE FACTOR 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 oral slope factor is a plausible upper bound on the estimate 

of risk per mg/kg-day of oral exposure.

2.3.1. Analysis of Carcinogenicity Data

As noted in Section 1.2.2, EPA concluded that there is “suggestive evidence of carcinogenic 

potential” for RDX. The Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a) state: 
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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. 

1 In the case of RDX, the carcinogenicity of the chemical has been evaluated in one oral 

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chronic/carcinogenicity bioassay in mice (Lish et al., 1984) and two bioassays in rats (Levine et al., 

1983; Hart, 1976). The data in Lish et al. (1984) demonstrated a statistically significant positive 

trend with dose6 in the incidence of liver and lung tumors in female, but not male, B6C3F1 mice 

associated with dietary administration of RDX. In the study by Levine et al. (1983), the incidence of 

liver tumors in male F344 rats showed a statistically significant positive trend with dose7. No 

increases in tumors were observed in Sprague-Dawley rats exposed to RDX (Hart, 1976). As 

discussed further below, the 2-year studies by Lish et al. (1984) and Levine et al. (1983) were wellconducted studies that support quantitative analysis. Considering these data along with the

uncertainty associated with the suggestive nature of the weight of evidence, EPA concluded that 

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) was 

performed in accordance with FDA Good Laboratory Practice regulations (FDA, 1979), included 

comprehensive histopathological examination of major organs, contained four dose groups and a 

control, used adequate numbers of animals per dose group (65/sex/group, plus interim sacrifice 

groups of 10/sex/group at 6 and 12 months) and a sufficient overall exposure duration (2 years), 

and 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 incidences 

from the PWG reanalysis of Lish et al. (1984) were used for quantitative dose-response 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 Table 2-4. 

 

6A two-sided asymptotic Cochran-Armitage test yielded p = 0.041 for liver tumors and p = 0.019 for lung 

tumors in female mice. 

7A two-sided exact Cochran-Armitage test yielded p = 0.032 for liver tumors in rat. An exact test was done 

because the incidence of tumors was too low for the asymptotic test to be reliable. 
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Table 2-4. Incidence of hepatocellular and alveolar/bronchiolar tumors in 

female B6C3F1 mice administered RDX for 2 years in diet

Tumor type Study/Analysis

Dose group

(mg/kg-day)

Control 1.5 7 35 107a

Hepatocellular adenomas or 

carcinomas

Parker et al. (2006) 1/67 4/62 5/63 10/64 4/31b

Alveolar/bronchiolar adenomas

or carcinomas

Lish et al. (1984) 7/65 3/62 8/64 12/64 7/31b

1 

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a

TWA dose, due to reductions in the highest dose from 175 to 100 mg/kg-day at week 11. b

Histopathology results are based on animals that survived more than 12 month. The smaller number of mice in 

the high-dose group reflects the high mortality at a dose of 175 mg/kg-day.

The incidence of liver carcinomas in male F344 rats from the study by Levine et al. (1983)

was also considered for quantitative dose-response analysis. Although the study was well 

conducted (see Section 1.1.5), EPA considered that the association between RDX exposure and rat 

liver tumors is not strong, reflecting the relatively low magnitude of the rat liver carcinoma 

response and reduced confidence that the high-dose group accurately reflects lifetime cancer 

incidence because, in part, of low survival. A candidate slope factor is provided in Appendix D, 

Section D.2. for comparison.

2.3.2. Dose-Response Analysis—Adjustments and Extrapolations 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 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. 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 similar across dose groups after the dose was reduced in 

the high dose group to 100 mg/kg-day; 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% extra risk was applied to both tumor sites in the mouse. 
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1 Given the finding of an association between RDX exposure in the female mouse and 

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increased tumor incidence at two tumor sites, basing the oral slope factor on only one tumor site 

could potentially underestimate the carcinogenic potential of RDX. Therefore, an analysis that 

combines the results from the mouse liver and lung tumor incidence is preferred. The MS-COMBO 

procedure (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 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, 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 1.1.5) 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 summarized briefly in Section 

2.1.3 and in more detail in Appendix C, Section C.2.5. 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 is 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). It was not necessary to adjust the administered doses to HEDs prior to BMD modeling 

because the relationship between the two dose metrics is linear and the same POD would be 

produced whether the adjustment was performed before or after modeling. Details of the BMD 

modeling can be found in Appendix D, Section D.2. 
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2.3.3. Derivation of the Oral Slope Factor

1 The lifetime oral cancer slope factor for humans is defined as the slope of the line from the 

2 

3 

4 

5 

6 

7 

BMR (10% extra risk) at the BMDL to the estimated control response at zero (slope factor = 

0.1/BMDL10-HED). This slope, a 95% upper confidence limit (UCL) on the true slope, represents a 

plausible upper bound on the true risk. The PODs estimated for each mouse tumor site are 

summarized in Table 2-5. Using linear extrapolation from the BMDL10-HED, human equivalent oral 

slope factors (OSFs) were derived for each tumor site individually and both sites combined and are 

listed in Table 2-5. 

Table 2-5. Model predictions and oral slope factors for hepatocellular and

alveolar/bronchiolar adenomas or carcinomas in female B6C3F1 mice 

administered RDX in the diet for 2 years (Lish et al., 1984a) 

Tumor type

Selected 

model BMR

BMD

(mg/kg-d)

BMDL

(mg/kg-d)

POD =

BMDL10-HEDa

(mg/kg-d)

OSFb

(mg/kg-d)−1

Hepatocellular adenomas or 

carcinomasc

Multistage 1° 10% ER 64.2 32.6 4.89 0.020

Alveolar/bronchiolar 

adenomas or carcinomas

Multistage 1° 10% ER 52.8 27.7 4.16 0.024

Liver + lung tumors Multistage 1°

(MS-COMBO)

10% ER 29.0 17.7 2.66 0.038

8 

9 

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a

BMDL10-HED = BMDL10 × (BWa

1/4/BWh1/4), where BWa = 0.035 kg, and BWh = 70 kg. b

Slope factor = BMR/BMDL10-HED, where BMR = 0.1 (10% extra risk). c

Incidence of female mouse liver tumors from Lish et al. (1984) are those reported in the PWG reevaluation (Parker 

et al., 2006). 

An OSF was derived from the BMDL10-HED based on significantly increased 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% extra risk) by the Liver + Lung BMDL10-HED and represents an upper bound 

on cancer risk associated with a continuous lifetime exposure: 

OSF = 0.1 ÷ (Liver + Lung) BMDL10-HED

= 3.8 × 10−2 (mg/kg-day)−1 

= 4 × 10−2 (mg/kg-day)−1, rounded to one significant figure
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2.3.4. Uncertainties in the Derivation of the Oral Slope Factor 

1 A number of uncertainties underlie the cancer unit risk for RDX. Table 2-6 summarizes the 

2 

3 

4 

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 2-6. 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., 1983) 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/dose group except 

highest dose where n = 31). Tumor data from 

the mouse provided a stronger basis for 

estimating the OSF than rat data, and yielded a 

higher (and therefore more health protective) 

estimate of risk than data from the rat bioassay.

Species/gender

Use of data sets from the male 

mouse would not support 

quantitative analysis of carcinogenic 

risk

OSF based on tumors in 

female mouse 

It is assumed that a positive tumor response in 

animal cancer studies indicates the agent can 

have carcinogenic potential in humans in the 

absence of data indicating 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 gender 

would be most relevant for extrapolating to 

humans, tumor data from the most sensitive 

species and gender were selected as the basis of 

the OSF.

Combined tumor types

Human risk would ↓ if OSF based on 

analysis using only a single tumor 

type

OSF based on liver and 

lung tumors in female 

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: use 

administered dose 

 

Lack of sufficient data on RDX metabolism and 

major uncertainties identified in the mouse 

PBPK model. 

 
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Consideration and 

impact on cancer risk value Decision Justification

Cross-species scaling

Alternatives could ↓ or ↑ slope 

factor (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 area under 

the curve, 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 expected neither to over- or 

underestimate human equivalent risks. 

BMD model uncertainty 

Alternative models could ↓ or ↑ 

slope factor

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 calculating plausible 

upper bound slope 

factor)

Lower bound is 95% CI on administered 

exposure at 10% extra risk of liver and lung

tumors.

Sensitive subpopulations 

↑ OSF to an unknown extent

Considered qualitatively No data are available to support a range of 

human variability/sensitivity in toxicokinetics or 

toxicodynamics for RDX, including whether 

children are more sensitive than other life 

stages.

1 

2 

3 

4 

5 

6 

7 

8 

2.3.5. Previous IRIS Assessment: Oral Slope Factor 

The previous cancer assessment for RDX was posted to the IRIS database in 1990. The oral 

slope factor 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. A slope factor of 

1.1 × 10−1 (mg/kg-day)−1 was derived using a linearized multistage procedure (extra risk). This 

differs from the slope factor for hepatocellular tumors in Table 2-6, because the current OSF is 

based on the combined incidence of hepatocellular and alveolar/bronchiolar adenomas or 

carcinomas, PWG reevaluation of female mouse liver tumors, and use of scaling by body weight to 
Toxicological Review of Hexahydro-1,3,5-trinitro-1,3,5-triazine

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1 the 3/4 power for cross-species extrapolation (whereas the previous assessment scaled by body 

2 

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4 

5 

6 

7 

8 

9 

10

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weight to the 2/3 power).

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 is a plausible upper bound on the 

estimate of risk per μg/m3 air breathed.

An inhalation unit risk value was not calculated because inhalation carcinogenicity data for 

RDX are not available. A PBPK model for inhaled RDX is not available to support route-to-route 

extrapolation 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 applied to account for early-life exposure to carcinogens that act 

through a mutagenic MOA. Because no chemical-specific data on life-stage susceptibility for RDX 

carcinogenicity are available, and because the MOA for RDX carcinogenicity is not known (see 

Section 1.1.5), ADAFs were not applied.
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