## DRAFT DELIBERATIVE: FOR INTERAGENCY REVIEW ONLY

. DO NOT DISTRIBUTE OUTSIDE YOUR AGENCY

. **EPA/635/R**

**- 14/302**  **I nteragency Review Draft**  **www.epa.gov/iris**



## T **oxicological Review of Hexahydro**

**- 1,3,5** 

**- trinitro** 

**- 1,3,5** 

**- triazine (RDX)** 

**( CASRN 121** 

**- 82** 

**- 4)** 

**In Support of Summary Informati**  **on on the**  **Integrated Risk Information System (IRIS)**

S eptember 2014

**NOTICE**

This document is an **Intera gency 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, DCToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **DISCLAIMER**

This document is a preliminary draft for review purposes only

. T his information is distributed solely for the purpose of pre

- dissemination peer review under applicable information quality guidelines. I t has not been formally disseminated by EPA

. I t does not represent and should not be construed to represent any Agency determination or policy

. M ention of trade names or commercial products does not constitute endorsement or recommendation for use.Toxicological Review of Hexahydro

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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

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............................................................................................. 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

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# TABLES

Table ES

- 1. Organ/system

- specific RfDs and proposed overall RfD for RDX

................................ .......... ES

- 2 Table ES

- 2. Summa ry 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 limitation s

................................ ................................ .... 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 u rinary 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 B6C3F 1 mice ad ministered RDX for 2 years in diet

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

- 16 Table 2

- 5. Model predictions and oral slope factors for hepatocellular and alveolar/bronc hiolar adenomas or carcinomas in female B6C3F 1 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 sy stem 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

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- 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 v alues with cor responding POD and composite UF

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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 Tech nical 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 concen tration 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 an d 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 ref erence 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 confidenc e limit UCM Unregulated Contaminant Monitoring UF uncertainty factor UF A animal

- to

- human uncertainty factor UF D database deficiencies uncertainty factor UF H human variation uncertainty factor UF L LOAEL

- to

- NOAEL uncertain factor UF S subchronic

- to

- c hronic uncertainty factor WBC white blood cells WOS Web of ScienceToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **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, DCToxicological Review of Hexahydro

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**Review ers** 
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 Pollu tion 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



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# PREFACE

This Toxicological Review critically reviews the publicly available studies on hexahydro



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- triazine (RDX) in order to identify its adverse health effects and to characterize

exposure

- response relationships

. I t was prepared under the auspices of EPA’s Integrated Risk Information System (IRIS) program

. T his 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 carcinoge nicity of RDX, as well as derivation of an oral slope factor to quantify the cancer risk associated with RDX exposure (posted in 1990). N ew information has become available and this assessment reviews information on all health effects by all exposure rout es

. O rgan/system

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

. T hese reference values may be useful for cumulative risk assessments that consider the combined effect of multiple agents acting on the sam e biological system. This assessment was conducted in accordance with EPA guidance, which is cited and summarized in the Preamble to IRIS Toxicological Reviews. T he findings of this assessment and related documents produced during its development are availab le on the IRIS web site ( h ttp://www.epa.gov/iris )

. A ppendices 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 (S ee 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 strategie s, evidence tables, and

exposure

- r esponse 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 doc k et at http://www.regulations.gov (Docket ID No. EPA

HQ

- ORD

- 2013

- 0430). In April 2011, the National Research Council (NRC) released its Review of the Environmental Protection Agency’s Draft IRIS Assessment of Form a ldehyde . 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 co nsistent with the Panel’s “Roadmap for Revision,” which viewed the full implementation of their recommendations by the IRIS Program as a multi

- y ear process. In response to the NRC’s 2011 recommendations, the IRIS Program has made changes to
streamline t he 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 r eviewing the IRISToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE as sessment development process and found that EPA has made substantial improvements to the 

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38 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 identific at ion, 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 e x perimental 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

- r esponse arrays. The hazard identification and dose

- r esponse sections include subsections based on organ/system

specific effects in which the evidence is synthesized within and integrated across all evidence for each target organ/systems.
In the draft R D X 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 yiel ding 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 o f 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 s tudy 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 t he 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 t h is 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

- r esponse relationship; specificity of the response to chemical exposure and wh et her 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

- ev idence approach to identify the potential human hazards associated with chemical

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— DO NOT CITE OR QUOTE Th e IRIS RDX assessment provides a streamlined presentation of information, integrated 

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35 hazard identification of all toxic effects, and derivation of organ/system

- specific reference va lues. 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 discussio n 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

- 5 66

- 1676 (phone), 202

- 56 6- 1749 (fax), or hotline.iris@epa.gov

.
**Chemical Properties**

R DX is a white, crystalline solid member of the nitramine class of organic nitrate explosives ( B oileau et al., 2005 ; Bingham et al., 2001 ) . I t is a synthetic chemical not found naturally in the
environment . T he 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 atm

m 3 /mole at 25°C, suggesting slow volatilization from water or moist soil ( ATSDR, 2012 ) . T he normalized soil organic carbon/water partition coefficient ( K OC ) 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 depo sit ion . R DX degrade s 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 ) . F urther 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

. I n the United States, RDX is produced at Army ammunition plants (AAP) and is not produced commercially

. RD X production peaked in the 1960 s; 180 million pounds per year were produced from 1969 to 1971. Y early total production dropped to 16 million pounds in 1984 ( A TSDR, 2012 )

. A ccording to the U.S. EPA Inventory Update Reporting program, the aggregated national production volume in 2006 was between 1 and
10 m illion pounds. RDX releases have been reported into the air, water, or soil ( ATSDR, 2012
, 1999
, 1993
, 1992 ) . R DX 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)

. De tectable levels of RDX have been observed in plants irrigated or grown wit h RDX

- contaminated water ( Best et al., 1999b ; Simini and Checkai, 1996 ; Harvey et al.,Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE 1991 ) . R DX has also been detected in indoor air samples from military facilities where RDX is 

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26 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

. O ral, inhalation, and
dermal rou t es of exposure may be relevant.
RDX h as been detected in surface water, groundwater, sediment, or soil at 34 current U.S. EPA National Priorities List (NPL) sites . Th e NPL serves as a list of sites with known releases or
threatened releases of hazardous su bstances, pollutants, or contaminants throughout the United
States and its territories

. T he NPL list aids the Agency in identifying the most serious sites that may warrant cleanup. Th e majority of the NPL sites where RDX was listed are associated with military
facilities , although the total number of sites where RDX is p resen t 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 . C ontaminants included in the UCM program are suspected of bei ng 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 . Th e 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 I**  **nternational Health Agencies** Toxicity values for RDX ha ve been e stablishe d by the Agency for Toxic Substances and Disease Registry (ATSDR), the American Conference of Governmental Industrial Hygienists (ACGIH), the Australian National Industrial Chemical s Notification and Assessment Scheme (NICNAS), the National Institute for Occupational Safety and Health (NIOSH), and th e Occupational Safety and Health A dministration (OSHA)

. T he se toxicity values and their basis are presented in Appendix A

. I t is impor tant to recognize that the assessments performed by other health agencies may have been prepared for differ ent 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

Soon after the EPA was established in 

### 1 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 39 40 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

. T he 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

- r esponse relationships

. I n terms set forth by the Nati o nal Research Council ( NRC, 1983 ) , IRIS assessments cover the hazard identification and dose

- r esponse assessment steps of risk assessment, not the exposure assessment or risk characterization steps that are conducted by th e EPA’s program and regional offic es and by other federal, state, and local health agencies that evaluate risk in specific populations and exposure scenarios. I RIS assessments are distinct from and do not address political, economic, and technical considerations that influence the design a nd 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

. T hese agents may be found in air, water, soil, or sediment. E xceptions are chemicals currently used exclusively as pesticides, ionizing and non

- i onizing radiation, and criter i a air pollutants listed under S e ction 108 of the Clean Air Act (carbon monoxide, lead, nitrogen oxides, ozone, particulate matter, and sulfur oxides). Peri o dically, 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 . A g ents may be considered for reassessment as si gnificant new studies are published

. S election is based on program and regional office priorities and on availability of adequate information to evaluate the potential for adverse effects. Ot her agents may also be

assess e d in response to an urgent public health need.



## 2. Process for developing and peer

reviewing IRIS assessments
The process for developing IRIS assessments (revised in May 2009 a nd enhanced in July 2013 ) involves critical analysis of the pertinent studies, opportunities for public input, and m u ltiple levels of scientific review . T h e 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

. T hen a public m eeting 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** 

. T he draft assessment considers all pertinent publ i cly available studies and appliesToxicological Review of Hexahydro

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23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 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**

. T he draft assessment is revised to address the comments from within the E PA. **Step 3**  **. Interagency science consultation with other federal agencies and the**  **Executive Offices of the President**

**.** T he draft assessment is revised to address the interagency comments

. T he science consultation draft, interagen cy comments, and th e EPA’s response to major comments become part of the public record. **Step 4. Public review and comment,**  **followed by external peer review**

. T he EPA releases the draft assessment for public review and comment

. A public meeting provides an o pportunity to discuss the assessment prior to peer review

. T hen the EPA releases a draft for external peer review. Th e peer review meeting is open to the public and includes time for oral public comments

. T he 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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64 65 66 67 68 69 70 71 72 73 74 75 76 77 78 T he peer review draft, written public comments, and peer review report become part of the public record. **Step 5. Revision of draft Toxicol**  **ogical Review and development of draft IRIS**  **summary**

. T h e 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

. T he 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**  T he draft assessment and summary are revised to address the E P A and interagency comments. T he science discussion draft, written interagency comments, and EPA’s response to major comments become part of the public record. **Step 7**  **. Completion and posting**

. T he Toxicological Review and IRIS summary are posted on the IR I S website ( h ttp://www.epa.gov/iris/ ).

The remainder of this Preamble addresses step 1, the development of a draft Toxicological Review

. I R IS 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). T ransparent application of scientific judgment is of paramount importance

. T o provide a harmonized approach across IRIS assessments, this Preamble summarizes concepts from these guidelines and emphasizes principles of general applicability.Toxicological Review of Hexahydro

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## 3. Identifying and selecting pertinent studies

**3.1. Identifying studies**  Before beginning an assessment, the EPA 

### 1 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 39 40 41 42 43 44 conducts a comprehensive search of the primary scientific literature

. T he literature search follows standard practices and includes the PubMed and ToxNet databases of the Nation al Library of Medicine, Web of Science, and other databases listed in the EPA’s HERO system (Health and Environmental Research Online, h ttp://hero.epa.gov/ )

. S earches for information on mechanisms of toxicity are inher e ntly specialized and may include studies on other agents that act through related mechanisms. Each assessment specifies the search strategies, keywords, and cut

- of f dates of its literature searches. T h e EPA posts the results of the literature search on t h e IRIS web site and requests information from the public on additional studies and ongoing research. The E PA also considers studies received through the IRIS Submission Desk and studies (typically unpublished) submitted under the Toxic Substances Control A ct or the Federal Insecticide, Fungicide, and Rodenticide Act

. M aterial submitted as Confidential Business Information is considered only if it includes health and safety data that can be publicly released

. I f a study that may be critical to the conclus io ns of the assessment has not been
peer

- r eviewed, the EPA will have it peer

reviewed. The EP A 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 mo deling. In assessments of ch emical mixtures

, mixture studies are preferred for their a b ility to reflect interactions among compo ne nts. The literature search seeks, in decreasing order of preference ( U .S. EPA, 2000b, §2.2 ; 1986b, §2.1 )]: – Studies of the mixture being assessed. 

### 45

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80 81 82 83 – Studies of a sufficiently similar mixture

. I n evaluating similarity, the

assessme n t considers the alteration of mixtures in the environment through partitioning and transformation.

– S tudies of individual chemical components of the mixture, if there are not adequate studies of sufficiently similar mixtures.



### 3.2. **S electing pertinent epidemiolo**  **gic 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

- ba sed surveys (for example, NHANES) provide the strongest e p idemiologic evidence, especially i f they collect information about individual exposures and effects.

– E cological studies (geographic correlation studies) relate exposures and effects by geographic area. T h ey 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.

– C ase reports of high or accidental exposure lack definition of the population at risk and the expected number of case s

. T hey 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**  Exposure route is a key design 

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31 32 33 34 35 36 37 38 39 consideration for selecting pertinent experimental animal studies or human clinical studies.

– S tudies of oral, inhalation, or dermal exposure involve passage through an absorption bar ri er and are considered most pertinent to human environmental exposure.

– I njection or implantation studies are

o ften considered less pertinent but may provide valuable toxicokinetic or mechanistic information

. T h ey also may be useful for identifying effect s in

a nimals 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.

– S tudies of effects from chronic

ex posure are most per tin ent to lifetime human exposure.

– S tudies of effects from less

- th an

chronic exposure are pertinent but less preferred for identifying effects from lifetime human exposure

. S uc h studies may be indicative of effects from less

- th an

- lifetime human
exposure.
Short- d uration 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

. A cco r dingly, specialized study designs are used for these effects ( U.S. EPA, 2006b

, 1 998
, 1996
, 1991 ). 

## 4. Evaluating the quality of individual studies
After the subsets of pertinent 

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76 77 epidemiologic and experimental studies have been selected from the literature searches, the assessment evaluates the quality of each individual study

. T his evaluation considers the design, methods, conduct, and documentation of each study, but not whether the results are positive, negative, or null. T he objective is to identify the stronger, more informative studies based on a uniform evaluation of quality characteristics across studies of similar design.



### 4.1. **E valuating the quality of**  **epidemiologic studies** The assessment evaluates design and methodological a spects 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.

– Defi nition and selection of the study group and comparison group.

– A scertainment of exposure to the chemical or mixture.

– A scertainment of disease or health effect.

– D uration of exposure and follow

- up and adequacy for assessing the occurrence of effects.

– C har acterization of exposure during critical periods.

– S ample size and statistical power to detect anticipated effects.

– P articipation rates and potential for selection bias as a result of the achieved participation rates.

– M easurement error (can lead to misclass ification of exposure, healthToxicological Review of Hexahydro

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40 41 types of information bias.

– Potential confounding and other sources of bias addressed in the study design or in the analysis of results

. T he basis for consideration of confounding is a re as onable 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 evalu a ting 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

- v itro study, or human clinical study ( U .S. EPA, 2005a

, 1998
, 1996
, 1991 )

. R esearch 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.

– N ature of the assay and validity for its intended purpose.

– C haracterization of the nature and extent of impurities and contaminants of the administered chemical or mixture.

– C haracteriza tion of dose and dosing regimen (including age at exposure) and their adequacy to elicit adverse effects, including latent effects.

– S ample sizes and statistical power to detect dose

- r elated differences or trends.

– A scertainment of survival, vital signs, dis eas e or effects, and cause of death. – Control of other variables that could 

### 42 43

44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60 61 62 63 64 65 66 67 68 69 70 71 72 73 74 75

76 77 78 79 80 81 82 83 84 85 86 87 88 influence the occurrence of effects.

The assessment uses statistical tests to evaluate whether the observations may be due to chance

. T h e standard for determining statistical s ignificance of a response is a trend test or comparison of outcomes in the exposed groups against those of concurrent controls

. I n some situations, examination of historical control data from the same laboratory within a few years of the study may improv e the analysis

. F or an uncommon effect that is not statistically significant compared with concurrent controls, his t orical controls may show that the effect is unlikely to be due to chance

. F or a response that appears significant against a concurrent con t rol response that is unusual, historical control s 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 experiment a l studies of these effects ( U.S. EPA, 2005a

, 1998
, 1996
, 1991 )

. I n multi

generation studies, agents that produce developmental effects at doses that are not toxic to the maternal animal are of special

concern

. E ffects that occur at doses associated with mild maternal toxicity are not assumed to result only from maternal toxicity

. M oreover , 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

. T h ere 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 consid ers the study quality characteristics in this section to identify the strongest studies or types of study

. T h e tables present details from these studies, and the assessment explains the reasons for not reporting details of other studies or groups of stu dies that do not add new information

. S upplemental informationToxicological Review of Hexahydro

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40 41 42 43 including those not summarized in the tables. The assessment discusses strengths and limitations that affect the interpretation of each study. I f 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 EP A asks peer reviewers to identify studie s 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. T his inference is based on information from pertinent human studies, animal studies, and mechanistic studies of adequate quality

. P o sitive, negative, and null results are given weight accord i ng to study quality. Causal inference involves scientific judgment, and the considerations are nuanced and complex

. S everal health agencies have develo p ed frameworks for causal inference, among them the U.S

. S u rgeon General (CDC, 2004; HEW, 1964), th e In ternational Agency for Research on Cancer ( I ARC, 2006 ) , the Institute of Medicine (IOM, 2008), and the EPA (2010, §1.6; 2005a, §2.5). A l though developed for different purposes, t he frameworks are similar in nature and provide an established structure and language for causal inference

. E a ch 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 t o chance, bias, or other factors

. M odest relative risks, 

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80 81 82 83 84 85 86

87 88 89 90 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

. R eproducibility of findings constitutes one of the strongest arguments for causation

. D iscordant results sometimes reflect differences in study design, exposure, or confounding factors. **Specificity of association:** A s originally intended, this refers to one cause associated with one effect. Current understanding that many agents cause multiple effects and man y 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. **Biologi c**  **gradient (exposure**

**- response**  **relationship):** E xposure

- response relationships strongly suggest causation

. A mo notonic increase is not the only pattern consistent with causation

. T he presence of an exposure

- r esponse gradient also weighs against bias and

co n founding as the source of an association. **Biologic plausibility:** A n inference of causation is strengthened by data demonstrating plausible biologic mechanisms, if available

. P lausibility may reflect subjective prior beliefs if there is insufficient under s tanding of the biologic process involved.

**Coherence:** A n inference of causation is strengthened by supportive results from animal experiments, toxicokinetic studies, and short

- te rm tests

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### 1 2 3

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29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 cha nging disease patterns in the population. **“Natural experiments”:** A c hange in exposure that brings about a change in disease frequency provides strong evidence, as it tests the hypothesis of causation

. A n

ex am ple would be an intervention to reduce exposure in the workplace or environment that is followed by a reduction of an adverse effect. **Analogy:** In f or mation on structural analogues or on chemicals that induce similar mechanistic events can provide insight into causation.

These considerations are consiste n t w ith guidelines for systematic reviews that evaluate the quality and weight of evidence

. C onfidence is increased if the magnitude of effec t is large, if there is evidence of an
exposure

- response relationship, or if an

a ss ociation was observed and the p la us ible biases would tend to decrease the magnitude of the reported effect. C on fidence is decreased for study limitations, inconsistency of results, indirectness of evidence, imprecision, or reporting bias ( Guy att 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

. T he o bjective is to determine whether a credible association has been observed and, if so, whether that association is consistent with causation

. I n d oing this, the assessment explores alternative explanations (such as chance, bias, and confounding) and draws a conclusion about wheth er 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 ev idence of association between exposure to the agent and occurrence of a health effect. Sufficient epidemiologic evidence of an 

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88 89 90 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

as sociation 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: T h e available studies do not permit a conclusion regarding the presence or absence of an association. Epidemiologic evidence consistent with no causal association: S e veral adequate studies covering the full range of human exposures and considering susceptible populations, and for which alternative explanations such as bi a s and confounding can be ruled out, are mutually consistent in not finding an association.



### 5.3. **Eva luating evidence in animals** 
For each effect, the assessment evaluates the evidence from the animal experiments as a whole to determine the extent to which the y indicate a potential for effects in humans

. C o nsistent results across various species and
st rains increase confidence that similar results would occur in humans

. S everal concepts discussed by Hill ( 1 965 ) are pertinent to the weight of experimental resu lts: consistency of response, dose

- re sponse relationships,
st r ength of response, biologic plausibility, and

c oherence ( U.S. EPA, 2005a, §2.2.1.7 ; 1994, Appendix C ). In weighing evidence from multiple experiments, U.S. EPA ( 2 005a, §2.5 ) distinguishes : Conflicting evide nce (that is, mixed positive and negative results in the same sex and strain using a similar study protocol) fromToxicological Review of Hexahydro

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42 43 negative results are in different sexes or strains or use different study protocols). Negat ive or null results do not invalidate positive results in a different experimental system . T h e EPA regards all as valid observations and looks to explain differing results using mechanistic information (for example, physiologic or metabolic differences

ac r oss 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 d evelopmental and reproductive effects ( U .S. EPA, 2006b

, 2005a

, b

, 1998
, 1996
, 1991 )

. A ccordingly, the assessment determines whether critical periods have been adequately investigated

. S i milarly, the assessment determines whether the database is adequate to evaluate other critical sites and effects. In evaluating evidence of genetic tox icity:

– D emonstration of gene mutations, chromosome aberrations, or aneuploidy in humans or experimental mammals ( in v ivo ) provides the strongest evidence.

– Th is is followed by positive results in lower organisms or in cultured cells ( in v itro ) or for other genetic events.

– N egative results carry less weight, partly because they cannot exclude the possibility of effects in other tissues ( I ARC, 2006 ). For germ

- cell mutagenicity, The EPA has defined categories of evidence, ranging from positive re s ults 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. – The biolog ic plausibility of a causal 

### 44 45

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79 80 81 82 83 84 85 86 87 88 89 interpretation of human studies.

– The generalizability of animal studies to humans.

– Th e susceptibility of particular populations or lifestages. The focus of the analysis is to describe, if possible, me chanistic pathways that lead to a health effect . T h ese pathway s encompass:

– T oxicokinetic 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.

– T oxicodynamic 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 m odes of action an d associated key events ( key events being empirically observable, n ecessary precursor steps or biologic markers of such steps; m ode of action being a series of key events involving interaction with cells, operational and anatomic changes, and resulting in disease). P ertinent information may also come from studies of me t abolites or of compounds that are structurally similar or tha t act through similar mechanisms

. I n formation 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 ab out 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 c a n come from experimental challenge to the hypothesized mode of action, in which studies that suppress a key event observe suppression of the effect

. S upport for a mode of action is meaningfully strengthened by consistent results in different experimental mo dels, muchToxicological Review of Hexahydro

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22 23 24 25 26 27 28 29 30

31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 48 the same model. T he assessment may consider various aspects of causation in addressing this question. 2) **Is the hypothesized mode of action relevant to humans?**  Th e assessment reviews the key events to id e ntify critical similarities and differences between the test animals and humans. S ite concordance is not assumed between animals and humans, though it may hold for certain effects or modes of action

. I n formation suggesting quantitative differences in do s es where effects would occur in animals or humans is considered in the dose

- r esponse analysis. C urrent 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

m ay entail higher exposures. 3) **W hich populations or lifestages can be**  **particularly susceptible to the hypothesized mode of action?**  T he assessment reviews the key events to identify populations and lifestages that might be susceptible to their occurrence

. Q ua n titative differences may result in separate toxicity values for susceptible populations or lifestages.

The assessment discusses the likelihood that an agent operates through multiple mode s of action

. A n uneven level of support for different modes of actio n can reflect disproportionate resou rces spent investigating them ( U .S. EPA, 2005a, §2.4.3.3 )

. I t 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

. K ey data include the ability of the agent or a metabo l i te to react with or bind to DNA, positive results in multiple test systems, or similar properties and structure

- a ctivity relationships to mutage ni c carcinogens ( U.S. 

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83 84 85 86 87 88 89 90 91 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). I n doing this, the assessment develops a narrative that integrates the evidence pertinent to causation. T o provide clarity and consistency, the narrative includes a standard hazard descriptor

. F or example, the following standard descriptors co mb ine epidemiologic, experimental, and mechanistic evidence of carcinoge nicity ( 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 ani mal 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: T he evidence demonstrates a potential hazard to humans but does not meet the criteria fo r carcinogenic

. T here 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: T he evidence raises concern for effect s in humans but is not sufficient for a stronger conclusion

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nadequate information to assess 

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. C onflicting 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

. D iffering 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. T here 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 b elo w a defined dose. Multiple descriptors may be used if there is eviden ce that carcinogenic effects differ by dose range or exposure ro u te ( U.S. EPA, 2005a, §2.5 ). Another example of standard descriptors

come s from the EPA’s Integrated Science Assessments, which evaluate cau s ation for the effects of the criteria polluta nt s in ambient air (U.S. EPA, 2010, §1.6). Causal relationship: S ufficient evidence to conclude that there is a causal relationship

. O bservational studies cannot be explained by plausible alternatives, or they ar e supported by other lines of evidence, for example, animal st udies or mechanistic information. Likely to be a causal relationship: S ufficient evidence that a causal relationship is likely, but important uncertainties remain

. F o r example, observational studies show an association but co

- e xposures 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. Suggestive of a causal relationship: At least 

### 48 49 50 51

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66 67 68 69 70 71 72 73 74

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84 85 86 87 88 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 associat i on. Not likely to be a causal relationship: S everal 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 E PA is investigating an d 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

. T h e decision to derive toxicity values may be linked to the hazard descriptor. Dose

- r esponse analysis requires quantitative measures of dose and response

. T h en, other factors being equal:

– E pidemiologic studies are preferred over animal studies, if quantitative measures of exposure are available and effects can be attributed to the agent.

– A mong experimental animal models, those that respond most like humans are preferred, if the comparability of response can be determined.

– S tudies by a route of human environmental exposure are preferred, although a validated toxicokinetic model can be used to extrapolate across exposure routes.Toxicological Review of Hexahydro

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### 1 2 3 4 5

6 7 8 9

10 11 12 13 14

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33 34 35 36 37 38 39 40 41 42 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

o f t he exposure

- response curve.

– S tudies 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

- mo notonic exposure

response relationships are n ot 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 la r ge number of studies are suitable for dose- r esponse analysis, the assessment considers the study characteristics in this section to focus on the most informative data

. T h e assessment explains the reasons for not analyzing other groups of studies

. A s a check on the selection of studies for dose

- r esponse analysis, t he EPA asks peer reviewers to identify studies that were not adequately considered.



## 7. Deriving toxicity values

**7.1. General framework for dose** 

**** **response analysis**

Th e EPA uses a two

- step approach th at distinguishes analysis of the observed dose

response data from inferenc e s 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. T he modeling yields a point of depart u re (an exposure level near the lower end of the observed range, without significant extrapolation to lower dose s) (Sections 7.2

- 

### 43 44 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60 61 62 63 64 65 66 67 68 69

70 71 72 73 74 75 76 77 78 79 80 81 82 83 84 85 86 87 88 7.3). Extrapolation to lower doses considers what is known about the modes of action for each effect ( S e ctions 7.4

- 7.5) . I f re sponse 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 o f uncertainty and variability (S ec tion 7.6). For a group of agents that induce an effect through a common mode of action, the dose

response analysis may derive a relative potency factor fo r each agent. A full dose

response analysis is conducted for one well

studied i ndex chemical in the group, then the potencies o f other members are expressed in relati v e terms based on relative toxic effects, relative absorption or metabolic rates, quantitative structure

- a ctivity relationships, or receptor binding characteristics ( U.S. EPA, 2005a, §3.2.6 ; 2000b, §4.4 ). Increasingly, the EPA is basing toxicity values on combi n ed analyses of multiple data sets or multiple responses

. T he EPA also considers multiple dose

- r esponse approaches if t hey can be supported by robust data.



### 7.2. **M odeling dose to sites of biologic**  **effects** The preferred approach for analysis of dose is t oxicokinetic modeling because of its ability to incorporate a wide range of data

. T h e preferred dose metric would refer to the active agent at the site of its biologic effect or to a cl o se, reliable surrogate measure. T he active agent may be the adminis t ered chemical or a metabolite. C o nfidence in the use of a toxicokinetic m odel 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 EP A has developed standard ap p roaches that can be applied to typical data sets. T hese standard approaches also facilitate comparison across exposure patterns and species.Toxicological Review of Hexahydro

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38 39 40 41 42 43 44 standardized to a daily average over the duration of exposure

. F or chronic eff e cts, daily exposures are averaged over the lifespan. E xposures during a critical period, ho wever, 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.

– O ral doses are scaled allometrically using mg/kg 3/4 -d ay as the equivalent dose metric across species

. A l lometric scaling pertains to equivalence across species, not across lifestages, and is not used to scale doses from adult humans or mature animals to infants

o r children ( U.S. EPA, 2011 ; 2005a, §3.1.3 ).

– Inhalation exposures are scaled using dosimetry models that apply species

specific physiologic and anatomic factors and c onsider 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

. I f this is done, the assessment descri bes 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 valu e s for use in dose

- r esponse 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

. S uch models require sufficient data to ascertain a mode of action and to quantitatively support model parameters associated with its key events

. B e cause different models may provide equivalent fits to the observed data but 

### 45 46 47 48 49 50 51 52 53 54 55 56 57 58 59 60 61 62 63 64 65 66 67 68 69 70 71 72 73 74 75 76

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87 88 89

90 91 diverge substantially at lower doses, critical biologic parameters should be measured from laboratory studies, not by model fitting

. C onfide nce in the use of a toxicodynamic model depends on the robustness of its validation process and on the results of sensitivity analyses

. P eer 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 avai lab le, the EPA has developed a standard set of empirical (“curve

- f itting”) models ( h ttp://www.epa.gov/ncea/bmds/ ) that can be applied to typical data sets, including those that are nonlinear

. T he EPA has also developed guidance on modeling dose

response 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 d o ses ( U.S. EPA, 2005 a, §3.2.3 ). M odeling i s 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 depa r ture 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 exa m ple, 10% extra risk for cancer bioassay data, 1% for epidemiologic data, lower for rare cancers).

– Fo r nonlinear approaches, both statistical and biologic considerations are taken into account.

– F or dichotomous data, a response level of 10% extra risk is ge nerally used forToxicological Review of Hexahydro

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### 44 45

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82 83 84 85 86 m inimally adverse effects, 5% or 

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41 42 43 lower for more severe effects.

– For continuous data, a response level is ideally based on an established definition of biologic significance

. I n the absence of such definition, one control standard deviatio n from the control mean is often used for minimally adverse effects, one

- h alf 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. **E xtrapolating to lower doses an**  **d response levels**  The purpose of extrapolating to lower doses is to estimate responses at exposures below the observed data

. L ow

- d ose extr ap olation, 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) L inear extrapolat ion is used if the dose

response curve is expected to have a linear component below the point of departure

. T his includes:

– A gents or their metabolites that are DNA

- r eactive and have direct mutagenic activity.

– A gents or their metabolites for which human ex p osures 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 descr i bed by an oral slope factor or an inhalation unit risk, which is the slope of the dose

- r esponse curve at lower doses or concentrations, respectively. 3) N onlinear 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. N onlinear approaches generally should not be used in cases where mode of action has not ascertai ne d

. I f nonlinear extrapolation is appropriate but no model is developed, an alternative is to calculate reference values. 4) B oth linear and nonlinear approaches may be used if there a multiple modes of action

. F o r example, modeling to a low response level c an be useful for estimating the response at doses where a high

- d ose 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

. T hese results are arra y ed, using common dose metrics, to show the distribution of relative potency across various effects and experimental systems. T he assessment then derives or selects an overall slope factor and an overall unit risk for the agent, considering the various do se

- r esponse analyses, the study preferences discussed in S ec tion 6, and the possibility of basing a more robust result on multiple data sets.



### 7.5. **C onsidering susceptible**  **populations and lifestages** The assessment analyzes the available information on populations and lifestages that may be particularly susceptible to each effect

. A t iered approach is used ( U.S. EPA, 2005a, §3.5 ). 1) If an epidemiologic or experimental study reports quantitative results for a susceptible po p ulation or lifestage, these data are analyzed to derive separate toxicity values for susceptible individuals.Toxicological Review of Hexahydro

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### 46 47 48 49 50 51 52 53 54

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70 71 72 73 74 75 76 77 78 79 80 81 82 83 84 85 86 87 88

89 90 91 92 93 2 ) If data on risk

- related parameters allow 

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19 20 21 22

23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 comparison of the general population and susceptible individuals, these data are used to adjust the gene ral- population toxicity values for application to susceptible individuals. 3) In the absence of chemical

- specific data, the EP A has developed age -dependent adjustment factors f or early

- life exposure to potential carcinogens that have a mutagenic mode of actio n

. T here is evidence of early

- l ife susceptibility to various carcinogenic agents, but most epidemi o logic studies and cancer bioassays do not include early

- l ife
exposure

. T o address the potential for early- l ife susceptibility, the EPA

recom me nds ( 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. **R eference values and uncertainty**  **factors**

An oral reference dose or an inhalation reference concentration i s an estimate of an
exposur e (including in susceptible subgroups) that is likely to b e w ithout an appreciable risk of adverse health effects over a lifetime ( U.S. EPA, 2002, §4.2 )

. R eference 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

. R efer ence 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

. F or de velopmental 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 f actors to the point of departure

. I f a point of departure cannot be derived by modeling, a no

observed

- ad verse

- effect level or a lowest

observed

- ad verse

- effect level is used instead

. T h e assessment discusses scientific considerations involving several areas of variability or uncertainty.

**Human variation**

**.** T he 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 e ffect

. A factor of 10 is generally used to account for this variation

. T his 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 susc ep tible 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** 

**.** I f animal results are used to make inferences about humans, the assessment adjusts for cross

species d ifferences

. T hese may arise from differences in toxicokinetics or toxicodynamics

. A c cordingly, if the point of departure is standardized to equivalent human terms or is based on toxicokinetic or dosimetry modeling, a factor of 10 1/2 (rounded to 3) is ap p lied to account for the remaining uncertainty involving toxicokinetic and toxicodynamic differences

. I f a biologically based model adjusts fully for toxicokinetic and toxicodynamic diff e rences across species, this factor is not used

. I n 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 -effec**  **t level to no**

**- observed** 

**** **adverse**

**- effect level** 

**.** I f a point of departure is based on a lowest

- o bserved

adverse- e ffect level, the assessment must infer a dose where such effects are notToxicological Review of Hexahydro

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. T his can be a matter of great 

### 1 2 3 4 5 6 7 8 9 10 11

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28 29 30 31 32 33 34 35 36 37 38 39 40 41 42

43 44 45 46 47 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 ma y be used, depending on the magnitude and nature of
t he response and the shape of the dose

response 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** 

. I f a point of departure is based on subchronic studies, the assessment considers whether lifetime exposure could have effects at lower levels of exposure

. A f actor of 10 is applied to account for the uncertainty in using subchronic studies to make inferences about lifetime exposure

. T his factor may a lso be applied for developmental or reproductive effects if exposure covered less than the full critical period

. A factor other t han 10 may be used, depending on the duration of the
stu d ies and the nature of the response ( U .S . EPA, 2002, §4.4.5 ; 1998, §4.2 ; 1994, §4.3.9.1 ). **Incomplete database.** I f an incomplete database raises concern that further studies might identify a more sensitive effect, o rgan system, or lifestage, the

assess ment 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 )

. T he size of the factor depends on the nature of the database deficiency

. F or ex ample, the EPA typically follows the suggestion that a factor of 10 be applied if both a prenatal toxicity study and a two

generation reproduction st udy are missing and a factor of 10 1/2 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

. T hese results are arrayed, using

common d ose metrics, to sh ow where effects occu r across a range of exposures ( U.S. EPA, 

### 48 49 50 51 52 53 54 55 56 57 58 59 60 61 62 63 64 65 66 67 68 69 70 71

72 73 74 75 76 77 78 79 80 81 82

83 84 85 86 87

88 89 90 1994, §4.3.9 ). The assessment derives or selects an organ - o r system

- specific reference value for each organ or system affected by the agent

. T h e assessment explains the rationale for each organ/system

- s pecific reference value (based on, for example, the highest quality studies, the most sensitive outcome, or a clustering of values)

. B y providin g these organ/system

specific reference values, IRIS assessments facilitate subsequent cumulative risk assessments that consider the combined effect of multiple agents acting at a common site or through common mechanisms ( N RC, 2009 ). The assessment then selects an overall reference dose and an overall reference concentration for the agent to represent lifetime human exposure levels whe re effects are not anticipated to occur. T his is generally the most sensitive organ/system

- s pecific reference value, though consideration of study quality and confidence in each value may lead to a different selection.



### 7.7. **Co nfidence 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

. C onfidence in reference values is based on quality of th e studies used and comp l eteness of the database, with more weight given to the latter

. T h e level of confidence is increased for reference values based on human data su p ported 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 c**  **onfidence:** This is a matter of judgment, between high and low confidence.Toxicological Review of Hexahydro

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### 1 2 3

4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 especially vulnerable to change with further testing.

These criteria are consistent with guidelines for systematic reviews that evaluate the q u ality of evidence

. T hese also focus on whether further research would be likely to change confidence in the estimate of effect ( G uyatt et al., 2008b ). All assessments discuss the significant uncertainties e ncountered in the analysis

. T he EPA provides guidance on characterization of uncertainty ( U .S. EPA, 2005a, §3.6 )

. F or example, the discussion distinguishes model uncertainty (lack of knowledge about the most appropriate experimental or analytic model) and parameter uncerta inty (lack of knowledge about the parameters of a model)

. A sse ssments also discuss human variation (interpersonal differences in biologic susceptibility or in exposures that modify the effects of the agent).

A ugust 2013Toxicological Review of Hexahydro

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# E XECUTIVE SUMMARY

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— DO NOT CITE OR QUOTE Occurrence and Health Effects

Hexahydro- 1,3,5

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- triazine (RDX) is a synthetic chemical used primarily as a military explosive . RDX releases have been reported in air, water, and soil . E xposure to RDX is likely limited to individuals in or around military facilities where RDX is or was produced, used , or stored . O ral exposure may occur from drinking contaminated groundwater or ingestion of crops irrigated with contaminated water. I nhala tion 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 . A nimal studies demonstrate toxicity, including nervous system effects, kidney and other urogenital effects, and male reproductive effects. Results from animal studies provide suggest ive evidence of carcinogenic potential for RDX based on evidence of positive trends in liver and lung tumor incidence in experimental animals . T here are no data on the carcinogenicity of RDX in humans. 

### 1

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18 **Effects Other Than Cancer Observed Following Oral Exposure** EPA identified nervous system effects as a human hazard of RDX exposure. S everal human case reports and animal studies provide consistent ev idence of associations between RDX exposure
and e ffects on the nervous system, including seizures or convulsions . I ncreased mortality was generally observed at RD X d oses that induced nervous system effects, and several studies documented that deaths in most cases were preceded by tremors and convulsions

. Alt hough
mechanistic data ar e insufficient to establish a mode of action (MOA) for RDX

- induced convulsions,
the av ailable information suggests that nervous system effects are mediated by RDX binding to the picrotoxin convulsant site of the GABA A channel, resulting in disinhibition that leads to the onset of
seizures.
EPA identified ki dney and other urogenital eff ects as a potential human hazard of RDX exposure based on observation s i n 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.
A n increased incidenc e of suppurative prostatitis was identified, and is considered a marker for
RDX

- r elated urogenital effects. There is no established MOA for RDX

- related effects on the urogenital system.
Based on t h e finding of testicular degeneration in male mice exposed to RDX in diet for 2 years, in the only mouse s tudy conducted of that duration

, EPA identified suggestive evidence ofToxicological Review of Hexahydro

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27 for male reproductive effects of RDX exposur e. Evidence for effects on other organs/systems, including the liver a nd develop ment al effects

, was more limited than for the endpoints summarized above

. EP A concluded that the evidence does not support effects on other organs/systems, including liver an d d evelopmental effects, as a potential human hazard of R DX exposure

.
**Oral Reference Dose (RfD) for Effects Other Than Cancer**

O r gan

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

- 1 ) . T hese 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 e xposure description Confidence

Nervous system Convulsions 9 × 10 − 4 Subchronic M edium
Kidney/urogenital Suppurative prostatitis 2 × 10 − 3 Chronic L ow Male reproductive Testicular degeneration 2 × 10 − 2 Chronic L ow Proposed overall RfD Nervous system effects 9 × 10 − 4 Subchronic M edium

The overall RfD ( s ee Table ES -2 ) is derived to be protective of all types of hazards associated with RDX exposure. T h e effect of RDX on the nervous system was chosen as the basis for
the o verall RfD because nervous system effects were observed most consistently across studies, species, and exposure durations, and because it represent s t h e most sensitive human hazard of RDX

exposure . I n c idence of seizures or convulsions as observed in a subchron ic gavage stud y ( 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

- r e sponse curve

. B enchmark dose (BMD) modeling was
utilized to derive t h e point of departure (POD) for RfD derivation (expressed as the BMDL 01 ) . A 1% response level was chosen because of the severity of the endpoint ; t his is supported by the
observatio n i n 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 w as used to extrapolate the BMDL 01 to a human equivalent dose (HED) ba s e d on RDX arterial blood concentration

, which was then used for RfD derivation. The proposed overall RfD was calculated by dividing the BMDL 01 -H ED for nervous system effects b y a composite uncertainty factor of 300 to account for extrapolation from animals toToxicological Review of Hexahydro

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27
28
29 subchronic study to a chronic study (3), and deficiencies in the toxicity database (3).

**Table ES -2** 

**. Summary of**  **r eference**  **d ose**  **(RfD) derivation** 

Critical effect Point of departure* UF Chronic RfD
Nervous system effects (convulsions) 90

- d F344 rat study Crouse et al. (2006) BMDL 01-H ED : 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

. T h e resulting POD was converted to a BMDL 01

- HED using a PBPK model based on modeled arterial blood concentration

. T h e concentration was derived from the area under the curve (AUC) of modeled RDX concentration i n a rterial blood, which reflects the average blood RDX concentration for the exposure duration normalized to 24 hours.

**Effects O**  **ther Than Cancer Observed Following Inhalation Exposure**

N o studies were identified that provided useful information on effects observed following inhalation exposure to RDX. Of t h e available human epidemiological studies of RDX, none provided
data that could be used for dose

- r e sponse analysis of inhalation exp o sures . T he single experimental animal study involving inh alat i on exp osure is not public ly 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. T h e refore

, the
available health effects li t e rature does not support the identification of hazards following inhalation exposure to RDX.

**Inhalation R**  **ef erence Concentra**  **tion (RfC ) for**  **Effects Other Than Cancer**

An RfC for RDX could not be derived based on the available health effects data. A dditionally, a PBPK model for inhaled RDX is not available to support route

- to

- r o ute extrapolation from the RfD.

**Evidence for Human Carcinogenicity**

U n der EPA’s Guidelines for Carcinogen Risk Assessment ( U.S. EPA, 2005a ) , the database for RDX provides “suggestive evidence of carcinogenic potential” based on the f inding of statistically significant trends for hepatocellular adenomas or carcinomas and alveolar/bronchiolar adenomas or carcinomas in female, but not male, B6C3F 1 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,
F 3 44 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. T he MOA for liver and lung tumors in experimental animals is not known . A vailabl e in vitro and in vivo genotoxicity assay s w er e largely negative for RDX, suggesting that parent RDX does not interactToxicological Review of Hexahydro

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. N

- nitroso metabolites of RDX generated anaerobically have tested positive in 

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33 some genotoxicity assays ; their contribu tion to the overall carcinogenic potential of RDX is not known

.

**Quantitative Estimate**  **of Carcinogenic Risk**  **f rom Oral Expos**  **ure**
A quantitative estimate of carcinogenic risk from oral exposure to RDX was based on the

i n creased incidence of hepatocellular adenomas or carcinomas and alveolar/bronchiolar adenomas
or carcinomas in female B6C3F 1 mice observed in the carcinogenicity bioassay in mice ( Lish et al.,
1984 ) . T his 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) . T h e initial high dose (175 mg/kg

- day) was reduced t o 100 m g/kg

- day at week 11 due to high mortality.
Althou gh EPA concluded that there is “suggest ive evidence of carcinogenic potential ” for
RDX
, t he Agency determined that quantitative analys is of the mouse tumor data may be useful for providing a sense of the magnitude of potential carcinogenic risk.
EPA c alculated a single oral slope factor (OSF) that considered the combination of tumors

. P o int 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. T he single BMDL 10 so derived from the mouse tumors was extrapolated to the HED us ing BW 3/4 scaling, and a n OSF was derived by linear extrapolation from the BMDL 10

- HED . T he OSF is 4 × 10

− 2 per mg/kg -d ay , based on the liver and lung tumor response in female mice ( L ish et al., 1984 ).

**Quantitative Estimate**  **of Carcinogenic Risk**  **f rom**  **Inhalation Exposure** The carcinogenic potential of RDX by inhalation has not been investigated . A PBPK model to support route

- to

- r oute extrapolation of an inhalation unit risk based on oral carcinogenicity data
was not available

.
**Susceptible Populations an d Lifestages**  **for Cancer and Noncancer Outcomes** Little information is available on populations that may be especially vulnerable to the toxic
effects of RDX . L if estage, and in particular childhood susceptibility, has not been observed in
human or animal st udies of RDX toxicity . In rats, t ransfer of RDX from the dam to the fetus during gestation an d to pups via maternal milk has been reported. Data to suggest males may be more susceptible than females to noncancer t oxicity associated with RDX exposure are limited

. S p ecifically, u rogenital 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 polymor phisms in these metabolic enzymes to affect susceptibility to RDX . S imilarly, individuals with epilepsy or other seizure syndromes that have their basis in genetic mutation to
GABA A receptors may represent another group that may be susceptible to RDX exp osure ; h owever,Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE t here is no information to indicate how genetic polymorphisms may affect susceptibility to RDX. 

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25 **Key Issues Addressed in Assessment**  In most instances, the spectrum of effects associated with chemical exposure will range in severity
, with re latively less severe effects generally occurring at doses lower than those associated with more severe or “frank” toxicity . C o nvulsions 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) e mphasizes 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 t he severity of the effect. F or convulsions

, because of the severity of the effect itself and the strong association with mortality, a b enchmark response (BMR) leve l of 1% was selected for modeling, balancing the quantitative limitations of the available animal bioassays and the severity
of the effe ct . U s e of a BMR of 1% extra risk of convulsions resulted in extrapolation below the range of experimental data and c ould potentially increase uncertainty in the BMD and BMDL values

. The candidate RfD for kidney and other urogenital effects is based on s uppurative prostatitis

. T his organ/system

- specific RfD is based on a dose

- related increase in suppurative prostatitis as reported in a 2

- y ear feeding study in male F344 rats ( Levine et al., 1983 ) , the only
2- year st udy in rats that examined the prostate. S ome reports have hypothesized that the observed suppurative prostatitis was secondary to a bacterial infection unrelate d to RDX toxicity ( ATSDR,
2012 ; Sweeney et al., 2012a ; Crouse et al., 2006 )

. I n 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 assoc ia ted with RDX

exposure

. H istopathological 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. A lthough 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

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# LITERATURE SEARCH STRATEGY | STUDY SELECTION AND EVALUATION

**L iterature Search and Screening**  **Strategy**

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

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28 characterizing the health effects of RDX. T his strategy consisted of a search of online scientific databases and other sources, casting a wide net in order t o identify all potentially pertinent studies

. I n 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 li terature search for RDX was conducted through January 2014 using the databases and general k eywords listed in Table LS -1 ( see Appendix B for further details of the literature search strategy) . Mo re specifically, the literature search for RDX was conducted in four online scientific
databases

— Pubmed, Toxline, Toxcenter, and TSCATS . T he detailed search approach for these databases, including the search strings and number of citations identified per database, is provided

in A ppendix B , Table B-1 . G iven the m ilitary applications of RDX, the Defense Technical Information Center (DTIC) database , a central online repository of defense

- related scientific and technical
information w ithin 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 , T able B-2

. T his search of the five online
databases identified 99 5 citations (after electronically eliminating duplicates)

. T he computerized database searches were supplemented by review of onli n e regulatory sources, “forward” and
“backward” searches of Web of Science (Appendix B , T able B

- 3) , as well as additional refe rences added during development of the toxicological review ( including guidance documents and other references that provide context for evaluating RDX health effects) ; 1 1 3 citations were obtained using these additional search strategies

. I n total, 110 8 ci tations 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 th is 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 ( E PA

- HQ

- ORD

- 2013

- 0430 ; available at www.regulations.gov ) maintained through the development of the assessment.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **Ta ble 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

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- triazine" OR "Hexahydro

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

- trinitro

- 1,3,5

- triazine" OR "1,3,5 -T riaza

- 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

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- triazina" OR "Hexah ydro

- 1,3,5

- trinitro

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- triazin" OR "Perhydro

- 1,3,5

- trinitro

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- 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 keywo rds

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

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

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4 title, abstr act, and in limited instances, full text for pertinence to examining the health effects of
RDX exposure

. T he process for screening the literature is described belo w and is shown graphically in Figure LS

- 1. 1

• 2 1 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 stud ies; these included 16 studies describing phys iologically- based pharmacokinetic (PBPK) models and other toxicokinetic information, 25 studies providing genotoxicity and ot her mechanistic information, 7 acute toxicity studies, and 17 human case reports

. S tudies investigating the effects of acute expos ures and

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

. S tudies 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) d id not preclude its use in one or more other categories. For example, Woody et al. (1986)

, a case re port 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.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE case reports are generally less pertinent for characterizing health hazards associated with chronic o ral and inhalation exposure

. T herefore, information from these studies was not considered for extraction into evidence tables. N evertheless, the se studies were still evaluated as possible sources of supporting health effects information.

• 27 7 references were identified as secondary sources of health effects information (e.g., reviews and other agency assessments) or as studies providing potentiall y useful contextual information (e.g., studies providing information on exposure levels) ; these references were kept as additional resources for development of the Toxicological Review.

• 7 45 references were identified as not being pertinent to an evaluatio n of the health effects of RDX and were excluded from further consideration (see Figure LS

- 1 for exclusion categ ories).

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

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4 Health and Environmental Research Online (HERO) w ebsite ( http://hero.epa.gov/index.cfm/project/page/project_id/2216 ).Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **Fig ure LS**  **-1**

**. Summary of**  **literature search and screening process for RDX** 

**.**Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **S election of Critical Studies for Inclusion in Evidence Tables** 

Selection of Critical Studies

The 21 studies retained after the literature search and screen (Figure LS

- 1) were evaluated 

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19 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 potenti al

. M uch 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

. I mportantly, the evaluation at this stage does not consider study results, or more specificall y, the direction or magnitude of any reported effects.
To facilitate th is ev aluation , evidence tables we re constructed that systematically summarize the important information from each study in a standardized tabular format as recommended by the N R C (2011) . T he 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.
Studi es 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 )

. S u ch study design f laws 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 b ec ause of fundamental issues with study design, conduct, or reporting

. T h e specific studies and basis for exclusion are summarized in Table LS

- 2.

**Table LS -2** 

**. Studies determined not to be informative because of significan**  **t 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

. T he study was conducted 7 yrs after residents were provided the opportunity to connect to a municipal water supply

. O nly one target

- area household reported using private well water for bathing and cooking at the time of the health study

. T he st udy was not considered informative because the design was not able to adequately define the exposed population.Toxicological Review of Hexahydro

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Unpublished report from the DTIC database; Human and animal data One section of the report describes a human case series with no referent gro up

. I ssues 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.

The health effects literature for RDX is not extensive

. All human and experimental animal 

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29 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 e ffects associated with chronic exposure to RDX

. T he se 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 s hort

- term duration or routes of exposure other than oral and inhalation, w ere not included in evidence tables . N evertheless, these studies were considered, where relevant, in the evalua t i on of RDX health hazards

.

**S t**  **udy Evaluation**  In evaluating the evidence to determine whether RDX exposure may pose a hazard for each of the health effects considered in this assessment, me thodological aspects of a study’s design,
cond u ct, and reporting wer e considered in the overall evaluation and synthesis of the pertinent
data

. I n general the relevance and informativeness of the available studies were evaluated as
outlined in the Preamble and in EP A 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 review s of two 2

- year bioassays ( Lish et al., 1984 ; Levine et al., 1983 ) and

o ne 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 ).
T he general findings of this evaluation are presented in the remainder of this section . S tudy evaluation considerations that are outcome specific are discussed in the relevant health effect
sections in Section 1.1.

Human Studies

T he body of literature on RDX includes three studies of populations occupationally exposed to RDX (one case

- c ontrol and two cross

- sectional studies) ( West and Stafford, 1997 ; Ma and Li,
1992 ; Hathaway and Buck, 1977 ).
T o varying degrees, t hese epidemiology studies of RDX are limited by their study design

, uncertainty in estimates of ex posure, inadequate reporting

, and/or failure to account for potential confounding exposures

. A ll t hree studies were based on a relatively small number of participantsToxicological Review of Hexahydro

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‒ 69 exposed workers in the cross

- sectional studies and 32 cases in the case

- control study)

. T he 

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37 study by M a and Li (1992) of Chinese industrial workers provide d 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 B uck (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 perfo r med . I n the case

- control study by West and S tafford (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. I n the study by Hathaway and Buck (1977) that in cluded 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 a ssociated with liver and hematological system toxicity

. T h e case

- control study by West and Stafford (1 997) , which examined hematology
outcomes, did not perform statistical analyses to adjust for other risk factors or occupational exposures (including TNT) . F u rther, the impact of age or gender could not be assessed as the cases
and controls were not mat c hed. The methodological limitations in these three studies we re 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 studi es , the human health effects literature includes 16 case reports that describe effects following acute exposure to RDX. C ase reports are often anecdotal and typically describe unusual or extreme exposure situations, providing little information that would b e useful for characterizing chronic health effects.
T h erefore, 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 A nimal 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

- t erm studies, and four reproductive/developmental toxicity studies in rats and rabbits (including a t wo

- generation reproductive study ) . O nl y one inhalation study of RDX was identified . A s discussed in Appendix B
and Table LS -2 , this inhalation study was considered uninformative and excluded from consideration in the development o f the Toxicological Review because of study design issues ( i ncluding lack of a control group, incomplete information on exposure levels, and inadequate reporting) . Th erefore, evaluation of the experimental animal database for RDX is limited to studies

o f

ora l toxicity

. A n evaluation of the oral toxicity literature, organized by general methodological
feature s , is provided in the remainder of this section.Toxicological Review of Hexahydro

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Th e RDX database consists of health effect studies conducted in multiple strains of ra ts 

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34 (F344, Sprague

- Dawley, CD ), mice (B6C3F 1 ), dog (beagle), and monkey. T he species and strains of animals used are consistent with those typically used in laboratory studies . A ll of these species or strains were considered relevant to assessing the potential human health effects of RDX. S everal studies in the RDX database provided inadequate information on test animals. T he strain of monkey (rhesus or cynomolgus) used in the study by M artin and Hart (1974) was not clear ly specified . I n 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

. O ther studies of RDX were ide ntified that used nonstandard species

, including deer mice ( P eromyscus maniculatus ), western fence lizards ( Sceloporus occidentalis ), prairie voles ( Microtus ochrogaster ), a nd northern bobwhite quail ( Colinus virginianus )

. T hese studies provide information
relevant to RDX toxicokinetics and mechanis m of action on the nervous system , but not health
effects data. T h e refore, these studies are not included in evidence tables, but are discussed where
relevant in the assessment.
Experimental setup

Ge n eral 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 . K ey features of the experimental setup, including the periodici ty and duration of

exposure , t iming of exposure (e.g., gestational days for developmental studies)

, experimental group sample sizes

, a nd interim sacrifices are summarized in the evidence tables . N ote that sample size
was not a basis for exclud i ng a study from con sideration . 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) . E lements of the experimental setup that could influence interpretation of study findings are discussed in the relevant hazard identification
sections of th e as sessment.
Exposure

P ro perties of the test material were also considered in determin ing whether the exposures were sufficiently specific to the compound of interest

. T wo 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

. T h e purity of RDX used in health
effects studies varied from 84

- 9 9.99%

. T he major contaminants were octahydro

- 1,3,5,7

- tetranitro

1,3, 5,7

- t etrazocine (HMX) and water, which are the primary contaminants of RDX produced through the Bachmann process. T he majority of studies used RDX with ~10% impurities; only
C rouse et al. (2006) used 99.99% pure RDX as a test material in their study

. T he toxicity of HMX was reviewed by the IRIS P rogram in 1988 (www.epa.gov/iris) ; histopathological changes in theToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE l iver in male F344 rats and in the kidney in female rats were reported in a 13

- week feeding study. 

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33 N o chronic studies were available to evaluate the carcinogenicity of HMX. T he presence of the impurities introduces some uncertainty in attribution of toxicity to RDX

. H owever, consistenc y in
the doses at which some toxic effects were seen across studies suggest s that the uncertainty associated with the use of less pure test materials may be relatively small. E vidence 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

. I t should be noted that the test
material s employed in these studies are considered representative of RDX that would be released
into the environment.
Differenc es in milling procedures used to generate the test material resulted in the use of RDX of varying particle sizes across studies

. S o me studies utilized a relatively fine particle size (majority of particles were <66 μm in size) while others used test m aterial with comparatively coarse particle size (~200 μm particle size). D ifferences 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 obse r ved across studies, including neurotoxicity

. I nformation on test
material p urity 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 tes t material in the studies by Hart (1974)

, H art (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. O utcome

- specific methodological co nsiderations are discussed in the relevant health effect sections in Section 1.1.
F or 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

. W hile 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 outco me s. Outcomes and data reporting

I n evaluating studies, consideration was given to whether data were reported for all pre

specified endpoints and study groups, and whether any data were excluded from presentation or analysis . Fo r example, it was noted wh ere histopathological analysis wa s limited to control and high

- d ose groups, a study reporting feature that limited the ability to identif y dose- related trends.
I n limited cases, EPA performed additional statistical analysis to identify trends or refine an alyses consistent with EPA guidance (e.g., analyzing developmental data sets on a per litter basis rather than individual fetus) . D ata from studies have been extracted and presented in evidence tables.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Notable features of the RDX database

T hree two

- year toxicity bioassays of RDX are available as unpublished laboratory studies

. 

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35 T he bioassays by L evine 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

. B oth studies included interim sacrifices (at 6 and
12 mo nths)

. C omplete histopathological examinations were performed on all animals in the control and high

- d ose groups; however, only a subset of tissues was examined in the mid

- dose groups, limiting the ability to identify dose

- r elated trends for tissues with incomplete histopathology

. I n the m ouse 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 y ears of dosing (see Table LS -3)

. B ecause they were available only as laboratory r eports , peer reviews of the Levine et al. (1983) and Lish et al. (1984) studies were conducted by EPA in 2012

. T he peer reviewers generally concluded that the reports provided useful information on the toxicity of RDX, noting that there were limitations in interpretati on due to the histopathological analysis and the statistical approaches employed in the reports. A n earlier

two

- ye ar study in the rat by Hart (1976) used a dose range that was low er 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

. H istopathology findings were limited by the lack of pathology examinations in the
mid

- d ose groups and the lack of individual time of death, which impacts the ability to interpret the histopathology data. I n addition, a heating system malfunction on day s 75

- 76 of the study resulted
in the death o f 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 o f study design, conduct, and reporting reflects that range

. A ll but one of the eight short

- term and subchronic toxicity studies of RDX are available as unpublished laboratory studies; only V on Oettingen et al. (1949) was published

. T he majority of studies conducted histopathological examinations on only some of the experimental groups (e.g., control and hig h dose)

. O ne subchronic study Crouse et al. (2006) was peer

- reviewed by EPA in
2012

. T he peer reviewers determined that the report provided useful information on the toxicity of RDX, including an array of endpoints for neurotoxicity and immunotoxicit y

. L imitations 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, con duct, and reporting of experimental animal toxicity studies of RDX are summarized in Table LS -3 . L imitations of these studies were taken into consideration in evaluating and synthesizing the evidence for each of the
health effects in Section 1 .1.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **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 initia l high dose (175 mg/kg

- d) was reduced to 100 mg/kg

- d at wk 11 due to high mortality

. M ortality 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 . D ead 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.

Ch olakis et al. (1980) 13

- wk mouse study ( Experiment 1 ) Dose range was too low to produce effects in mice . H istopathological 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 wk s

. F or the next 11 wk s, 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

. O nly gross observations made at autopsy.
Von Oettingen et al. (1949) 6

- wk dog study The breed of dog was not reported

. O nly gross obs ervations made at autopsy.
Martin and Hart (1974) 90

- d monkey study Species of monkey is unclear (either Cynomolgus or Rhesus)

. S ome test subjects may have had variable dosing due to emesis

. S mall sample size per dose group (n=3).Toxicological Review of Hexahydro

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# 1. HAZARD IDENTIFICATION

**1.1. PRESENTATION AND**  **SYNTHESIS OF EVIDENCE BY ORGAN/SYSTEM** 

**1.1.1. Nervous System**  **E ffects** 

N ervous system effects following RDX exposure have been observed in multiple case 

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27 reports, and the association between RDX exposure and neurobehavioral effects has been examined in a single occupational epidemiology study

. In formation relevant to an examination of the association between RDX exposure and nervous system effects also co mes 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. M emory retention and block design scores 2 were significantly lower among exposed workers (mean concentrations of RDX in two exposed groups: 0.407 and 0.672 mg/m 3 ) compared to unexposed workers from the same plant. H owever , n o significant differences were observ e d between the groups on other neurobehavioral tests (e.g., simple and choice reaction times, block design, and letter cancellation test) (Table 1

- 1 )

. T his study
did not c onsider potential confounders such as alcohol consumption or co

- exposure to TNT.
Ca se reports support an association between RDX exposure and neurological effects (see Appendix C
, Section C. 3)

. S evere neurological disturbances include tonic

- clonic seizures ( formerly known as grand mal seizures) in factory workers ( T estud 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/non

occupational e xposures ( Kas uske 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 t he mother’s clothing ( Woody et al., 1986 )

. Nervous system effects in experimental animals, including seizures and convulsions (used interchangeably by study authors) , t remors, behavioral changes, irritability

, and hyperactivity, have
been observed in the ma jority of chronic, subchronic, and developme ntal studies following oral exposure to RDX ( s ee Table 1

- 2 and Figure 1 -1)

. I n a 2- year dietary stud y in F344 rats ,

2 The 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.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE administration of 40 mg /kg

- day RDX resulted in c onvulsions ( Levine et al., 1983 ) ; convulsion s were 

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38 not observed at lower doses in the same study ( ≤ 8 mg/kg -day) ( Levine et al., 1983 ) or in Sprague

Dawley rats following chron ic dietary administration of 10 mg/kg -day , the highest dose tested ( Hart, 1976 )

. C onvulsions were observed in B6C3F 1 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 )

. S ubchronic dietary exposure was also associated with convulsions in the rat, although doses reported to increase convulsive activity were inconsistent across studies

. C onvulsions were reported in RDX

- ex posed rats at subchronic d oses as low as 8 and 25 mg/kg -day ( Crouse et al.,
2006 ; Von Oettingen et al., 1949 ) . I n contrast, Lev ine 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 th is study

. N o evidence of seizures, convulsions or

tremors w as reported in three subchronic rat studies that used relatively lower doses of RDX (highest administered doses: 10

− 5 0 mg/kg

- day) ( MacPhail et al., 1985 ; Cholakis et al., 1980 )

. R DX 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. C onvulsions were also reported in dogs exposed to 50 mg/kg

- d ay 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 c h allenged with an audiogenic stimulus 8 and 16 hours after treatment. H owever, no audiogenic seizures were observed at the earlier 2 - a nd 4

- hour post

- dosing test periods even though RDX plasma concentrations were elevated throughout the testing period. I n a complementary experiment, Long Evans rats treated daily with 6 mg/kg

- d ay RDX for up to 18 days required fewer stimulation trials to exhibit amygdaloid kindled seizures compared to controls

. N ei ther 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; interpret a tion of these observation s is limited to some extent because the nature and severity of convulsions and seizur e s were not more fully characterized . T he 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 i n male and female F344 rats associated with RDX exposure via gavage (see Table 1-2)

. T r emors were reported following administration of ≥ 12 mg/kg

- day, persisting throughout the
90

- d ay study

. C onvulsions were observed at ≥ 8 mg/kg

- day in male and female rats; information on
duration and onset w as n ot reported ( Crouse et al., 2006 ) . In general, gavage dosing ( Crouse et al.,Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE 2006 ; Cholakis et al., 1980 ) induced convulsio ns a t lower doses than did dietary administration, 

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38 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 freq uently preceded by convulsions or seizures. C rouse 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

. I n a 2

- year study in rats, Levine et al. (1983) observed that tremors and/or convulsions were often seen in high

- d ose animals prior to their death

. F urther , 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 app ea rance (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 d eat h.
Additional neurobehavioral effects associated with RDX exposure in rats included increased hyperactivity, hyper

- r eactivity, fighting, and irritability at doses similar to those that induced tremors, convulsions, and seizures (10

− 1 00 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 )

. H yperactivity and nervousness were also reported in male mice that
received a subchroni c exposure to 320 mg/kg

- day RDX ( Cholakis et al., 1980 ) . N o changes in motor activity, flavor aversion, scheduled

- control led behavior, or acoustic startle response were observed
in a 30

- d ay 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/kg

day m ay have been caused by the oral dosing procedure (gavage) alone. Observations of changes in absolute and relative brain weight were mixed across studies.
A m ong chronic oral studies, a decrease in absolute brain weight of female B6C3F 1 mice (3

− 4%
relativ e to control) was reported at doses ≥35 mg/kg -day ( Levine et al., 1984 ; Lish et al., 1984 )

. C onversely, an increase in absolute brain weight of 2% relative to control was observed in F344 rats at 40 mg/kg

- d ay in another two

- year oral bioassay ( Levine et al., 1983 ; Thompson, 1983 )

. S imilarly elevated absolute brain weights were reported in subchronic assays in B6C3F 1 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 measureToxicological Review of Hexahydro

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effects of RDX on brain weights . Based on an evaluation of the relationship betwee n organ 

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28 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 a nalysis methods should be utilized.
A cross the studies summarized in Table 1 -2 , n ervous system responses to RDX did not show

a p redicted relationship with duration of exposure

. F or example, seizures or convulsions were observed in F 344 rats in some subch ronic studies at doses lower than in studies of chronic
duration , an d at even lower doses in dams exposed for approximately 2 weeks during gestation. I n some studies, seizures appeared soon after dosing, suggesting that seizure induction was more strongly correlated with dose level rather than with d uration 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 seizur e onset. Similarly, nervous system effects across studies did not show a consistent relationship with
dose . Th is lack of consistency may, at least in part, be attributed to d ifferences in the purity or particle size of the test material across studies

. A s suming that increased particle size results in slowed absorption and distribution to the brain, studies that used a larger particle size m ay be expected to produce l ess neurotoxicity in test animals

. T he 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. T his 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 m g/kg -day ) failed to report seizures or convulsions . F inally, differences in study design may
have contributed to d ifferences in reported neurological responses in subchronic and chronic duration studies; i n 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.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **Table 1-1** 

**. Evidence pertaining to**  **nervous system effects in humans** 

Reference and s tudy d esign 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; expose d workers were divided into two groups based on RDX concentration in the air: Concentration (mg/m 3 ) Group A 0.407 (± 0.332) Group B 0.672 (± 0.556) Effect measures a : Five neurobehavioral function tests and five additional memory subtests. Analysis : Varian ce (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 a Symptom data were not included in evidence table because of incomplete reporting.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **Table 1-2** 

**. Evidence pertaining to**  **nervous system effects in animals** 

Reference and s tudy d esign Results
Convulsions and neurobehavioral effects
Lish et al. (1984) ; Levine et al. (1984) Mice, B6C3F 1 , 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 μ m 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 neurolo gical 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, B6C3F 1 , 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 follow ed 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 females a ; no incidence data were reported.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Reference and s tudy d esign Results
Cholakis et al. (1980) Rats, F344, 1 0/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.

C rouse 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 a 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 conv ulsions 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; p article 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 groups a ; no incidence data were reported.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Reference and s tudy d esign 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 10 a
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/gr oup (control);
7 females/group (exposed) 90

− 97% pure
, with 3

− 10% HMX; p article size not specified 0 or 50 mg/kg

- d Diet 6 d/wk for 6 wks Treated dogs exhibited convulsions , excitability, ataxia, and hyperactive reflexes a ; 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 siz e 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 Dos es 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.Toxicological Review of Hexahydro

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- 9 DRAFT

— DO NOT CITE OR QUOTE Reference and s tudy d esign Results
Angerhofer et al. (1986) Rats, Sprague

- Dawley, 39

− 51 mated females/group Purity 90%; 10% HMX and 0.3% acetic acid

o ccurred as contaminants 0, 2, 6, or 20 mg/kg

- d Gavage GDs 6

− 15 Convulsions and hyperactivity a were observed at 20 mg/kg

- d ay;

no incidence data were reported.

Brain w eight
Lish et al. (1984) ; Levine et al. (1984) Mice, B6C3F 1 , 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 b rain w eight

M 0%

− 0.2% 0.61% 0.81%

− 1%

F 0%

− 2%

− 2%

− 4%*

− 3%*

Relative b rain w eight

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 b rain w eight

M 0% 2%

− 1% 2% 2%

F 0%

− 0.3%

− 0.4% 1% 2%*

Relative b rain w eight

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 b rain w eight

M 0%

− − − 2% 2%

F 0%

− − − 4% 2%

Relative b rain w eight

M 0%

− − − 6% 2%

F 0%

− − − 0% 3%Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Reference and s tudy d esign 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 b rain w eight

M 0% 0% 2% 10%

F 0% 0% 4% 2%

Relative b rain w eight

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 a nd 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 b rain w eight

M 0%

− − − 3% 0%

F 0%

− − − 0% 0%

Relative b rain w eight

M 0%

− − − 7%* 10%*

F 0%

− − − 5% 6%

C rouse 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 b rain w eight

M 0%

− 1%

− 0.3% 2% 5%* 7%*

F 0%

− 2% 6% 1% 4% 6%

Relative b rain w eight

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 b rain w eight

M 0% 1% 0.53%

− 6%

− −

F 0%

− 1% 1% 2%

− −
Relative b rain w eight

M 0% 4% 7% 14%

− −

F 0% 0.3% 2% 5%

− − 

### 1 2 3 4 5 6 7 8 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

- r esponse array are those provided in the results section of the technical report ( L evine 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

- da y (abstract and executive summary), or that mortality was observed inToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE rats receiving 100 mg/kg

- d and that hyperactivity to approach, tremors, and convulsions were observed in 

### 1 2 3 4 animals exposed to lethal doses (discussio n).

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 ; Lev ine et al., 1981b ) present subsets of the data provided in the full laboratory report.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUO TE M M

M M

M M

M M

M

M M

M M M

M M M

M

0.1

1 10 100 1000Lish 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 WeightDose (mg/kg-day) significantly changed
not significantly changed

M

- Mortality observed at this dose and above

**Fig ure**  **1-1**

**. Ex**  **posure response array of nervous system effects following oral exposure** 

**. 3** 

3 Due to the severity of the endpoint for convulsions and/or seizures

, a response in treated groups w as determined to be significant (filled circles) in the
exposure

- response array whe re there was an observati on of convulsions and/or seizures reported in the stud y.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE M

echanistic Evidence

T he few studies that have explored the MOA of RDX on the central nervous system have 

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37 focused on potential impacts on neurotransmission

. T hese studies suggest that the MOA for RDX

induced 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 . T he strongest mechanistic information for RDX neurotoxicity comes from documented
inte ractions with the GABA A receptor . GABA is a major inhibitory neurotransmitter in the brain,
and the GABA A receptor has been implicated in susceptibility to seizures ( G alanopoulou, 2008 ) . I t is also a target of many anticonvulsant therapies (e.g., benzodiazepines, propofol, barbiturates) ( Meldrum and Rogawski, 2007 ; Möhler, 2006 ) . T he affinity of RDX for the GABA A 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, W illiams 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

. R DX ( 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 a lter ations 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 t o act through inhibition of acetylcholinesterase (AChE) activity ( M cdonough and Shih, 1997 )

. S ome 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 ) . H owever, the limited data available for RDX do not support AChE inhibition as a primary mechanism because: 1) c ommon AChE

- induced symptoms such as salivation and lacr imation have not always been
observed ( Williams et al., 2011 ) ; 2) blood and brain levels o f 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 targ ets 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 GABA A receptors ( W illiams et al., 2011 ; Williams and Bannon, 2009 ) . S pecifically, RDX showed a significant affinity for the picrotoxin convulsant site of the GABA channel . T he authors demonstrated that RDX treatment in brain slices from the basolateral amygdala inhibit GABA A − mediated inhibitory postsynaptic currents and initiated seizu r e- like neuronal discharges . RDX exposure may reduce theToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE i nhibitory effects of GABAergic neurons, resulting in enhanced excitability that could lead to 

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37 seizures ( W illiams 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

- i nduced 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) ( J efferys 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 pro

convulsant activity in response to amygdaloid kindling at a dose that was approximately half the dose necessary for RDX to induce spontaneous seizures . P o tential limbic system involvement is also suggested given its role in integrating emotional and behavioral responses (including aggression) and the anecdotal observations of hyperactivity, hyper

- r esponsiveness, and irritability
noted across several studies of R DX 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 . Gen es encoding proteins involved in the glutamate pathway were also underex p ressed, indicating a possible mechanism of RDX via excessive glutamate stimulation. T h e authors speculated that this depression of the major excitatory neurotransmitter system could be a negative response to the
increase in seizure likelihood from RDX in f lux 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 B6C3F 1 mice fed 5 mg/kg

- day for 28 days . O ne miRNA, miR

206, was upregulated 26

- f old in RDX

- exposed brains; brain

- derived neurotrophic factor (BD NF) was identified as a downstream gene target of this miRNA, along with two other miRNAs that were upregulated in RDX

- e xposed brains (miR

- 30a and miR

- 195) ( Zhang and Pan, 2009a

, b ) . B DNF is a
member of the neurot rophin family of growth factors, and promotes the survival and differentiation of existing and new neurons . E ffects of RDX on BDNF expression may play a role in RDX neurotoxicity, but the utility of miRNAs a s predictors of toxicity has not been established, and the
contribution , i f 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 si gnaling in the limbic system could represent a likely mechanism for RDX

- in duced hyperactivity and
seizures . Ho wever, the available data are insufficient to identify any specific mode(s) of action for the nervous system effects observed following RDX expos ure.

Summary o f Nervous System Effects

Evidence for nervous system effects associated with exposure to RDX comes from studies in
both humans and animals. O ne occupational study reported memory impairment and decrementsToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE in certain neurobehavioral tests in workers exposed to RDX compared to controls ( Ma and Li, 

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36 1992 ) , and human case reports provide other evidence of an association be tween acute RDX exposure and neurological effects . E leven of 1 6 repeat

- dose animal studies reported neurological effects, including seizures, convulsions, tremors, hyperirritability, hyper

- r eactivity 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 )

. I n most of these studies, the occurrence of neurological effects was dose related

. I n those studies that found no
evide n ce 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 accou nt for the
lack of effect . A lthough the specific mode(s) of action for RDX

- induced nervous system effects remains unknown, evidence t hat RDX exposures may lead to seizures through binding to the GABA A receptor p rovides biological support for this association . EPA identified ne rvous system effects as a human h azard of RDX exposure

.

**1.1.2. Kidney and**  **Other Urogenital**  **System Effects** 

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

was e xamined in a single occupational epidemiology study. Case reports involving accidental exposure to ingested or inhaled RDX provide s ome information on the potential for acute exposures

to RD X to affect the kidney in humans. Organ weight and histopa thology findings from experimental animal studies involving subchronic and chronic ex posure to ingested RDX also provide data r elevant to an examination of the association between RDX exposure and kidney and other urogenital system effects

. A summary of k idney 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 a ffect the kidney and urogenital
system

. R e ported 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 )

. I n many of these case reports, renal parameters returned to normal within a few days following exposure

. N o changes in renal parameters were reported in other individuals exposed to unknown amounts of RDX ( Stone et al.,
1969 ; Kaplan et al., 1965 )

. I n 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 m g/m 3 ), no statistically significant differences in BUN or total serum protein between nonexposed and RDX

- ex posed groups were observed ( Hathaway and Buck, 1977 ) (Table 1 -3 ).Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE S tudies in experimental animals provide some evidence that RDX exposure is associated 

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38 with kidn ey and other urogenital effects (Table 1 -4 and Figure 1 -2)

. D ose

- related increases in absolute and relative kidney weights (19

− 27% compared to control) were observed in male B6C3F 1 mice exposed to RDX in the diet for 2 years ( L ish et al., 1984 ) and a dose

- related increase in relative kidney weights (up to 19%) was observed in female mice

. R elative, but not absolute, kidney weights were increased (20

− 2 1% compared to control) in male and female F344 rats exposed to
40 m g/kg

- day RDX in the diet for 2 years ( Levine et al., 1983 )

. C hanges 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

- r elated 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

- br ain weight rati o) 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.
H istopathological 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 p yelitis, uremic mineralization

, bladder distention and/or cystitis, and suppurative prostatitis were observed in high

- d ose (40 mg/kg

- day) male rats that died spontaneously or were sacrificed in
moribund condition ( L evine 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 B6C3F 1 mice exposed to 320 mg/kg

- d ay 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

- d ay 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 i n 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 RD X studies, HMX, was associated with h istopathological changes in the kidney and alterations in renal function in femaleToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE r ats fed doses ≥ 450 mg/kg

- day HMX for 13 weeks. No effects were observed at doses ≤ 115 mg/kg



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38 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 to tal incidence of suppurative prostatitis was reported in male F344 rats exposed to ≥1.5 mg/kg

- d ay 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 ga v age) exposure to RDX ( Crouse et al., 2006 ; Levine et al., 1990 ; Levine et al.,
1981a ; Levine et al., 19 81b ) . 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 bacteria l infection unrelated to RDX toxicity ( ATSDR, 2012 ; Sweeney et al., 2012a ; Crouse et al., 2006 ) . For example, Crouse et al. (2006)

con cluded 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 like ly that the condition was a result of a n 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 i n SDMS rats, there was an increasing trend with dose in both the scheduled sacrifice (SS) and SDMS groups ; t h e incidence of suppurative prostatitis in the control group was 4% when the SS and SDMS groups are combined

. Additionally, t he dose

related nature of the increase d in cidence suggests that the primary cause (potentially leading to bacterial infection) was treatment

- r elat ed since a more uniform distribution of rats with suppurative prostatitis would be expected with a spontaneous or age

- r elated lesion. The dose

responsiveness could be explained if the infections were secondary to treatment -r elated immunotoxicity, but ther e 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 ass o ciated 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

- d ay RDX for 24 months, although one rat in the
0.3 m g/kg

- day dose group also exhibited these lesions

. Treatment

- related effects o n the kidney (necrosis ) and bladder (distension/obstruction and hemorrhagic cystitis) were also identified in
the 12

- m onth pathology report (see Table s 1 -5 to 1 -7)

. The absence of these observations in the
6- mo nth interim pat ho logy report suggests that an exposure duration of greater than 6 months may be required before RDX

- i nduced effects on the urogenital system are observed

. Suppurative prostatitis was observed with increasing incidence i n each dose group in the study at 24 months.Toxicological Review of Hexahydro

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This document is a draft for review purposes only and does not constitute Agency policy. 1-6 DRAFT

— DO NOT CITE OR QUOTE Considered as a group, treatment

- related kidney and urogenital lesions may have led to a blockage 

### 1
2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19 that resulted in urinary stasis. Reduced urinary flow and/or retrograde flow may have contributed
to an envir onment 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 hypothesi s is consistent with the
observed dose

- related increase in incidence of the suppurative prostatitis ( AT SDR, 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 o b s ervations 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 Levi ne 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 obse rvation of a gross abnormality. Bladder distension was reported sporadically among the lower dose groups (0.3, 1.5, or 8.0 mg/kg

- da y ), 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 a n 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 s tudy d esign Results Hathaway and Buck (1977) Cross

- sectional study, 2,022 workers, 1,491 participated (74% response rate)

. Analys is 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; exp osed subjects assigned to two groups: undetected (<LOD) or ≥ 0.01 mg/m 3 (mean 0.28 mg/m 3 ). Effect measures : Renal function tests (blood) Analysis : Types of statistical tests were not reported (assumed to be t

- tests for comparison of means and χ 2 tests for

c omparison of proportions). Renal function tests in men: mean (standard deviation not reported)

Test Referent (n = 237) RDX exposed Undetected

(n = 22) >0.01 mg/m 3 (n = 45)
B UN 15.5 15.6 16.4 Total protein 7.2 7.2 7.3 No differences were statistically significant. Similar results in

women.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **Table 1-4** 

**. Evidence**  **pertaining to kidney and other urogenital**  **system effects**  **in animals**

Reference and s tudy d esign Results
Kidney weight
Lish et al. (1984) ; Levine et al. (1984) Mice, B6C3F 1 , 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
conta minant ; 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, B6C3F 1 , 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

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This document is a draft for review purposes only and does not constitute Agency policy. 1

- 8 DRAFT

— DO NOT CITE OR QUOTE Reference and s tudy d esign 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%

− −Toxicological Review of Hexahydro

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- 9 DRAFT

— DO NOT CITE OR QUOTE Reference and s tudy d esign 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 te st 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, B6C3F 1 , 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

. H owever, at 12 and 2 4 mo of treatment, this lesion was observed as frequently in control animals as animals treated with RDX.

Hart (1976) Rats, Sp rague

- 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 tr eatment; 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 s acrificed moribund (SDMS); incidence data were not reported for females.

Doses 0 0.3 1.5 8.0 40
Kidney, m edullary 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, s uppurative pyelitis; 24 mo (incidence)Toxicological Review of Hexahydro

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- 10 DRAFT

— DO NOT CITE OR QUOTE Reference and s tudy d esign 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, u remic 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. (1 980) Mice, B6C3F 1 , 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 a nd 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

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— DO NOT CITE OR QUOTE Reference and s tudy d esign 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; ~20 0 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

−

C rouse 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 Histopatholog ical 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, Cyn omolgus 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 subset s of the data provided in the full laboratory report.Toxicological Review of Hexahydro

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— 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 0/15Toxicological Review of Hexahydro

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This document is a draft for review purposes only and does not constitute Agency policy. 1- 13 DRAFT

— DO NOT CITE OR QUOTE 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

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

. 

### 1 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 0/15Toxicological Review of Hexahydro

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This document is a draft for review purposes only and does not constitute Agency policy. 1- 14 DRAFT

— DO NOT CITE OR QUOTE 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 0/15Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE 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: L evine et al. (1983)

.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE M

M M

M M M

M M M M M M

M M M

0.1

1 10 100 1000Lish 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) 1luminal distention; Levine et al., 1983 (rat)cystitis; Levine et al., 1983 (rat)2suppurative prostitis: Levine et al., 1983 (rat)tubular nephrosisCholakis et al., 1980 (mouse)Cholakis et al., 1980 (rat)Cortical cystsCholakis et al., 1980 (rat)subacute prostate inflammationCrouse et al., 2006 (rat)Levine et al., 1990, 1981a,b (rat)Hart, 1974a (dog)medulla mineralizationMartin and Hart, 1974b (monkey)

Chronic Subchronic Chronic Subchronic
↑ Relative kidney weight ↓Relative kidney weight Histopathological lesionsDose (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.

**Fig ure**  **1-2**

**. Exposure** 

**- response array of**  **kidney and urogenita l**  **system effects.** Toxicological Review of Hexahydro

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echanistic Evidence

N o MOA information is available for RDX

- induced kidney and other urogenital effects, 

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37 including suppurative prostatitis . H owever , mechanistic information underlying the neurotoxicity observed with RDX exposure, and the specific affinity of RDX to the GABA A receptor

- convulsant site ( W illiams et al., 2011 ; Williams and Bannon, 2009 ), suggests a biologically plausible role for the
GABA A receptor in RDX

- related effects on the urogenital system and provides some p otential modes
of action f or the effects reported in Levine et al. (1983)

. Alterations in hormonal signaling or circulating levels of estrogen or prolactin may lead to prostatitis

. P rostate inflammati on has been associated with endocrine disruptors in the
environment ( C owin 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 )

. T ypically the inflammation seen is chronic and does not reverse over tim e ( Robinette,
1988 )

. F unctional GABA A receptors have been identified in the anterior pituitary( Zemkova et al.,
2008 ; Mayerhofer, 2001 ) , which also serves as the primary source of prolactin. T hus, the prostate
inflammation observed in the rat in the 2

- year study by Levine et al. (1983) could have be en produced by disruption of pituitary prolactin or other hormonal signal via interference with normal regulatory GABA

- r elated hormonal control

. H owever, no direct evidence for this hypothesized
MOA is a vailable. Levine et al. (198 3) did not evaluate serum endocrine measures or pituitary weights, and pituitary adenomas that could account for higher prolactin levels were not observed

. A M OA based on pituitary

- mediated alterations in endocrine signaling also does not explain the

o t her urogenital lesions observed by Levine et al. (1983)

. Another hypothesis is that the prostate effects could be mediated through an autoimmune inflammatory response

. G ABA A receptor transcripts have been identified in immune cells of mouse models ( R eyes

- García et al., 2007 ; Tian et al., 2004 ) , and GABA A receptor agonists have decreased cytotoxic immune responses and hypersensitivity reactions ( Tian et al., 1999 ; Bergeret et al., 1998 )

. I n 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 thi s effect. H owever, picrotoxin on its own did not significantly alter cytokine production, suggesting the effects are limited to reversal of agonist -i nduced GABAergic activity . I f an autoimmune mechanism was contributing to the effects observed with
RDX e xposure, 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 m ediated through localized
interaction with GABA A receptors, another possibility is that effects would result from direct
interactions with GABA A receptors located on the prostate. G ABA A receptors have been identified on the prostate ( N apoleone et al., 1990 ) , providing a potential mechanism by which RDX could interact directly with the prostate . However, this would require that the prostate is activelyToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE m aintained in a non

- inflamed state, mediated by GABA; RDX bi nding to GABA A receptor

- convulsant 

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36 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 GABA A receptors mediating inputs to the urogenital system . G ABA is believed to play a role in the regulation of urination and bladder capacity (r ev iewed in Fowler et al. (2008) and Yoshimura and de Groat (1997) )

. I n rats, injection of a GABA A receptor agonist inhibits the
urination reflex ( Igawa et al., 1993 ; Kontani et al., 1987 ) . G ABA A 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 )

. R DX would be expected to act like an antagonist and increase bladder activity (which would not result in urinary st asis), although the impact of chronic exposure to RDX acting as a GABA A receptor antagonist is not known . E vidence of GABAergic signaling regulating bladder function, and the hypothesized disruption of that regulation by RDX via interaction with GABA A receptors, may plausibly account for the kidney and other u rogenital lesions, including suppurative prostatitis, observed by Levine et al. (1983) ; however, no evidence to support this hypothesized MOA is
available. In summary, t here are no studies available that inform mechanistically how RDX might lead to kidney and o ther urogenital effects. T here is evidence that RDX binds to GABA A receptors in
neuronal tissues ( W illiams et al., 2011 ; Williams and Bannon, 2009 ) , and i t is biologically plausible
that binding to the GABA recepto r could occur in other tissues as well, account ing for the observed kidney and urogenital effects. Am ong 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 . Ho wever, the ways GABA A receptors work in non

- n euronal tissues and organs is still not well understood, and the MOA by which RDX induces kidney and other ur ogenital 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,

e vi dence for k idney 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 . N o RDX

- related changes in kidney parameters were found in a sm all cross

sectional study of RDX

- ex posed workers ( Hathaway and Buck, 1977 ) . T reatment

- related increases in relative kidney weight were consistently observed in rat s 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 ofToxicological Review of Hexahydro

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36 dose- related chang es associated with RDX exposure

. Histopathological changes in a two

- y ear study in F344 rats, including a dose -related increase in the incidence of suppurative prostatitis in male rats ( L ev ine et al., 1983 ; Thompson,
1983 ) , provides the strongest evidence of RDX

- associated kidney and other urogenital effects . A s discussed above, the incidence of suppurative prostatitis is considered to b e an indicator for the broader array of kidney and other u r ogenital effects seen in this study . A second 2

- year study in Sprague

- Daw l ey 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) . I n light of
the dose

- related increase in suppurative prostatitis and lack of support for an alternative (i.e., non

RDX

- r elated) basis for this effect, EPA identified kidney and o ther urogenital effects as a potential
human h azard of RDX exposure

.

**1.1.3. Reproductive and Developmental**  **Effects**

N o human studies were identified that evaluate the potential of RDX to cause reproductive or developmental effects

. I n formation relevant to an examination of the association between RDX exposure and reproductive and developmental e f fects comes from a 2

- generation study in rats and
studies in rats and rabbits i n volving gestational exposure to ingested RDX. In addition, oral subchronic and chronic studies in experimental animal s p r ovide information useful for examining
the a ssociat i on between RDX exposure and effects on the male reproductive system. A summary of the developmental and reproductive effects a ss ociated with RDX exposure is presented in Tables
1-8 a nd 1 -9 and Figures 1 -3 and 1 -4.

Developmental Effects Animal studies report effects of RDX on offspring sur vival . P up survival rates in the F0 and F1 generations were statistically significantly decreased in RDX

- ex p osed CD rats compared to controls in the only available two

- g en eration reproductive toxicity study of RDX ( Cholakis et al.,
1980 )

, but only at the high est dose tested (50 mg/kg

- day) that also produced toxicity in adults ( neurotoxicity, mortality, and reduced body weights and f oo d consumption ) . D ecreased fetal viability was observed at 20 mg/kg

- d ay 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 . I ncreased resorptions were similarly limited to the highest dose tested (20 mg/kg

- d ay ), i.e., a dose assoc iated with maternal toxicity ( Cholakis et al., 1980 ) . There was no evidence of mater nal toxicity, embryotoxicity or decreased fetal viability in a teratology study of pregnant rabbits exposed to RDX by gavage from GD 7 to 2 9 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 an d length were reported in Sprague

- D aw ley rats exposed to RDX by gavage from GD 6 to 15 ( Angerhofer et al.,Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE 1986 ). 4 Maximum decreases in fetal body weight (9%) and body length (5%) were observed at 

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34 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 R DX at 20 mg/kg

- d ay, a dose associated with maternal mortality ( Cholakis et al., 1980 ) . The larger Sprague

Dawley litter sizes and number of fetuses, compared to F344 rats, may account for the greater statistical power to observe treatment

- r el ated effects. D ose

- related reductions in fetal body weight
were not observed in rabbits at dos e s 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 m g/ kg

- day RDX, a dose that resulted in approximately 30% maternal mortality ( A ngerhofer et al., 1986 ; C holakis et al., 1980 ) . E xamination 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 ) . I ncreased incid ence s 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. G estational exposure to RDX did not result in any
other skeletal abnormalities.

Reproductive Effects Evidence of male rep roductive toxicity is provided by the finding of testicular degeneration in male mice ( T ab le 1 -9 and Figure 1 -4)

. An increased incidence of testicular degeneration was
observed in male B6 C3F 1 mice exposed to ≥ 35 mg/kg

- day RDX for 2 years in the diet (10

− 11%) compared to c on current (0%) and historical (1.5%) controls ( Lish et al., 1984 ) . R eductions in absolute testicular weight were observed, but the magnitude of the effect was small (≤6% compared to controls) and not dose

- r el ated

. An increased incidence of g erm cell degeneration was observed in rats exposed to 40 m g/ kg -da y (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

- t r eated groups ( Levine et al., 1983 ) . N o dose

related 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 )

. C hanges 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 ; L evine et al., 1981a ; Levine et al., 1981b ; Cholakis et al., 1980 ; Hart, 1976
, 1974 ) in testicular weight; in most cases the change s in testicular weight were small (≤10% change compared to control) and no t dose

- related . Based on an evaluation of the relationship between organ weight and body/brain weight to determine which endpoint (organ weight, organ

- to

- bo d y weight ratio, or organ

- to

- b r ain 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 sta tistically significant decreasing trends.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE toxicity, Bailey et al. (2004) concluded that testes weights are not modeled well by any of the 

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14 choices, and that alternative analysis methods should be utilized. Reproductive function was assessed in two separate studies reported by Cholakis et al. (1980) . I n 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

- d ay RDX for 15 weeks prior to mating, although this effect was attributed to decreased
well

- b eing of the males in t h is high

- dose group ( Cholakis et al., 1980 ) . N o specific effects on reproductive fun ction were observed in F0 and F1 rats exposed to ≤ 16 mg/kg

- day RDX in a two

generation study . Th e h ighest dose tested, 50 mg/kg

- day, was associated with reductions in fertility (specifically a decreased number of pregnancies) in the F0 generation, althoug h t hese changes were not statistically significant . Th e fi nding of lower fertility rates only at the 50 mg/kg

day dose, a dose associated with reduced body weight and feed consumption and increased mortality, suggests that effects on reproductive function w er e 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 s tudy d esign 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.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Reference and s tudy d esign Results
Cholakis et al. (1980) Rabbits, New Zealand W hite, 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 mortali ty (16/51) occurred at 20 mg/kg

- d.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Reference and s tudy d esign Results
Offspring growth
Cholakis et al. (1980) Rabbits, New Zealand W hite, 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 morta lity (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 W hite, 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 2/12Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Reference and s tudy d esign 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 tr eatment

- 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 G Ds 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 mortalit y (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

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— DO NOT CITE OR QUOTE M

M M M M M M

0.1

1 10 100Cholakis 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)1Cholakis et al., 1980 (rabbit)Cholakis et al., 1980 (rat)Angerhofer et al., 1986 (rat)

↓ Offspring survival Offpsring growth Morphological developmentDose (mg/kg-day) signficantly changed
not signifcantly changed

M

- Maternal m ortality observed at the highest dose

1 Statistically signficant dose

- related trend (p <= 0.05) by linear trend test, performed for this assessment.

**Fig ure**  **1-3**

**. Exposure response array of**  **reproductive and developmental effects following oral exposure.**Toxicological Review of Hexahydro

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— 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, B6C3F 1 , 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 3/27 a
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
conta minant ; 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

Tes tes, germ cell degeneration; 12 mo b (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, B6C3F 1 , 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)Toxicological Review of Hexahydro

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— 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 microscop ically in control and 320 mg/kg d ay groups; no effects were observed.
Cholakis et al. (1980) Rats, F3 44, 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 avail able.

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.

C rouse 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%Toxicological Review of Hexahydro

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— 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

- d ay dose groups were “notable” when compared to concurrent (0%) and histori ca l (1.5%) incidences.

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

with controls)

. By 24 month s , all male rats (including controls) had testicular masses; testes weights were not recorded, a nd an increased incidence of testicular degeneration was not observed.

c Doses were calculated by the study authors.

d Levine et al. (1981a) is a l aboratory 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

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— DO NOT CITE OR QUOTE M M

M M

M M

M

0.1

1 10 100 1000Lish et al., 1984 (mouse)Hart, 1976 (rat) 1Cholakis 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) 1Lish et al., 1984 (mouse)2Hart, 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 degenerationDose (mg/kg-day) signifcantly changed not significantly changed

1 Increased absolute weight of testes and epididymis

2 Although the study authors did not observe a statistically significant increase in the incidence of testicular degeneration, the y 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.

**Fig ure**  **1-4**

**. Exposure response array of**  **male reproductive effects following oral exposure.**Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Summary of Reproductive and Developmental Effects Developmental studies in rats ( Angerhofer et al., 1986 ; Cholakis et al., 1980 ) and rabbits 

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34 ( Cholakis et al., 1980 ) suggest that developmental effects related to offspring survival, growth, and morphological development were likely associated with severe maternal toxicity. D evelopmental effects were observed only at doses that caused mater n a l mortality

. As noted in EPA’s Guidelines for Developm en tal 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 ca using excessive toxicity, as is the case with RDX, information on developmental effects may be difficult to interpret and of
limited value

. T herefore, EPA concluded that the evidence does not support developmental effects as a potential h uman hazard of RDX exposure.
Testicular effects were reported in male B6C3F 1 mice chronically exposed to RDX in the diet
for 24 months ( L ish 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 m onths, 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 suggest ive evidence of male reproductive effects as a potential human hazard of RDX

exp os ure.

**1.1.4. Liver Effects** 

T he association between RDX exposure and changes in serum liver enzymes was examined in a single occupational epidemiology study. C ase reports involving accidental exposure to RDX provide information on the potential for acute expos ur e to RDX to a ffe ct the liver in humans. In addition, o r gan weight

, histopathology , and serum chemistry findings from experimental animal studies involving subchronic and chronic ex posure to ingested RDX provide data relevant to an examination of the association between RDX exposure and liver e ffects . A summary of the liver effects associated with RDX exposure is presented in Tables 1

- 1 0 and 1

- 11 and Figure 1

- 5. Reports in humans provide limited evidence of liver toxicit y associated with acute exposure
to RDX

. E levated serum levels of aspartate aminotransferase (AST) and alanine aminotransferase (ALT) were reported in several case reports of individuals who ingested unknown amounts of RDX ( Kü çükardali et al., 2003 ; Woody et al., 1 986 ; Knepshield and Stone, 1972 ; Hollander and Colbach,
1969 ; Stone et al., 1969 ; Merrill, 1968 ) ( see Appendix C, Section C.3 )

. L iver biopsies did not reveal any abnormal observations ( Stone et al., 1969 )

. I n other case reports, no significant changes in serum levels of liver enzymes were observed ( Testud et al., 1996b ; Ketel and Hughes, 1972 )

. I n 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 av erage exposure concentration was 0.28 mg/m 3 ) ( Hathaway and Buck, 1977 ) , serum chemistry analysis (including the serum liver enzymes AST,Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE A LT, and alkaline phosphatase (ALP)) revealed no statistically significant differences between 

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38 exposed and unexposed workers (Table 1 -10 ).
In ex perimental animals, the most consistent noncancer liver effect associated with RDX exposure is elevated liver weight in studies of subchronic exposure ( T able 1 -1 1 and Figure 1 -5)

. D o se

- related increases in absolute and relative liver weight were observed in male and female
B6C3F 1 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 ) . I n another 90

- day study, only absolute liver weights were increased
in female F344 rats exposed to RDX by gavage ( Crouse et al., 2006 )

. T he magnitude of liver weight

increases in B6C3F 1 mice and female F344 rats across these studies ranged fro m 4

− 29% in the high

- d ose groups

. M ale 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 )

. I n male and female monkeys exposed subchronically to RDX, a bsolute 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 m g/kg

- day) ( Hart, 1974 ) . C hronic RDX exposures in B6C3F 1 mice and F344 or Sprague Dawley

rats showed a less consistent pattern of liver weight increases

. I nterpretation of liver weight
increases in 2- y ear 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 d iet ( Lish et al., 1984 ) was reduced when mice with liver adenomas or carcinomas were removed from the analysis. I n a 2

- year rat study, absolute live r weight showed no dose

- related changes; however, relative liver weights were increased in high

- d ose (40 mg/kg

- day) males and females (by 11 and 18% compared to controls, respectively) ( Levine et al., 1983 ) . T he cha nges in relative liver weight likely reflected the depressed weight gain in the high

- dose rats (2

- 30% in males and 10

- 15% in females)

. B ased on an evaluation of the relationship between organ weight and body/brain weight to determine which endpoint (organ w eight, 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

- br a in weight ratios

. Nonneoplastic histopathological changes in the liver were not associated with RDX exposure in the majority of experimental animal studies ( C rouse et al., 2006 ; Levine et al., 1990 ; Lish et al., 1984 ; Lev ine 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 )

. T he 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 )

. F or example, the incidence of liver portal inflammation was increased inToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE female but not male rats exposed to 40 mg/kg

- day in the diet for 90 days ( Cholakis et al., 1980 )

. 

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36 T h ere 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 )

. I n 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

. I t should be noted that exposure to HMX, the primary contaminant in several of the RDX studies, w a s associated with histopathological changes
in the livers of male rats fed doses ≥ 4 5 0 mg/kg

- day for 13 weeks. S imilar findings were not
observed in the RDX st u d ies, where the doses of RDX employed in the studies would have result ed in HMX exposure s o f ≤60 mg/kg -day . The contribution of HMX exposure to the overall liver findings in the studies of RDX toxicity is t herefore expected to be negligible.
Clinical chemistry parameters, including serum ALT, AST, and ALP, showed no treatment

related changes indicative of liver toxicity

. S t at istically significant changes in these parameters in some subchronic and chronic toxicity studies in rats and mice were relative l y small (generally <50% of the control mean), were not dose

- r e lated 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

- r el ated
cha n ges in serum cholesterol and triglyceride levels; however, these changes were not consistently

o bserved in males and females within the same study, and patterns of changes were not consistent
across studies

. S p e cifically, serum triglyceride levels were elevated (up to 41%) in female B6C3F 1 mice exposed to RDX in the diet for 2 years, although increases were not dose

- r elat ed ( Lish et al.,
1984 ) ; male mice in the same study did not show a similar increase in triglycerides . I n contrast,

s erum 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 )

. I n 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

- d a y); however, triglyceride levels across the four dose groups in this study did not sh o w a dose

- related response.
Serum cholesterol levels showed a dose

- related increase (38% at the high dose of
100 m g/kg

- d ay) in female B6C3F 1 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. C hanges in
serum cholesterol in male and female F344 r at s exposed to RDX in the diet for 2 years at doses up to 40 mg/kg -d a y ( 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 sma ll (within 38% of control mean) and were not dose related.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **Table 1-**  **10. Evidence**  **pertaining to liver effects in humans** 

Reference and s tudy d esign 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, H MX, 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/m 3 (mean 0.28 mg/m 3 ). Effect measures : Liver function tests. Analysis : Types of statistical tests were not reported (assumed to be t

- tests for comparison of means and χ 2 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/m 3 (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/m 3
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.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **Table 1-**  **11. Evidence pertaining to liver effects in animals**

Reference and s tudy d esign Results
Liver weight

Lish et al. (1984) ; Levine et al. (1984) Mice, B6C3F 1 , 85/sex/group; interim sacrif ices (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 sacrif ices (10/sex/group) at 6 and
12 mo 89.2

− 98.7% pure
, with 3

− 10% HMX as
conta minant; 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, B6C3F 1 , 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 320Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Reference and s tudy d esign 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 conta minants; ~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%

−

C rouse 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 %*

− −Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Reference and s tudy d esign 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, B6C3F 1 , 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 0/4Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Reference and s tudy d esign Results

89.2

− 98.7% pure
, with 3

− 10% H MX 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, B6C3F 1 , 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 wit h a moderate

- sized focus of basophilic cytoplasmic alteration; neither finding was attributed by study authors to RDX treatment.

Levin e 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

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— DO NOT CITE OR QUOTE Reference and s tudy d esign Results
Hart (1974) Dogs, Beagle, 3/se x/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 a nd 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/k g-d RDX; however, the study authors considered the toxicological significance to be uncertain.

Serum chemistry
Lish et al. (1984) ; Levine et al. (1984) Mice, B6C3F 1 , 85/sex/group; interim sacrif ices (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 %*

C rouse 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

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— DO NOT CITE OR QUOTE Reference and s tudy d esign Results
Levine et al. (1981a) ; Levine et al. (1990) ; Levine et al. (19 81b) 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 te st 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 su bsets of the data provided in the full laboratory report.Toxicological Review of Hexahydro

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— 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 1000Lish 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 changesDose (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.** Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Summary of Liver Effects There is limited evidence from reports of human exposure and from studies in experimental 

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35 animals that RDX may affect the liver. S everal human case reports of short- term elevat ions 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

. I n 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 s upported by
lifetime studies in mice and rats

. C hanges in serum liver enzymes were not consistent across studies and the magnitude of change relative to concurrent controls was not indicat ive of liver damage . N onneoplastic 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** 

T h e relationship between exposure to RDX and cancer in human popula tions has not been investigated. The carcinogenicity of RDX has been examined in one ora l chronic/carcinogenicity bioassay in mice ( L i sh 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/ do s e 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 tem p erature sp ike 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 inc i dent and ≥50 rats/sex/group at termination, and it seems unlikely that the mortality associated with the temperature spike wou ld have affected a tumor response in the rats. A su m mary of the evidence for liver and lung tumors in experimental
animals fr om these three bioassays is provided in Tables 1

- 12 an d 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 ( Lis h et al., 1984 ) , the combined incidences of hepatocellular adenomas or carcinomas were increased with increasing RDX doses in
female B6C3F 1 mice as compared to concurrent controls, but not in male B6C3F 1 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 n o t 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 theToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE other hand, there were no increased incidences of hepa tocellular adenomas or carcinomas in 

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31 Sprague

- Dawley rats of either sex exposed to RDX via diet for two years at doses up to 10 mg/kg

day ( Hart, 1976 ). Incidences of hepatocellular neoplasms are presented in Table 1

- 12. The tumor responses are discussed in further detail below.
In the female B6C3F 1 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 n o ted by the authors, the incidence of hepatocellular adenomas or carcinomas at RDX doses ≥35 mg/kg

- da y (19% at both doses) was also statistically significantly elevated when compared to the mean historical control incidence for
female B6C3F 1 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) st u dy 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 )

. T here 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 in cidence 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 t he exposed groups (0/55, 0/52, 2/55, 2/31) relative to the control (1/55), as compared with the mice. T her e i s less confidence that the final incidence in the highest

- dose group accurately reflects lifetime cancer incidence bec a use of low survival and no time

- to

- death information to estimate mortality

- adj u sted incidences; the available informati on may underestimate lifetime cancer incidence by overestimating the nu m ber of rats truly at risk. Some perspective on the magnitude of response is provided by comparing with incidence rates in

5 Comparison 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 u sed in the Lish et al. (1984) study. EPA Guidelines for Carcinogenic Risk Assessment ( U.S. EPA, 2005a ) al so 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.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE historical controls, despite the limitation of this comparison due to the historical data originating 

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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 ra t s 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 s tudy d esign Results
Lish et al. (1984) ; Levine et al. (1984) Mice, B6C3 F 1 , 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 3/31 b
Hepatocellular carcinomas (incidence) a

M 13/63 20/60 16/62 18/59 6/27

F 0/65 4/62 3/64 6/64 3/31 a
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 2/31 b
Hepatocellular carcinomas (incidence) a

F 0/67 1/62 3/63 2/64 2/31 b
Hepatocellular adenoma or carcinoma combined (incidence) a

F 1/67 c 4/62 5/63 c 10/64 4/31 b
Hart (1976) Rats, Sprague

- Dawley, 100/sex/group Doses 0 1.0 3.1 10
Neoplastic nodules (incidence) aToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Reference and s tudy d esign 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 2/31 b

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. 0 5), identified by the authors.

a T he incidences reflect the animals surviving to m onth 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 an d 64, respectively).Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Lung t umors 

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16 Cochran

- Armitage trend test s (as performed for this review, p = 0.019 ) found s tatistically significant pos it ive trend s in the incidence s of alveolar/bronchiolar adenomas in female B 6C3F 1 mice, alveolar/bronchiolar car c inomas in male mice, a nd alveolar/bronchiolar adenomas or
carcinomas combin e d in female mice

. T he combined incidence in male B6C3F 1 mice did not show a statistically significant tre n d ( 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 g roups in the mouse study, Lis h 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

. T he i n cidence of alveolar/bronchiolar adenomas or carcinomas was not
increased in either se x of Sprague

- Dawley rats exposed chronically to RDX at doses up to 10 mg/kg

day ( Hart, 1976 ) or in F344 rats of either sex exposed chronically to RDX at doses up to 40 mg/kg

day ( Levine et al., 1983 )

.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **Table 1-**  **13. Lung tumors observed in chronic animal bioassays** 

Reference and s tudy d esign Results
Lish et al. (1984) ; Levine et al. (1984) Mice, B6C3F 1 , 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
3/31 b
Alveolar/bronchiolar carcinomas (incidence) a

M

F 3/63
3/65 6/60
1/62 3/62
3/64 7/59
3/64 5/27 b
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
7/31 b
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 T he incidences reflect the animals surviving to m onth 12.

b Statistically significant trend ( p < 0.05) was identified using Cochran

- Armitage trend test performed by EPA.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE M

echanistic Evidence There are few mechanistic data to support a MOA determination for either liver or lung 

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The increase in l iver 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 B6C3F 1 mice ( Lish et al., 1984 ; Cholakis et al., 1980 ) raise s the possibility of RDX

- related liver cell proliferation as a precursor to tumorgenicity

. Sweeney et al. (2012b) reviewed hypothe sized MOA’s for carcinogenicity and conclude d that a MOA involving a proliferative response generated by tissue

- derived oxidative metabolites of RDX was the most plausibl e of the MOAs consider ed

, but acknowledged that the overall support for th is MOA was
limited . T he f ollowing lines of evidence do not support a metabolite

- based proliferative response as the MOA for RDX carcinogenicity:

• t he 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 p recursor to these tumors ;

• i ncreased liver weight was also observed in rats and male mice where tumors did not occur;

• n o studies were available that directly measured RDX

- induced cell proliferation rates; and

• n o information was available to rule out non -p recancerous 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. Sw eeney et al. (2012b) propose d that the increased incidence of liver adenomas
and carcinomas in female mice ( Parker et al., 2006 ; Lish et al., 1984 ) may result from liver

generated metabolites as the most likely agents responsible for liver tumors

. Sweeney et al. (2012b) estimated a n appr oximately 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. Th es e authors hypothesized a non

- linear, cell proliferati on MOA in conjunction with t he 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 multi ple site s and species . T his observation is inconsistent with the finding in Lish et al. (1984) that showed positive trend s in the incidence of both alveolar/bronchiolar adenomas or carcinomas and liver tumors. In contrast to the neg ative results for RDX oxidative metabolites, there are some positive genotoxicity results for the N

- n it roso metabolites of RDX, specifically hexahydro -1- nitroso

3,5

- d initro

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- triazine [MNX) and hexahydro

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- triazine (TNX)

. M NX 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 humansToxicological Review of Hexahydro

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. M NX has tested positive in some in vitro assays, including 

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38 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 )

. M NX was not mutagenic in S. typhimurium (strains TA98, TA100, TA1535, TA1537, and TA1538), with or wi thout the addition of the S9 metabolic activating mixture ( Pan et al., 2007 ; Snodgrass, 1984 )

. W hen 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) , M N X and TNX were mutagenic

. N

- n it roso metabolites, includin g MNX and TNX, are generated anaerobically and are likely a result of bacterial transformation of parent RDX i n the gastrointestinal tract to various N

nitroso derivatives ( Pan et al., 2007 )

. E xposure 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 . Ho w ever, in pigs the N

- nitr oso metabolites of
RDX have been identified o nl y in trace amounts in urine compared to the major metabolites, 4

- nitro

- 2 ,4

- d iazbutanal and 4

- nitro

- 2,4

- diazbutanamide . T hus

, the contribution of the N

- nitroso metabolites to the overall carcinogenic potential o f R DX is unclear. Aberrant expression of microRNAs (miRNAs) was observed in the brains and livers of
female B6C3F 1 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

. H o wever, the pattern of induction was not always c onsistent in the livers of RDX

- treated mice (e.g.

, m IR

- 9 2a was downregulated in liver tissue samples when it is typically upregulated in hepatocellular carcinomas) ( Sween ey 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 a berrant expression of a specific miRNA (or suite of miRNA’s) plays a role in the MOA o f RDX carcinogenicity is unknown

. M icroarray analysis of gene expression in male Sprague

- Daw ley rats after exposure to a single oral (capsule) dose of RDX revealed a gen e r al 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 summ ary , the available evidence indicates that RDX is not mutagenic ( s ee Appendix C,
Section C .4 ) ; h o wever, anaerobically

- derived N

- nitroso metabolites have demonstrated some genotoxic potential. W h ile these metabolites have been measured in the m ouse ( Pan et al., 2007 )
and mini pig ( 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

- d e rived oxidative metabolites of RDX has been proposed, but is not supported by t he available data. I n light of limited information on precursor events leading to the observed
liver and l ung tumor response in R DX

- exposed rodents and lack of toxicokinetic information on RDX metabolites, n ei ther a cell proliferative MOA or a mutagenic N- nitroso metabolite MOA isToxicological Review of Hexahydro

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. T hus, the MOA leading to the increased incidence of liver and lungs tumors is not 

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**1.1.6. O**  **ther Toxicological Effects** There is limited evidence that RDX can produce systemic effects in several organ s/systems

, includi n g the eyes, and the musculoskeletal, cardiovascular, immune, and gastrointestinal systems.
H o wever, t here 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
Table s 1 -14 a nd 1 -15.

Ocular Effects

T h ere are no reports of ocular effects in human case reports or epidemiological studies.
T h e incidence of cataracts was st atistically 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% i n female F344 rats exposed to 40 mg/kg

- day RDX in the diet for 2 years, compared to 32% in the control group ( L ev ine et al., 1983 ) . A ft er 76 weeks of

exposure , the incidence of cataracts in female rats at 40 mg/kg

- day (23%) was also elevated

com p ared to controls (6%) . T h e 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

. O nly 2 rats (dose groups not reported) were observed to have mild
cataracts in a 90

- d a y study of male and female F344 rats exposed to RDX at doses up to 15 mg/kg

day by gavage; however, t h e authors noted that these observations are common in F344 rats at
4 m onths of age and should not be attributed to treatment ( Crouse et al., 2006 ) . F urthermore,

c ataracts 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 B6C3F 1 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 ca taracts in male mice was initially noted by Lish et al. (19 84) , but was not confirmed when mice used for orbital bleedings were excluded from the analysis, suggesting
the effect was not treatment r elat ed

.
Cardiovascular Effects

H uman evidence for cardiovascular effects is limited to case reports that include
observations of t r ansient arterial hypertension in male Italian workers following inhalation of RDX during manufacturing ( B ar sotti 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 weigh t ( Küçükardali et al., 2003 ) (see A ppendix C, Section C.3 ).
Incons istent observations of cardiovascular effects have been reported in animal studies.
A n increase in the relative heart

- to

- b ody weight ratio was observed at the highest dose tested inToxicological Review of Hexahydro

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35 dietary administration of RDX ( Lish et al., 1984 ; Levine et al., 1983 ) ; how ever , this dose also resulted in reductions in body weight in both males and females . D ose

- related decreases in absolute heart weight were reported following su b chronic 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 m g/kg

- d ay) ( 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 ex p osure to RDX in the diet
for 9 0 day s ( Cholakis et al., 1980 ) . In each study, the finding of myocardial degeneration w as 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 m usculoskeletal effects in humans consists of c ase reports that include
observations o f muscle twitching, myalgia/muscle soreness, a nd muscle injury as indicated by elevated levels of AST or myoglobinuria ( Kü ç ükardali 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 alte rations in
mice ( L ish 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) . T h r ee cases of SLE were initially identified among workers at

one U. S. Army mu nitions plant ; however, upon further investigation of 69 employees at five U.S

. Army munitions plants with p o tential exposure to RDX, no additional cases of SLE were identified ( H athaway and Buck, 1977 )

. I ncreased 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 ).
I n animal studies , increased WBC count in female rat s following subchronic dietary exposure to RDX was the only dose

- r el ated 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 ) . C hanges in spleen weights were observed across studies, but the responses were not consistent and did not appear to be dose

- r elat ed . F or 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, whileToxicological Review of Hexahydro

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- day (not statistically 

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35 significant) . Across studies, there wa s no significant or dose

- dependent pattern of response to

s u ggest that the WBC changes reflect RDX

- in duced immunotoxicity

. N o dose

- related immun e effects from o r al exposure to RDX were observed in other animal studies, including a 90

- day study

in F 344 r ats specifically designed to evaluate immunotoxic ity (paramet ers included evaluation of red and white blood cell populations, proportion of cell surface markers , c e llularit y in proportion to organ weight, B and T cells in the spleen

, an d CD4/CD8 antigens of maturing lymphocytes in the thymus ) ( Crouse et al., 2006 ) . R outine clinical and histopathology evaluations o f 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 i n creased WBC counts in female rats ( Levine et al., 1981a ; Levine et al.,
1981b ) . E vidence 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 ; Le vine 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 RD X poisonings , and generally concurrent with severe neurotoxicity ( Kas uske et al., 2009 ; Dav ies 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 ) . I n animal studies, n ausea and vomiting have also been observed following oral exposure of swine ( Musick et al., 2010 ) , dogs ( Hart , 1974 ) , and monkeys ( Martin and Hart, 1974 ) . O ne 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, n one of the subsequent subchronic or chronic animal studies report ed histological findings of the gastrointestinal t r act 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 abnormalit ies were observed in a

case

- c ontrol study of males (32 exposed, 322 controls) exposed to RDX in an occupational setting ( W est and Stafford, 1997 ) (see Table 1

- 14) . T he 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 o f low intensity exposure, while the prevalence OR w as 1.2 (95% CI 0.3

− 5.3) for men with

> 50 hours of high intensity exposure

. T h e ORs from this study must be interpreted with caution given the small sample size and wide confidence intervals

. N o changes i n hematological parametersToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE (including hemoglobin, hematocrit, and reticulocyte count) were observed in a cross

- sectional 

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28 epidemiologic study of 69 workers exposed to RDX by inhalation (average of 0.28 mg/m 3 ) ( Hathaway and Buck, 1977 ) . H umans 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, hemat u ria, 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 individual s ( 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 B6C3F 1 mice ( see Table 1-15)

. I ncreases
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 ( Li sh et al., 1984 ; Levine et al., 1983 ; Cholakis et al., 1980 ); h owever, findings were generally inconsistent across studies and were not necessarily dose

- dependent . Similarly, decreased hemoglobin levels/anemia were observed in
some chronic and subchronic st u dies ( Levine et al., 1983 ; Cholakis e t 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 sh ow a consistent relationship with dose . Other hematological parameters, including W B C counts, reticulocyte counts, and hematocrit, showed conflicting results between studies, marginal responses, or inconsistent changes with increasing dose . O t h er subchro nic studies in rats and dogs ( Crouse et al., 2006 ; Hart, 1974 ; Von Oettingen et al., 1949 ) did not identify any changes in hematologica l parameters. In summary, evidence for hematological effects associated with RDX exposure in humans
comes from several case reports t h at found transient fluctuations in hematological endpoints after
acute exposures . H e m atological findings from two epidemiological studies were inconsistent and difficult to interpret because of small sample sizes ( T able 1 -14)

. I n general , animal studies of chronic and subchronic durations showed no consistent, dose

- r elat ed pattern of increase or decrease in hematological p ar am eters

.

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

Reference and s tudy d esign Results
Hematological e ffects
West and Stafford (1997) (United Kingdom) Case

- control study, 32 cases with abnormal and 322 controls with normal hematology test drawn from 199 1 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

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— DO NOT CITE OR QUOTE Reference and s tudy d esign Results
Hematological e ffects
cases, 93% of controls

. Analysis limited to men (29 cases, 282 controls). Exposure measure s: Exposure determination based on employee interviews and job title analysis; data included frequency ( hrs /d, d/ yr ), duration ( yr s), 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 neutro penia (2.0 × 10 9 /L ), low platelet count (<150 × 10 9 /L ), or macrocytosi s (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/m 3 (mean 0.28 mg/m 3 ). Effect measures : Hematology tests. Analysis : Types of statistical tests were not reported (assumed to be t

- tests for comparison of mea ns and χ 2 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/m 3 (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/m 3
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.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **Table 1-**  **15. Evidence**  **pertaining to systemic**  **effects in animals** 

Reference and s tudy d esign Results
Ocular effects
Lish et al. (19 84) ; Levine et al. (1984) Mice, B6C3F 1 , 85/sex/group; interim sacrif ices (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. (1 980) 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/kg

d animals).

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 35 175/100Toxicological Review of Hexahydro

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Mice, B6C3F 1 , 85/sex/group; interim sacrif ices (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 Die t 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, F 344, 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, B6C3F 1 , 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

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— DO NOT CITE OR QUOTE Reference and s tudy d esign 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 (mi croscopic 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

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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

− − 1/3Toxicological Review of Hexahydro

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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, B6C3F 1 , 85/sex/group; interim sacrif ices (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

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— DO NOT CITE OR QUOTE Reference and s tudy d esign 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
conta minant ; 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, B6C3F 1 , 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

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- 1,3,5

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— DO NOT CITE OR QUOTE Reference and s tudy d esign Results

88.6% pure
, with 9% HMX a nd 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 conta minants; ~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

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- 1,3,5

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— DO NOT CITE OR QUOTE Reference and s tudy d esign 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; p article 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 te st 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 effectsToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Reference and s tudy d esign Results
Lish et al. (1984) ; Levine et al. (1984) Mice, B6C3F 1 , 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/g roup) 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; p article 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
g roups 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

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- 1,3,5

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Hematological e ffects
Lish et al. (1984) ; Levine et al. (1984) Mice, B6C3F 1 , 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

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— DO NOT CITE OR QUOTE Reference and s tudy d esign 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, B6C3F 1 , 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%Toxicological Review of Hexahydro

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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; p article si ze not specified Doses 0 15 25 50

RBC count (percent change compared to control)

M + F 0%

− 23%

− 12%

− 14%Toxicological Review of Hexahydro

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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 d s 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% 

### 1 2 3 4 *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.Toxicological Review of Hexahydro

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This document is a draft for review purposes only and does not constitute Agency policy. 1- 67 DRAFT

— DO NOT CITE OR QUOTE c Levine et al. (1981a) is a laboratory report of a 13

- week study of RDX in F344 rats; two subsequently published 

### 1 2 3 4 papers ( Levine et al., 1990 ; L evine et al., 1981b ) present su bsets of the data provided in the full laboratory report.

d 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.Toxicological Review of Hexahydro

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systems h ave been reported in some studies. EPA concluded that the evidence does not support

t hese effe cts as a potential human hazard of RDX exposure.



## 1.2. **INTEGRATION AND EVALUATION**

**1.2.1. Effects**  **Other Than Cancer**

T he majority of evidence for the health effects of RDX comes from oral toxicity studies

. T he available health effects literature does not support identification of hazards by the inhalation route of exposure

. T hree e pidemiological studies that document possible inhalation exposure are limited by v arious study d esign features, including inability to distinguish exposure to TNT ( associated with liver and hematological system toxicity )

, u n certainty in i dentifying exposure levels, small sample sizes, and inadequate reporting. T h e s ingle animal inhalation study identified in the
literature search h a d deficiencies that precluded its inclusion in this assessment ( s ee Literature Search Strategy | Study Selec t io n a nd Evaluation ). The strongest evidence for h azar d s following exposure to RDX is for ner vous system effects

. A h uman 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 Append ix C.3 )

. O ther nervous system effects identified in human case reports include dizziness, headache, confusion, and hyperirritability

. E v idence from toxicity studies in multiple animal species involving chronic, subchronic and gestational exposures is co n sistent with the effects seen in humans . E ff ects included d ose

- r elated increases in seizures and convulsions, as well as observations of tremors, hyperirritability, hyper

- r e activity, 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 )

. I n 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 e ffects, 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

. T h e s trength of a direct association between mortality and ne rvous system effect s i s less clear in
most of the earlier studies because t h e 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 w ith duration of exposure

. I t is unclear if nervous system effects increased in severity (e.g., from behavioral change to seizures and convulsions) with increasing dose because many of th e studies that reported more subtle neurobehavioral changes did not p r ovide detailed dose

- response information

, a nd the majority of studies were not designed to capture this information . A dditionalToxicological Review of Hexahydro

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37 consistent evidence of neurotoxicity a cross taxa, including several species of wildlife ( Quinn et al.,
2013 ; Garcia

- Reyero et al., 2011 ; McFarland et al., 2009 ; Gogal et al., 2003 )

. A lthough 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 )

. A dditionally, the affinity of RDX for the picrotoxin convulsant site of the GABA A 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 i s more limited than evidence for neurotoxicity

. I n creased 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 )

. S imilar histopathological changes of the urogenital system were not observed in mice, and no other rat
studies of similar duration tha t ex amined the prostate were available . Among the lesions identified in the rat, t he incidence of suppurative pro st atitis is considered a marker for RDX -related urogenital e f fects . The plausibility of a MOA that shares a common molecular initiating event (binding to the GABA A receptor convulsant

- site) with the neurotoxic effects of RDX increases support for an association between RDX exposure and kidney and other urogenital effects

. E PA
identified t he urogenital system a s a potential human hazard of RDX exposure. Evidence for male reproductive toxicity comes from the finding of testicu lar degeneration in
male B6C3F 1 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. H o w ever, studies in different rat strains did not consistently repo rt te s ticular effects . Although the available data are limited, given the dose

- related findings of mouse testicular degeneration

, E PA i dentified suggestive evidence of male reproductive
effects as a potential h um an hazard of RDX exposure

. Evidence for developmental toxicity and liver toxicity was more limited than that for the endpoints discussed above . I n a n im al studies

, embryotoxicity and other developmental effects
were observed only at d oses associated with maternal mortality ( Angerhofer et al., 1986 ; Cholakis et al., 19 80 ) . Evidence for h epat ic effects comes from observations of increases (generally dose

related) 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 )

. H owever, these elevations in liver weight were not accompanied by RDX

- related histopathological change s in the liver or increases in serum liver enzymes

. I n 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 liverToxicological Review of Hexahydro

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- response analysis and the derivation of reference values

.

**1.2.2. C**  **arcinogenicity**

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 trend s f o r hepatocellular adenomas or carcinomas and alveolar/bronchiolar adenomas
or carcinomas in female

, but not male, B6C3F 1 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, t here

was n o evidence of carcinogenicity in Sprague

- Dawley r ats in a 2- year dietary stud y of RDX ( Hart,
1976 )

. N o human studies are available to assess the carcinogenic potential of RDX

. EPA’s Guidelines for Carcinogen Risk Assessment ( U.S. EPA, 2005a ) e mphasizes the importance of weighing the totality of evidence in reaching conclusions about the human carcinogenic potential of agents under evaluation

. I nf ormation taken into consideration in weighing the evidence for the human carcinogenic potential of RDX includes the m ag nitude of response in rats, availability of a PWG reevaluation of tumors, and potential differences in test
material a cross 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. H o wever, as discussed in Section 1.1.5, the
association of live r t umors in rats with RDX exposure was judged not to be strong for several

reasons , including the small numbers of carcinomas o bserved across the study and the low surviva l rate in the high

- d os e group that reduces confidence that the final incidence in th at 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 th e t wo

- year bioassay in t he Sprague

- Dawley rat ( Hart, 1976 ) . T he
incidence of liver tumors in the Hart (1976) study was not increased relative to contr ols at a dose of 10 mg/kg

- day, a dose that fell in the range of doses in the Levine et al. (1983) s tudy 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 )

. T he available evidence for RDX suggests that it could be considered a borderline case between two descriptors

— “li kely to be carcinogenic to humans” and “suggestive evidence of carcinogenic potential.” O n e of the criteria identified in EPA ’s Guidelines for Carcinogen Risk Assessment ( U.S. EPA, 2005a ) that support s the likely descriptor is “an agent that has tested positive in animal experiments in more than one species, se x, strain, site, or exposure route, with or without evidence of carcinogenicity in humans” may qualify as a likely carcinogen . Bi o assay data provide evidence for RDX carcinogenicity in one sex of one species (female mouse)

, w e aker evidence for carcinogeni city in one sex of a second species (male rat), andToxicological Review of Hexahydro

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37 criteria for the “likely to be carcinogenic to humans” descriptor. The “suggestive evidence of carcinogen i c 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. T h is descriptor

c o vers a wide spectrum of evidence associated with varying levels of concern for carcinogenicity,

in clud in g a positive cancer result in the only study on an agent, a single positive cancer result in an

ex tensive database that includes negative studies in ot h er 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 f o r increased trends in tumor incidence in two tissues and possibly a second species r a ises a concern for carcinogenic effects in humans

. H owever, 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 i n male B6C3F 1 mice , female F344 rats, and Sprague

- Dawley

rats of b oth sexes, E PA concluded that there is “suggestive evidence of carcinogenic potential” for
RDX.
U.S. EPA’s Guidelines for Carcinogen Risk As s es sment ( U.S. EPA, 2005a ) indicate that for

tumor s 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

. A n e xception occurs when there is convincing tox icokinetic data that absorption does not occur by other routes

. I n formation available on the carcinogenic effects of RDX via the
oral r oute demonstrates that tumors occur in tissues remote from the site of absorption

. I n formation on the carcinogenic effe ct s of RDX via the inhalation and dermal routes in humans or
animals is not available

. B as ed on the observ ation 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. T h er efore, there is “suggestive evidence of carcinogenic potential” following exposure to RDX by a l l routes of exposure.

**1.2.3. Susceptible Populations and Lifestages**  **for Cancer and Noncancer Outcomes** Susceptib il ity refers to factors such as lifestage, genetics, sex, and health status that may predispose a group of individuals to greater response to an expos ur e

. T his greater response could be achieved either through differences in exposure to the chemical underlying toxicokinetic and toxicodynamic differences between susceptible and other populations. L it tle information is available on populations that may be especially vulnerable to the toxic effects of RDX

. L i festage, and in particular childhood susceptibility, has not been observed in human or animal studies of RDX toxicity

. R eproductive and developmental toxicity studies did not identif y effects in off spring

at d oses b elow those that also caused maternal toxicity ( Angerhofer et al., 1986 ; Cholakis et al.,
1980 )

. T ransfer of RDX from dam to the fetus du ring 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

- RuthToxicological Review of Hexahydro

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. B ecause RDX is neurotoxic in adult animals, eviden ce of gestational transfer of RDX to 

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have not been conducted

. L i m ited data suggest that male laboratory animals may be more susceptible to noncancer toxicity associated with RDX exposure

. W h ile no sex

- based differences in neurotoxicity were observed, urogenital effects have been observed in males at lower doses tha n in
females ( Levine et al., 1983 ; L evine et al., 1981a ; Levine et al., 1981b ; Cholakis et al., 1980 )

, suggesting a possible sex

- based difference in susceptibility to RDX toxicity

. T here 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

. T his susceptibility may also be influenced by differential expression of these enzymes during development. I n d ividuals with epilepsy or other seizure syndromes , and in particular those th a t have their basis in genetic mutation to GABA A receptors, may represent another group that may be susceptible to RDX exposure. H o wever, there is currently no information to support predictions of how genetic polymorphisms may affect susceptibility.Toxicological Review of Hexahydro

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# 2. DOSE

- RESPONSE ANALYSIS

**2.1. ORAL REFERENCE DOSE FOR EFFECTS OTHER THAN CANCER** 

The RfD (expressed in units of mg/kg

- day) is defined as an estimate (with uncertainty 

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lifetime

. I t can be derived from a no

- observed -adverse- effect level (NOAEL), lowest

- observed

adverse- effect level (LOAEL), or the 95% lower bound on the benchmark dose (BMDL), with uncertainty factors (UFs) generally applied to reflect limitations of the data used.

**2.1.1. I**  **dentification 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 av ailable epidemiological studies of worker populations exposed to RDX examined the relationship between certain h ealt h endpoints and inhalation exposure; no epidemiological studies of ingested RDX are available

. Th erefore, epidemiological studies could not be used for oral dose

- r es ponse analysis and as the basis for the RfD. M ultiple case reports provide some evidence of effects in humans associated with acute exposure to RDX; however, while c as e

r eports can support the identification of hazards associated with RDX exposure, data from case reports are inadequate for dose

- r e sponse analysis and subsequent derivation of a chronic reference value because of short exposure durations and incomplete o r missing quantitative exposure
information

. As discussed in Section 1.2.1, b a sed on findings from oral studies in experimental animals,
EPA identified nervous system effe c ts as a human hazard of RDX exposure

, and effects on the

u r ogenital system ( including the kidney) as a potential human hazard of RDX exposure. EPA also identified suggestive evidence of m a le reproducti ve effects as a potential human hazard of RDX

exposure . E x perimental animal studies within each health effect category were eval uated 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. R a t ionales for selecting the studies and effects to represent each of these hazards are summarized below.

Nervous S y

stem E ffe cts Nervous system effects following oral exposure to RDX, including convulsions, seizures, and hyper

- r eactivity, were observed in m u ltiple studies in rats, mice, monkeys, and dogs . O nly threeToxicological Review of Hexahydro

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— Crouse et al. (2006)

, Cholakis 

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— 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 f ollowing a 90

- day oral (gavage) exposu re to RDX . A dditionally, 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), s i x dose groups (including the control) that provided good resolution of
the dose

- r esponse 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 GD s 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

- r esponse analysis. The study in monkeys by Martin and Hart (1974) was not selected for dose

- response analysis. This stud y provided supporting evidence of nervous system effects (trembling, shaking,
ataxia
, an d hyperactive reflexes) with 66% incidence at the high dose of 10 mg/kg -day ; however, this study was not selected for d ose

- response analysis b ecause it used small group sizes (n = 3 / sex) and the exposures were relatively variable or uncertain (e .g., p urity of the test material was not specified , an d reported emesis in some animals likely influenced the amount of dose received ).
Other chronic and subchronic studies report ed nervous system effects as clinical
observations ( A ngerhofer et al., 1986 ; Lish et al., 1984 ; Levine et al., 1983 ; Levine et al., 1981a ; Le vine et al., 1981b ; Von Oettingen et al., 1949 ), but without incidence data

. A s discussed in Section 1.1.1, these studies did not systematically monitor or evaluate nervous system effects induced by
RDX
, l eading to possible underestimates of incid ence of such effects . A s such, there is some uncertainty associated with identification of NOAELs and LOAELs for nervous system effects from
these studies

. Further, t hese 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 O ther U rogenital E ffects 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 . A s 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

- ye ar study in F344 rats by L evine et al. (1983) and in a 13

- week study in B6C3F 1 mice by Cholakis et al. (1980)

. H istopathological changes of the kidney and bladder (medullary pa p illary necrosis, sup purative pyelitis, uremic mineralization, and luminal distention and cystitis of the urinary bladder) wereToxicological Review of Hexahydro

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- dose (40 mg/kg

- day) males . The incidence of suppurati ve 

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s howed a dose

- relat ed trend beginning at doses below 40 mg/kg

- day (see Section 1. 1.3 )

. T herefore, suppurative prostatitis was selected for dose

- r e spons e 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

- res p onse analys is.

Male R eproductive T oxicity Male reproductive effects were identified in mice following chronic administration of RDX

in the d iet. Lish et al. (1984) observed an increased incidence of testicular degeneration in mice given RDX in diet for two years compared t o controls

. T he response was shown to be dose -related
and was selected for do s e

- response modeling . C hanges in other reproductive outcomes were not
dose- related or consistently observed across studies, and therefore were not considered for dose

r e sponse 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. T he g eneral approach for the estimation of PODs is presented in Figure 2 -1 a n d described further below.

Human external d ose (HED)
Benchmark dose modeling Animal PBPK model Human PBPK model Animal
extern al d ose BMDL or NOAEL Animal internal d ose

**F ig**  **ur e**  **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

- re sp onse models thought to be consistent with underlying biological processes to determine how best t o e m pirically model the dose

- response relationship in the range
of the observed data. C on s istent 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). E PA

es t imated the benchmark dose (BMD) and 95% lower confidence limit on the BMD (BMDL) using a benchmark response (BMR) sele c ted for each effect

. A summary of BMD modeling, includingToxicological Review of Hexahydro

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Incidence d ata from Crouse et al. (2006) and Cholakis et al. (1980) were amenable to modeling

. F or 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 incid en ce data from male and female rats were used for modeling convulsion data from this study . A B MR of 1 % extra risk for convulsions was used to address the relative severity of this endpoint; a cr oss the expe rimental animal database for RDX , convulsions and seizures were generally associated with m ortality . I n 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 appre cia ble risk

. L ess severe nervous system outcomes that precede
convulsions and associated mortality would be preferred, b ut none were identified for RDX

.

K

idney/ U rogenital and M ale R eproductive E ffects 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 minimal l y biological ly significant degree of effect. U ncertainty in this characterization should be taken into account in comparisons with PODs from other effects.

Human E xtrapolation

EPA guidance ( U.S. EPA, 2011 ) advocates a hierarchy of approaches for deriving h uman equivalent doses (HEDs) from data in laboratory animals, with the preferred approach being physiologically- b ased toxicokinetic modeling . O ther approaches can include using chem ical

specific information in the absence of a complete physiologically

- b ased toxicokinetic model . I n lieu
of either reliable ch emical

- specific models or data to inform the derivation of human equivalent oral exposures, a body weight scaling to the 3⁄4 power ( i.e., BW 3/4 ) 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 fo r 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.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **Table 2-1** 

**. Summary of derivation of**  **PODs following oral exposure to RDX**

Endpoint and

r eference (exposure duration/route) Species/sex Model a BMR BMD (mg/kg

- d) BMDL (mg/kg

- d) POD HED (mg/kg

- d)

Administered dose b RDX AUC c

N ervous 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) ( GD s 6

− 19/gavage ) Female F344
rat Quantal

linear 1 % ER 0. 18 0. 12 0. 03 0. 06

K idney/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

M ale reproductive system

Testicular degeneration Lish et al. (1984) (2

- yr /diet) Male B6C3F 1
mouse LogProbit 10% ER 56.0 16.3 2.4 0.08



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b POD was convered to an HED using a standard DAF based on BW 3/4

.
c POD was converted to an HED based on the equivalence of internal RDX dose (expressed as area under the curve

[AUC] f or RDX concentration in arterial blood ) derived using PBPK models.

ER = extra risk

Physiologically- based pharmacokinetic models for RDX in ra ts, humans

, and mice have been published ( S weeney et al., 2012a ; Sweeney et al., 20 12b ; Krishnan et al., 2009 ) based on RDX

specific data . E PA evaluated and further developed these models for extrapolating doses from
animals to h umans (see Appendix C , Section C.2.5 )

. A s 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 effect s w ith 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 )

.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE T h erefore, PBPK model

- derived arterial blood concentration of RDX is considered a better dose



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31 metric for extrapolation of health effects than administered dose when there is adequate
confidence in the est imated value . T he PBPK model s for RDX w ere used to estimate the area under the curve (AUC) for RDX concentration in arterial blood , which represents the average blood RDX concentration for the exp osure duration normalized to 24 hours

. It appears logical to use RDX concentration levels in the brain as the internal dose metric for analyzin g convulsions as the health effect . N evertheless, the blood concentration of RDX was preferred as the dose metric due to greater confidence in modeling this variable. Th is 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 concentratio n s because, as discussed in Appendix C, Section C.2.5 , the

rate co ns tant for oral absorption (KAS) is uncertain , and peak concentrations are more sensitive to variations in this parameter than average values. F urthermore, a more consistent dose

- response for c onvulsions is observed in chronic studies than for the higher exposures in subchronic studies.
The rodent PBPK model was ap plied to the BMDL s generated from BMD modeling to
determine the animal internal dos e , expressed as the AUC of RDX blood concentration, and representing the cross

- sp ecies toxicologically eq uivalent dose. The human PBPK model was then applied t o derive the corresponding HEDs (see Figure 2

- 1) . B ecause the AUC is linear with exposure level, at least in the exposure range of interest, t he value of the HED would be the same
whether the rat or m o use 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 c o nsistent 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 = (BW a 1/4 /BW h 1/4 ),
where
BW a = animal body weight
BW h = human body weight

Using a BW a of 0.25 kg for rats and 0.035 kg for mice and a BW h of 70 kg for humans ( U.S. EPA, 1988 ) , the resulting DAFs for rats and mice are 0.24 and 0.15, respectively

. A pplying the DAF
to the POD identified for effe cts in adult rats or mice yields a POD HED as follows (see Table 2

- 1):

POD HED = Laboratory animal dose (mg/kg

- day) × DAFToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE F urther details of the BMDL modeling, BMDS outputs, and graphical results for the best fit 

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37 model for each dataset included in Table 2 -1 can be found in Appendix D, Section D.1. D etails of the PBPK model evaluation used for extrapolation from BMDL values can be found in Appendix C,
Section C.2.5. T able 2- 1 summarizes the results of the BMD modeling and the POD HED for each data
set discussed above.

**2.1.3. Derivation of Candidate**  **Values** Pharmacokinetic models are useful to examine species differences in pharmacokinetic processing . Bec ause of relatively high confidence in the rat and human PBPK modeling, these
models were used to derive reliable in ternal dose metrics for extrapolation . F or datasets selected from the rat bioassays, t he candidate RfDs were calculated assuming cross

- species toxicological equivalence of the AUC of RDX blood concentration derived from the PBPK mo del ing . H owever, there were major uncertainties identified in the mouse PBPK modeling. Th erefore, for endpoints selected from the mouse bioassay, the preferred approach for determining the candidate RfDs is
that based on the administered dose of RDX ex t rapolated to humans using allometric BW 3/4 scaling

. T h e evaluation of confidence in the PBPK model results is summarized in Summary of confidence in PBPK models for RDX i n Appendix C, Section C.2.5.
Under EPA’s A Review of the Reference Dose and Referen ce 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

. A n explanation follows

. An intraspecies uncert ainty factor

, UF H , 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 u ncertainty factor

, UF A , of 3 (10 1/2 = 3.16, rounded to 3) was applied to all
PODs t o account for uncertainty in characterizing the toxicokinetic and toxicodynamic differences
between rodents and humans. For the testicular degeneration dataset from the mouse bioassay, a
UF A of 3 was applied because BW 3/4 scaling is used to extrapolate oral doses from laboratory animals to humans. Although BW 3/4 scaling addresses some aspects of cross

- species extrapolation of toxicokinetic and toxicodynamic p rocesses, some residual uncertainty remains. In the absence of
chemical

- sp ecific data to quantify this uncertainty, EPA’s BW 3/4 guidance ( U.S. EPA, 2011 ) recommends use of an uncertainty f actor 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 r educes toxicokinetic uncertainty in extrapolating from the rat to humans, but does not account for interspecie s d ifferences due to toxicodynamics. A UF A 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, UF s , differs depending on t he exposure duration

. An UF s of 1 was applied to the POD values for kidney/urogenital effects and testicular degeneration
derived from the 2

- y ear bioassays in the rat ( Levine et al., 1983 ) and mouse ( Lish et al., 1984 ) . PODToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE v alues for nervous system effects were derived from st udies of subchronic duration or gestational 

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38 exposure; a UF S of 3 was applied to these PODs. Typically, a UF S 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 fr om 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 UF S 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 r odent bioassays provide s up port for an UF S no greater than 3. D ose levels associated with convulsions in chronic dietary studies of RDX a re ≥35 mg/kg -day and are higher than doses that induced convulsions in
the 14 - an d 90

- day ( gavage ) studies that were used to derive candidate P ODs for nervous system effects (i.e., 2 m g/kg

- day in Cholakis et al. (1 980) 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 UF S that is less than the default of 10.
A LOAEL to NOAEL uncertainty factor

, UF L , 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 a s one of the considerations in selecting a BMR f or benchmark dose modeling . In this case, the BMR for modeled endpoints w as selected under the assumption that the BMR represents a minimal, biologically significant change for these effects. A database uncertainty factor

, UF D , of 3 was applied to all POD values. T he oral toxicity
database for RDX i n cludes subchronic and chronic toxicity studies in the rat and mouse, a two

generation 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 monke y . D eficiencies in the database related to neurobehavioral and neurodevelopmental testing were identified. T he database for neurotoxicity is characterized primarily by observations of frank effects (convulsions)

. A d ditional observations of neurobehavior al 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, tre atment -related behavioral effects. F urther, 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 . G i ven 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
transfe r of RDX to offspring . Given the potential for exposure during gestation and lactation and the neurotoxic potential of RDX , t he lack of a developmental neurotoxicity study was identified as a data gap. A UF D of 3 was applied to all PODs to account for li mitations in neurobehavioral and neurodevelopmental testing

.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Table 2

- 2 is a continuation of Table 2

- 1 and summarizes the application of UFs to each 

### 1
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5 POD HED to derive a candidate value for each data set. T he candidate values presented in the table below are preliminary to the derivation of the organ/system

- specific reference values. T hese 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 POD H ED a POD type UF A UF H UF L UF S UF D Composite UF Candidate value (mg/kg

- d)

N ervous system (rats)

Convulsions C rouse et al. (2006) 0.27 BMDL 01 3 10 1 3 3 300 8

. 8 × 10 − 4

Convulsions Cholakis et al. (1980) 0. 06 BMDL 01 3 10 1 3 3 300 2

. 0 × 10 − 4

K idney/urogenital system (rats)
Prostate suppurative inflammation Levine et al. (1983) 0.23 BMDL 10 3 10 1 1 3 100 2.3 × 10 − 3

M ale reproductive system (mice)
Testicular degeneration Lish et al. (1984) 2.4 BMDL 10 3 10 1 1 3 100 2.5 × 10 − 2



### 6 7 8 9

10
11 a POD HED values based on data from the rat were derived using PBPK modeling; the HED POD based on data from the mouse was derived using BW 3/4 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 POD HED s, with each bar corresponding to one data set described in Tables 2

- 1 an d 2 -2.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **Figure 2-2** 

**. Candidate values with corresponding POD and composite UF.** Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **2.1.4. Derivation of Organ/System**  **-S pecific Reference Doses** 

Table 2- 3 distills the candidate values from Table 2- 2 into a single value for each organ or 

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. O rgan - 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 fo**  **r RDX** 

Effect Basis RfD (mg/kg

- day) Study e xposure 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 S ystem E ffects The organ/system

- specific RfD for nervous system effects was based on the incidence of convulsions in rats reported in C ro u se 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

m ultiple other studies . A lt hough the candidate value derived from Ch olakis 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 v er s us 14 days) , more systematic evaluation of neurobehavioral endpoints , and higher test compound purity

.
Kidney/ U rogenital E ffects A single data set for incidence of suppurative prostatitis in male B6C3F 1 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 uroge nital system . A s previously discussed, the data supporting RDX

- relat ed kidney and other urogenital effects are largely limited
to this 2

- year study in the mouse . A c cordingly, the candidate value for kidney and other urogenital
effects is based on the incid e nce of suppurative prostatitis in male mice ( Lish et al., 1984 ).

Male R eproductive E ffects A single dataset for male reproductive effects was b rought forward for quantitative analysis: the incidence of testicular degeneration as r ep orted in male B6C3F 1 mice exposed to RDXToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE in diet for 24 months ( Lish et al., 1984 )

. T he candidate value for male reproductive effects i s based 

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**2.1.5. S**  **election 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

, k i dney and other urogenital effects, and male re productive eff ec ts . E vidence for nervous system effects, and specifically convulsions, was observed in multiple studies, in multiple species, and following a range of exposure durations . I n addition, the organ/system

- specific RfD for nervous system effects was the lowest among the organ/system

- s p ecific RfDs derived for RDX. E vidence for dose -related effects on the urogenital system co me s primarily from a single 2- year toxicity study in male rats ( L evine 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 C r ouse et al. (2006) i s 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/system

specific RfD . T h i s overall RfD should provide an exposure level below which effects associated with RDX exposure ar e n ot expected to occur

. The overall RfD is derived to be protective of all types of effects for a given duratio n of exposure, and is intended to protect the population as a whole, including potentially susceptible subgroups ( U .S . EPA, 2002 ) . D ecisions concerning averaging exposures over time for comparison
with the RfD should consider the types of toxicological effects and specific lifestages of concern

. F lu ctuations in exposure levels that result in elevated exposures during these lifestages could

p otentially lead to an appreciable risk, even if average levels over the full exposure durati o n were less than or equal to the RfD. In t he 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

. T o derive the RfD, the UF approach ( U .S . EPA, 2000a
, 1994 ) was applied to a POD HED based on nervous system effects in rats expos ed to RDX for a subchronic duration

. U Fs were applied to the POD HED s to

accou n t for u ncertainties 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. T h ese 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 dif f e red inToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE purity, formulation, and particle size. T here is evidence that differences in test material 

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strain s ; however, there is some uncertainty associated with the incidence of reported neurological

e ffects in studies that employed a study design that did not monitor animals with sufficient frequency to accurately record neurobehavioral effects, including conv ul s ions.

**2.1.7. C**  **onfidence Statement**

A confidence level of high, medium, or low is assigned to the study used to derive the Rf D
, the overall database, and the Rf D it se lf, as described in Section 4.3.9.2 of EPA’s Methods for Derivation of Inhalation Reference Conc e n trations and Application of Inhalation Dosimetry ( U.S. EPA,
1994 ). The overall confidence in this RfD is medium. C onfidence in the principal study ( Crouse et al., 2006 ) is high

. T he study was we ll- conducted, utilized 99.99% pure RDX, and had five closely

spaced 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
neurobeh av i oral effects at doses ≥8 mg/kg

- day due to the timing of the dosing procedure and timing of the FOB screening. C o n fidence in the database is medium. T he database includes three chronic studies in rats and mice; eight subchronic studies in rats, mice, d o g s, and monkeys; two
short- t erm studies; and four reproductive/developmental toxicity studies in rats and rabbits ( including a two

- g e neration reproductive study)

. C onfidence is reduced largely because of limited examination of the potential for RDX to ind u ce neurobehavioral and neurodevelopmental effects
and the incomplete u nd erstanding of a MOA for convulsions

. R eflecting high confidence in the principal study and medium confidence in the database, overall co n fidence 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 L ev ine et al. (1983) . T he 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 . A n R fD of 3 × 10

− 3 mg/kg

- day was derived following application of an overall UF of 100 (UF A = 10, UF H = 10).



## 2.2. **INHALATION REFERENCE CONCENTRATION FOR EFFECTS OTHER THAN CANCER** 

The RfC (expressed in units of mg/m 3 ) 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

. I t 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 u se d.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE As noted in Section 2.1, human studies are generally preferred over animal studies as the 

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studies of RDX ( We st and Stafford, 1997 ; Ma and Li, 1992 ; Hathaway and Buck, 1977 ) , none provi ded data that could be used for dose

- response analysis

. T he studies by Ma and Li (1992) of
neurobehavioral effects in Chinese workers and West and Stafford (199 7) 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).
T h e 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 . M ultiple 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

- r e sponse anal ysis 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 experimen t al animal study involving inhalation exposure was identified in the DTIC database; the

s tudy is not publicl y a vailable . H owever, the study would not have provided useful data on responses to inhaled RDX, as the study was limited by small numbers of anima l s tested, a lack of controls, and incomplete reporting of exposure levels. T h erefore, the available health effects literature does not support the derivation of an RfC for RDX. F u rther, a PBPK model for inhaled RDX is not available to support route

- to -r o u te extrapolation from the RfD.

**2.2.1. P**  **revious IRIS Assessment**

A n 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 h azard potential of the substance in question

, an d quantitative estimates of risk from oral and inhalation exposure may be derived. Q uantitative risk estimates may be derived from the application of a low

- dose extrapolation procedure

. I f d erived, the oral slope factor is a plausible upper bound on the estimate of risk per mg/kg

- d ay of oral exposure.

**2.3.1. A**  **nalysis of Carcinogenicity Data**

A s noted in Section 1.2.2, EPA concluded that there is “suggestive evidence of carcinogenic potential” for RDX

. T h e Guidelines for Carcinogen Risk Assessment ( U.S. EPA, 2005a ) state:Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE When there is suggestive evidence, the Agency generally would not attempt a dose

response assessment, as the nature of the data generally would not support one; however, when the evidence includes a well

- conducted study, quantitative anal yses may be useful for some purposes, for example, providing a sense of the magnitude and uncertainty of potential risks, ranking potential hazards, or setting research priorities

.

In th e case of RDX , the carcinogenicity of the chemical has been evaluat ed in one oral 

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1983 ; Hart, 1976 ) . The data in Lish et al. (1984) demonstrated a statistically significant positive
trend with dose 6 in the incidence of liver and lung tumors in female , but not male, B6C3F 1 mice associated with dietary administration of RDX . I n t h e study by Levine et al. (1983) , the incidence of liver tumors in male F344 rats showed a statistically significant positive trend with dose 7

. N o increases in tumors were observed in Sprague

- D aw ley rats exposed to RDX ( Hart, 1976 )

. As discussed further below, the 2

- year studies by Lish et a l. (1984) and Levine et al. (1983) were well

conducted studies that support quantitative analysis. C onsidering the se data a long with the uncertainty associated with the suggestive nature of the weigh t 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.
T he incidence s of liver and lung tumors in female mice from the study by Lish et al. (1984) were selected for quantitative dose

- response analysis . T he study by Lish et al. (1984) was performed in accordance with FDA Good Laboratory Practice r egulations ( FDA, 1979 ) , included comprehensive histopathological examinati on of major organs, contained four dose groups and a control, used adequate numbers of animals per dose group ( 6 5 /sex/group , plus interim sacrifice groups of 10/sex/group at 6 and 12 months) an d a sufficient overall exposure duration (2 years), and adequat e l y 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 esta blished by Harada et al. (1999) . T he 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 l iver and lung tumors , benign and malignant tumors (i.e., adenomas and carcinomas) w er e combined for do se

- 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 th e malignant stage ( U .S . EPA, 2005a ; McConnell et al., 1986 ).
Female mouse l iver and lung tumor incidence s from the Lish et al. (1984) study are summarized in Table 2 -4.

6 A two

- sided asymptotic Cochran

- Armitage test yielded p = 0.041 for liver tumors and p = 0.019 for lung tumors in female mice.

7 A 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.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **Table 2-4**  **. Incidence of hepatocellular and alveolar/bronchiolar tumors**  **in female B6C3F**  **1 mice administered RDX**  **for 2 years in diet** 

Tumor type Study/ Analysis Dose group ( mg/kg

- day )

Control 1.5 7 35 107 a
Hepatocellular adenomas or carcinomas Parker et al. (2006) 1/67 4/62 5/63 10/64 4/31 b

Alveolar/bronchiolar adenomas or carcinomas Lish et al. (1984) 7/65 3/62 8/64 12/64 7/31 b



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27 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

- da y.

The incidence of liver carcinomas in male F344 rats from the study by Levine et al. (1983)
was also consider ed for quantitative dose

- response analysis

. Although the study was well conducted (see S ect i on 1.1.5)

, EPA considered that the association between RDX exposure and rat liver tumors is not strong

, re f lect ing the relatively low magnitude of the rat liver carcinoma response and reduced confidence that the high

- do s e group accurately reflects lifetime cancer incidence because, in part , of l ow 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 o f t he carcinogen and the shape of the cancer dose

- response curve

. T he linear approach is recommended when there are MOA d at a to indicate that the dose

- response curve is expected to have a linear component below the POD or w hen t he weight of evidence evaluation of
all available data are insufficient to establish the MOA for a tumor site ( U.S. EPA, 2005a )

. I n the case of RDX, the mode of carcinogenic action for hepatocellula r and alveolar/bronchiolar tumors is
unknown

. T her efo re, 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 wh ether time of death should be incorporated in the dose

- re s ponse analysis of tumors

. F or 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 w as modeled using the multistage

- cancer models in BMDS (versions 2.4
and 2.5) . A s t andard BMR of 10% extra risk was applied to both tumor sites in the mouse

.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE G iven the finding of an association between RDX exposure in the female mouse and 

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37 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. T he MS

- COMBO procedure (BMDS, version 2.5)

, extends the multistage

- cancer models to the case with multiple tumors assuming independence between tumor types

. T here is no known biological relationship between liver and lung tumors in RDX

- exposed mice , and therefore, as noted by the National
Research Council ( NR C, 1994 ) , t his 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 . Th e p rocedure derives a maximum likelihood estimate of the combined risk at a 95% confidence lev el b ased on the parameter values obtained for the individual tumor multistage model fits.
E PA’s preferred approach for extrapolating results from animal studies to humans is toxicokinetic modeling

. A s d escribed in Appendix C, PBPK models for RDX in mic e and humans published by S w eeney et al. (2012b) were evaluated and further developed by EPA. C onsideration was given to whether the available toxicokin etic information supported using an internal dose metric derived by PBPK modeling

. T h e available mechanistic data (Section 1.1. 5 ) point to some evidence, although not conclusive, that RDX

- g en erated metabolites may be implicated in the observed tumorigenic i t y in the female mouse. H owever, 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

. I n this case, as is to be expected from mass balance principles, the PBPK modeling provides

n o further information; t h e HED obtained from the model

- estimated amount of total RDX
metabolites s c aled by BW 3⁄4 was equal to that calculated using administered dose scaled by BW 3/4

. In addition to the lack of data on metabolism, o t her major uncertainties were identified in the

mouse PB PK m odel ing; EPA’s evaluation of these u ncertainties is summarized briefly in Section 2.1.3 and in more detail in Appendix C, Section C.2.5

. T h erefore, the PBPK model developed for the

mouse was no t u sed, 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 a llometric scaling to achieve toxicological equivalence across species.
As discussed in Section 2.1.1, the administered dose in animals is converted to a n HED on the basis of (body weight) 3/4 ( U.S. EPA, 1992 )

. T his 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 )

. I t 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. D et ails of the BMD modeling can be found in Appendix D, Section D.2.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **2.3.3. Derivation of the Oral Slope Factor**  The lifetime oral cancer slope factor for humans is defined as the slope of the line from the 

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7 BMR (10% extra risk) at the BMD L to the estimated control response at zero (slope factor = 0.1/BMDL 10

- HED ) . This slope, a 95% upper confidence limit (UCL) on the true slope, represents a plausible upper bound on the true risk

. T he PODs estimated for each mouse tumor site are
summarized in Table 2 -5 . U sing linear extrapolation from the BMDL 10

- HED , human equivalent oral slope factors (OSFs) were derived for each tumor site individually and both sites comb ined and are
listed in Table 2 -5.

**Table 2**  **-5 . M**  **odel predictions and oral slope factors for hepatocellular and alveolar/bronchiolar adenomas or carcinomas in female B6C3F**  **1 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 = BMDL 10

- HED a (mg/kg

- d ) OSF b (mg/kg

- d) − 1

Hepatocellular adenomas or carcinomas c 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 10 11 12

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17 a BMDL 10

- HED = BMDL 10 × (BW a 1/4 /BW h 1/4 ), where BW a = 0.035 kg, and BW h = 70 kg

. b Slope factor = BMR /BMDL 10

- HED , where BMR = 0.1 (10% e xtra r isk)

.
c Incidence of female mouse liver tumors from Lish et al. (1984) ar e those reported in the PWG reevaluation ( Parker

et al., 2006 ).

An OSF was derived from the BMDL 10

- HED based on significantly increased incidence of hepatocellular and alveolar/bronchiolar adenomas or carcinomas in female B6C3F 1 mice (i.e., the Liver + Lung BMDL 10

- HED from MS

- COMBO)

. T he OSF of **0.0 4**  **(mg/kg**

**- day)** 

**− 1**  is calculated by dividing the BMR (10% extra risk) by the Liver + Lung BMDL 10

- HED and represents an upper bound on cancer risk associated with a continuous lifetime exposure :

OSF = 0.1 ÷ (Liver + Lung) BMDL 10

- HED

= 3.8 × 10

− 2 (mg/kg

- day)

− 1

= 4 × 10

− 2 (mg/kg

- day)

− 1 , rounded to one significant figureToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE **2.3.4. Uncertainties in the Derivation of the Oral Slop**  **e Factor**

A number of uncertainties underlie the cancer unit risk for RDX. T able 2 -6 summarizes the 

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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 . EP A, 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 i mpact on c ancer r isk v alue Decision Justification
Selection of study The cancer bioassay in the rat ( Levine et al., 1 983 ) 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 p rovided 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 wo uld 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 indic ating animal tumors are not relevant to humans ( U.S. EPA, 2005a )

. A s 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 li kely 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

a dministered dose Lack of sufficient data on RDX metabolism and major uncertainties identified in the mouse PBPK model.Toxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE Consideration and i mpact on c ancer r isk v alue Decision Justification

Cross

- species scaling Alternatives could ↓ or ↑ slope factor (e.g., 3.5

- fold ↓ [scaling by body wei ght] or ↑ 2

- fold [scaling by BW 2/3 ]) BW 3/4 scaling (default approach) There are no data to support alternatives

. B ecause the dose metric was not an area u nder the curve, BW 3/4 scaling was used to calculate equivalent cumulative exposures for estimating equivalent human risks

. W hile 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

. T he multistage model has biological support and is the model most consistently used in EPA
cancer assessments ( Geh lhaus 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 )

. B ecause 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. 

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8 **2.3.5. Previous IRIS Assessment: Oral Slope Factor**  The previous cancer assessment for RDX was posted to the IRIS database in 1990. T he oral slope factor in the previous cancer assessment was based on the bioassay by L i sh et al. (1984) and analysis of data for hepatocellular adenomas or carcinomas in female mice . A slope factor of
1.1 × 1 0

− 1 (mg /kg

- day)

− 1 was derived using a linearized multistage procedure (extra risk). T his differs from the slope factor for hepatocellular tumors in Table 2 -6
, because the current OSF is
bas e d on the combined incidence of hepatocellular and alveolar/bronchiolar adenomas or

ca rcinomas, PW G r eevaluation of female mouse liver tumors

, and use of scaling by body weight toToxicological Review of Hexahydro

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— DO NOT CITE OR QUOTE the 3/4 power for cross

- species extrapolation (whereas the previous assessment scaled by body 

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## 2.4. **INHALATION UNIT RISK FOR CANCER**
The carcinogenicity assessment provides information on the carcino genic hazard potential of the substance in question and quantitative estimates of risk from oral and inhalation exposure may be derived. Q uantitative risk estimates may be derived from the application of a low

- dose extrapolation procedure

. I f d erived, th e inhalation unit risk is a plausible upper bound on the estimate of risk per μg/m 3 air breathed.
An i nh alation u ni t r isk value was not calculated because inhalation carcinogenicity data for
RDX ar e not available. A P BPK model for inhaled RDX is not avail able 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 MO A . B ec ause no chemical

- specific data on li fe

- stage susceptibility for RDX carcinogenicity ar e av ailable, and because the MOA for RDX carcinogenicity is not known (see Section 1.1.5), A DA Fs were not applied.Toxicological Review of Hexahydro

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Angerhofer, R; Davis, G; Balezewski, L. (1986). Teratological assessment of Trinitro

- RDX in rats. (75

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ATSDR (Agency fo r Toxic Substances and Disease Registry). (1992). Health Assessment for Nebraska Army Ordnance Plant (Former), Mead, Saunders County, Nebraska, Region 7. CERCLIS No. NE6211890011. Atlanta, GA. <br />: Department H.E.W. Office of Toxic Substance and Disease

. ATSDR (Agency for Toxic Substances and Disease Registry). (1993). Health Assessment for Milan Army Ammunition Plant, Milan, Carroll and Gibson Counties, Te nnessee, Region 4. CERCLIS No. TN0210020582. September 30, 1993. Atlanta, GA. <br />: Department H.E.W. Office of Toxic Substance and Disease.

ATSDR (Agency for Toxic Substances and Disease Registry). (1996). Symptom and disease prevalence with biomarkers health study Cornhusker Army Ammunition Plant Hall County, Nebraska. (ATSDRHS9672). Atlanta, GA. Div. of Health Studies. <br />. http://www.ntis.gov/search/product.aspx?ABBR=PB96187760

ATSDR (Agency for Toxic Substances and Disease Registry). (1999). Public Health Assessment for Iowa Army Ammunition Plant, Middletown, Des Moines County, Iowa, Region 7. CERCLIS No. IA7213820445. Atlanta, GA. Div. of Health Assessment and Consultation. <br />: Department H.E.W. Office of Toxic Substance and Disea se.
ATSDR (Agency for Toxic Substances and Disease Registry). (2012). Toxicological Profile for RDX (Update) [ATSDR Tox Profile]. (NTIS/10590213). Research Triangle Park, NC.; Agency for Toxic Substances and Disease Registry, Atlanta, GA. <br />: U.S. Department of Health and Human Services, Public Health Service.

Ba iley, SA; Zidell, RH; Perry, RW. (2004). Relationships between organ weight and body/brain weight in the rat: What is the best analytical endpoint. Toxicol Pathol 32: 448

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Barsotti, M; Crotti, G. (1949). [Attacchi epileptici come manifestazione di intossicazione professionale da trimetilen

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— DO NOT CITE OR QUOTE Best, EP; Sprecher, SL; Larson, SL; Fredrickson, HL; Bader, DF. (1999a). Environmental behavior of explosives in groundwater from the Milan Army Ammunition Plant in aquatic and wetland plant treatments. Removal, mass balances and fate in groundwater of TNT and RDX. Chemosphere 38: 3383

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- 1351.2003 Bingham, E; Cohrssen, B; Powell, CH. (2001). Patty's toxicology (5 ed.). Ho boken, NJ: John Wiley & Sons. Bishop, RW; Kennedy, JL; Podolak, GE; Ryea, JL. (1988). A field evaluation of air sampling methods for TNT and RDX. Am Ind Hyg Assoc J 49: 635

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- VCH Verlag. http://dx.doi.org/10.1002/14356007.a10_143 Burdette, L; Cook, L; Dyer, R. (1988). Convulsant properties of cyclotrimethylenetrinitramine (RDX): Spontaneous, audiogenic, and amygdaloid kindled seizure activity. To xicol Appl Pharmacol 92: 436

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- 8027). Kansas City, MO: Midwest Research Institute. Cowin, PA; Gold, E; Aleksova, J; O'Bryan, MK; Foster, PMD; Scott, HS; Risbridger, GP. (2010). Vinclozolin exposure in utero induces postpubertal prostatitis and reduces sperm production via a reversible hormone

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- 0982 Crouse, LCB; Michie, MW; Major, M; Johnson, MS; Le e, RB; Paulus, HI. (2006). Subchronic oral toxicity of RDX in rats. (Toxicology Study No. 85

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- 03). Aberdeen Proving Ground, MD: U.S. Army Center for Health Promotion and Preventive Medicine. Crouse, LCB; Michie, MW; Major, MA; Leach, GJ; Reddy, G. (2008). Oral bioavailability of cyclotrimethylenetrinitramine (RDX) from contaminated site soils in rats. Int J Toxicol 27: 317

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- phase optimization for bioremediation of munition compound

- contaminated soils. Appl Environ Microbiol 59: 2171

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- 011

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- 3 Gehlhaus, MW; Gift, JS; Hogan, KA; Kopylev, L; Schlosser, PM; Kadry, AR. (2011). Approaches to cancer assessment in EPA's Integrated Risk Information System [Review]. Toxicol Appl Pharma col 254: 170

- 180. http://dx.doi.org/10.1016/j.taap.2010.10.019
Gilbert, ME. (1994). The phenomeno logy of limbic kindling [Review]. Toxicol Ind Health 10: 343

- 358. Godejohann, M; Preiss, A; Levsen, K; Wollin, KM; Mügge, C. (1998). Determination of Polar O rganic Pollutants in Aqueous Samples of Former Ammunition Sites in Lower Saxony by Means of HPLC/Photodiode Array Detection (HPLC/PDA) and Proton Nuclear Magnetic Resonance Spectroscopy (1H

- NMR). Acta Hydrochim Hydrobiol 26: 330

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