A EPA
EPA/635/R-15/034a
Public Comment Draft
www.epa.gov/iris
Toxicological Review of Hexahydro-l,3,5-trinitro-l,3,5-triazine (RDX)
(CASRN 121-82-4]
In Support of Summary Information on the
Integrated Risk Information System (IRIS)
March 2016
NOTICE
This document is a Public Comment draft. This information is distributed solely for the purpose
of pre-dissemination peer review under applicable information quality guidelines. It has not been
formally disseminated by EPA. It does not represent and should not be construed to represent any
Agency determination or policy. It is being circulated for review of its technical accuracy and
science policy implications.
National Center for Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency
Washington, DC
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Toxicological ReviewofHexahydro-l,3,5-trinitro-l,3,5-triazine
DISCLAIMER
This document is a preliminary draft for review purposes only. This information is
distributed solely for the purpose of pre-dissemination peer review under applicable information
quality guidelines. It has not been formally disseminated by EPA. It does not represent and should
not be construed to represent any Agency determination or policy. Mention of trade names or
commercial products does not constitute endorsement or recommendation for use.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review of Hexahydro-l,3,5-trinitro-l,3,5-triazine
CONTENTS
AUTHORS | CONTRIBUTORS | REVIEWERS ix
PREFACE xi
PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS xiv
EXECUTIVE SUMMARY ES-1
LITERATURE SEARCH STRATEGY | STUDY SELECTION AND EVALUATION LS-1
1. HAZARD IDENTIFICATION 1-1
1.1. Overview of Chemical Properties and Toxicokinetics 1-1
1.1.1. Chemical Properties 1-1
1.1.2. Toxicokinetics 1-2
1.1.3. Description of Toxicokinetic Models 1-3
1.2. PRESENTATION AND SYNTHESIS OF EVIDENCE BY ORGAN/SYSTEM 1-3
1.2.1. Nervous System Effects 1-4
1.2.2. Kidney and Other Urogenital System Effects 1-21
1.2.3. Reproductive and Developmental Effects 1-38
1.2.4. Liver Effects 1-49
1.2.5. Carcinogenicity 1-60
1.2.6. Other Noncancer Effects 1-69
1.3. INTEGRATION AND EVALUATION 1-69
1.3.1. Effects Other Than Cancer 1-69
1.3.2. Carcinogenicity 1-72
1.3.3. Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes. 1-73
2. DOSE-RESPONSE ANALYSIS 2-1
2.1.ORAL REFERENCE DOSE FOR EFFECTS OTHERTHAN CANCER 2-1
2.1.1. Identification of Studies for Dose-Response Analysis of Selected Effects 2-1
2.1.2. Methods of Analysis 2-5
2.1.3. Derivation of Candidate Values 2-10
2.1.4. Derivation of Organ/System-Specific Reference Doses 2-16
2.1.5. Selection of the Overall Reference Dose 2-17
2.1.6. Comparison with Mortality LDoiS 2-18
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2.1.7. Uncertainties in the Derivation of the Reference Dose 2-22
2.1.8. Confidence Statement 2-23
2.1.9. Previous IRIS Assessment 2-24
2.2. INHALATION REFERENCE CONCENTRATION FOR EFFECTS OTHER THAN CANCER 2-24
2.2.1. Previous IRIS Assessment 2-25
2.3. ORAL SLOPE FACTOR FOR CANCER 2-25
2.3.1. Analysis of Carcinogenicity Data 2-25
2.3.2. Dose-Response Analysis—Adjustments and Extrapolation Methods 2-26
2.3.3. Derivation of the Oral Slope Factor 2-27
2.3.4. Uncertainties in the Derivation of the Oral Slope Factor 2-29
2.3.5. Previous IRIS Assessment: Oral Slope Factor 2-31
2.4. INHALATION UNIT RISK FOR CANCER 2-31
2.5. APPLICATION OF AGE-DEPENDENT ADJUSTMENT FACTORS 2-31
REFERENCES R-l
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TABLES
Table ES-1. Organ/system-specific RfDs and overall RfD for RDX ES-2
Table ES-2. Summary of reference dose (RfD) derivation ES-3
Table LS-1. Inclusion-exclusion criteria for health effect studies LS-5
Table LS-2. Studies determined not to be informative because of significant issues with design,
conduct, or reporting LS-7
Table LS-3. Considerations and relevant experimental information for evaluation of
experimental animal studies LS-8
Table LS-4. Summary of experimental animal database LS-10
Table LS-5. Experimental animal studies considered less informative because of certain study
design, conduct, or reporting limitations LS-14
Table 1-1. Chemical identity and physicochemical properties of RDX 1-1
Table 1-2. Evidence pertaining to nervous system effects in humans 1-9
Table 1-3. Evidence pertaining to nervous system effects in animals 1-10
Table 1-4. Evidence pertaining to kidney effects in humans 1-24
Table 1-5. Evidence pertaining to kidney and other urogenital system effects in animals 1-25
Table 1-6. Six-, 12-, and 24-month incidence of kidney endpoints in male F344 rats reported for
statistical evaluation in (Levine et al., 1983) 1-31
Table 1-7. Six-, 12-, and 24-month incidence of urinary bladder endpoints in male F344 rats
reported for statistical evaluation in (Levine et al., 1983) 1-32
Table 1-8. Six-, 12-, and 24-month incidence of prostate endpoints in male F344 rats reported
for statistical evaluation in (Levine et al., 1983) 1-33
Table 1-9. Evidence pertaining to male reproductive effects in animals 1-41
Table 1-10. Evidence pertaining to reproductive and developmental effects in animals 1-45
Table 1-11. Evidence pertaining to liver effects in humans 1-52
Table 1-12. Evidence pertaining to liver effects in animals 1-53
Table 1-13. Liver tumors observed in chronic animal bioassays 1-63
Table 1-14. Lung tumors observed in chronic animal bioassays 1-66
Table 2-1. Information considered for evaluation of studies that examined convulsions 2-2
Table 2-2. Summary of derivation of PODs following oral exposure to RDX 2-7
Table 2-3. Effects and corresponding derivation of candidate values 2-14
Table 2-4. Organ/system-specific RfDs and overall RfD for RDX 2-16
Table 2-5. Comparison of dose levels associated with mortality and convulsions in selected
studies 2-19
Table 2-6. Summary of dose-response evaluation for mortality following oral exposure to RDX 2-
20
Table 2-7. Model predictions and OSFs for hepatocellular and alveolar/bronchiolar adenomas or
carcinomas in female B6C3Fi mice administered RDX in the diet for 2 years (Lish et
al., 1984) 2-28
Table 2-8. Summary of uncertainty in the derivation of the cancer risk value for RDX 2-29
This document is a draft for review purposes only and does not constitute Agency policy.
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FIGURES
Figure LS-1. Summary of literature search and screening process for RDX LS-4
Figure 1-1. Exposure response array of nervous system effects following oral exposure 1-16
Figure 1-2. Exposure-response array of kidney and other urogenital system effects 1-35
Figure 1-3. Exposure response array of male reproductive effects following oral exposure.... 1-44
Figure 1-5. Exposure response array of liver effects following oral exposure 1-59
Figure 2-1. Conceptual approach to dose-response modeling for oral exposure 2-5
Figure 2-2. Candidate values with corresponding POD and composite UF 2-15
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Toxicological ReviewofHexahydro-l,3,5-trinitro-l,3,5-triazine
ABBREVIATIONS
AAP Army ammunition plant GI
ACGIH American Conference of Governmental GLP
Industrial Hygienists HED
AChE acetylcholinesterase HERO
ADAF age-dependent adjustment factor
AIC Akaike's information criterion HGPRT
ALP alkaline phosphatase
ALT alanine aminotransferase HMX
AST aspartate aminotransferase
atm atmosphere IARC
ATSDR Agency for Toxic Substances and
Disease Registry i.p.
AUC area under the curve IPCS
BDNF brain-derived neurotrophic factor
BHC beta-hexachlorocyclohexane IRIS
BMC benchmark concentration IUR
BMCL benchmark concentration lower i.v.
confidence limit LDH
BMD benchmark dose LOAEL
BMDL benchmark dose lower confidence limit LOD
BMDS Benchmark Dose Software miRNA
BMDU benchmark dose upper bound MNX
BMR benchmark response
BUN blood urea nitrogen MOA
BW body weight MRL
CASRN Chemical Abstracts Service Registry NAPDH
Number
CCL Contaminant Candidate List NAS
CI confidence interval NCE
CICAD Concise International Chemical NCEA
Assessment Document
CNS central nervous system NCI
CSF cerebrospinal fluid NCTR
CYP450 cytochrome P450
DAF dosimetric adjustment factor NHANES
DDT dichlorodiphenyltrichloroethane
d.f. degrees of freedom NICNAS
DMSO dimethylsulfoxide
DNA deoxyribonucleic acid NIEHS
DNX l-nitro-3,5-dinitroso-
1,3,5-triazacyclohexane NIOSH
DTIC Defense Technical Information Center
EEG electroencephalogram NOAEL
EHC Environmental Health Criteria NOEL
EPA Environmental Protection Agency NPL
ER extra risk NRC
FDA Food and Drug Administration NSCEP
FOB functional observational battery
FUDS Formerly Used Defense Sites NTP
GABA gamma-amino butyric acid NZW
GD gestational day OR
gastrointestinal
good laboratory practices
human equivalent dose
Health and Environmental Research
Online
hypoxanthine-guanine
phosphoribosyltransferase
octahydro-l,3,5,7-tetranitro-
1,3,5,7-tetrazocine
International Agency for Research on
Cancer
intraperitoneal
International Programme on Chemical
Safety
Integrated Risk Information System
inhalation unit risk
intravenous
lactate dehydrogenase
lowest-observed-adverse-effect level
limit of detection
micro RNA
hexahydro-l-nitroso-3,5-dinitro-
1,3,5-triazine
mode of action
Minimal Risk Level
nicotinamide adenine dinucleotide
phosphate
National Academy of Science
normochromatic erythrocyte
National Center for Environmental
Assessment
National Cancer Institute
National Center for Toxicological
Research
National Health and Nutrition
Examination Survey
National Industrial Chemicals
Notification and Assessment Scheme
National Institute of Environmental
Health Sciences
National Institute for Occupational
Safety and Health
no-observed-adverse-effect level
no-observed-effect level
National Priorities List
Nuclear Regulatory Commission
National Service Center for
Environmental Publications
National Toxicology Program
New Zealand White
odds ratio
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological ReviewofHexahydro-l,3,5-trinitro-l,3,5-triazine
ORD
Office of Research and Development
SGPT
glutamic pyruvic transaminase, also
OSF
oral slope factor
known as ALT
OSHA
Occupational Safety and Health
SLE
systemic lupus erythematosus
Administration
SS
scheduled sacrifice
PBPK
physiologically based pharmacokinetic
TLV
Threshold Limit Value
PCB
polychlorinated biphenyl
TNT
trinitrotoluene
PCE
polychromatic erythrocyte
TNX
hexahydro-l,3,5-trinitroso-
PEL
Permissible Exposure Limit
1,3,5-triazine
PND
postnatal day
TSCATS
Toxic Substances Control Act Test
POD
point of departure
Submissions
PWG
Pathology Working Group
TWA
time-weighted average
RBC
red blood cell
U.S.
United States of America
RDX
Royal Demolition eXplosive
UCM
Unregulated Contaminant Monitoring
(hexahydro-l,3,5-trinitro-
UF
uncertainty factor
1,3,5-triazine]
UFa
animal-to-human uncertainty factor
REL
Recommended Exposure Limit
UFd
database deficiencies uncertainty factor
RfC
inhalation reference concentration
UFh
human variation uncertainty factor
RfD
oral reference dose
UFl
LOAEL-to-NOAEL uncertain factor
SDMS
spontaneous death or moribund
UFs
subchronic-to-chronic uncertainty
sacrifice
factor
SDWA
Safe Drinking Water Act
WBC
white blood cell
SGOT
glutamic oxaloacetic transaminase, also
WHO
World Health Organization
known as AST
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review of Hexahydro-l,3,5-trinitro-l,3,5-triazine
AUTHORS | CONTRIBUTORS | REVIEWERS
Assessment Team
Louis D'Amico, Ph.D. (Assessment Manager)
Todd Blessinger, Ph.D.
Ravi Subramaniam, Ph.D.
U.S. EPA/ORD/NCEA
Washington, DC
Christopher Brinkerhoff, Ph.D.
ORISE Postdoctoral Fellow at U.S. EPA/ORD/NCEA
Currently at U.S. EPA/Office of Chemical Safety and
Pollution Prevention
Washington, DC
Contributors
Rob DeWoskin, Ph.D.
Karen Hogan, MS
Jordan Trecki, Ph.D.
U.S. EPA/ORD/NCEA
Washington, DC
Belinda Hawkins, Ph.D.
Scott Wesselkamper, Ph.D.
U.S. EPA/ORD/NCEA
Cincinnati, OH
Tammy Stoker, Ph.D.
Charles Wood, DVM, PhD, DACVP
U.S. EPA/ORD/National Health and Environmental
Research Laboratory (NHEERL)
Research Triangle Park, NC
Anne Loccisano, Ph.D.
ORISE Postdoctoral Fellow at U.S. EPA/ORD/NCEA
Washington, DC
Production Team
Maureen Johnson
Vicki Soto
U.S. EPA/ORD/NCEA
Washington, DC
Contractor Support
Heather Carlson-Lynch, S.M., DABT
Julie Melia, Ph.D., DABT
Megan Riccardi, M.S.
SRC, Inc., North Syracuse, NY
Pam Ross, M.S.
Robin Blain, Ph.D.
ICF International, Fairfax, VA
Executive Direction
Kenneth Olden, Ph.D., Sc.D., L.H.D. (Center Director) U.S. EPA/ORD/NCEA
John Vandenberg, Ph.D, (National Program Director, HHRA) Washington, DC
Lynn Flowers, Ph.D., DABT (Associate Director for Health)
Vincent Cogliano, Ph.D. (IRIS Program Director)
Samantha Jones, Ph.D. (IRIS Associate Director for Science)
Susan Rieth, MPH (Quantitative Modeling Branch Chief)
This document is a draft for review purposes only and does not constitute Agency policy.
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Internal Review Team
General Toxicity/Immunotoxicity/Cancer Workgroup U.S. EPA/ORD/NCEA
Epidemiology Workgroup
Neurotoxicity Workgroup
Pharmacokinetics Workgroup
Reproductive and Developmental Toxicity Workgroup
Scoping and Problem Formulation Workgroup
Statistics Workgroup
Toxicity Pathways Workgroup
Reviewers
This assessment was provided for review to scientists in EPA's program and regional offices.
Comments were submitted by:
Office of the Administrator/Office of Children's Health Protection
Office of Chemical Safety and Pollution Prevention/Office of Pesticide Programs
Office of Land and Emergency Management
Office of Land and Emergency Management/Federal Facilities Forum
Office of Water
Region 2, New York, NY
Region 8, Denver, CO
This assessment was provided for review to other federal agencies and the Executive Office of the
President Comments were submitted by:
Department of Defense
Department of Energy
Department of Health and Human Services/Agency for Toxic Substances and Disease Registry
Department of Health and Human Services/National Institute of Environmental Health Sciences/National
Toxicology Program
Department of Health and Human Services/National Institute for Occupational Safety and Health
National Aeronautics and Space Administration
Executive Office of the President/Council on Environmental Quality
Executive Office of the President/Office of Management and Budget
This document is a draft for review purposes only and does not constitute Agency policy.
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PREFACE
This Toxicological Review critically reviews the publicly available studies on hexahydro-
l,3,5-trinitro-l,3,5-triazine (RDX, Royal Demolition eXplosive, or cyclonite) in order to identify its
adverse health effects and to characterize exposure-response relationships. It was prepared under
the auspices of the U.S. Environmental Protection Agency (EPA) Integrated Risk Information System
(IRIS) Program. This assessment updates a previous IRIS assessment of RDX that included an oral
reference dose (RfD) for effects other than cancer (posted in 1988), a determination on the
carcinogenicity of RDX, and a derivation of an oral slope factor (OSF) to quantify the cancer risk
associated with RDX exposure (posted in 1990). New information has become available, and this
assessment reviews information on all health effects by all exposure routes.
A public meeting was held in December 2013 to obtain input on preliminary materials for
RDX, including draft literature searches and associated search strategies, evidence tables, and
exposure-response arrays prior to the development of the IRIS assessment. All public comments
provided were taken into consideration in developing the draft assessment The complete set of
public comments are available on the docket at http: //www.regulations.gov (Docket ID No. EPA-
HQ-0RD-2 013-0430).
Organ/system-specific reference values are calculated based on nervous system, kidney/
urogenital system, and male reproductive toxicity data. These reference values may be useful for
cumulative risk assessments that consider the combined effect of multiple agents acting on the
same biological system.
This assessment was conducted in accordance with EPA guidance, which is summarized in
the Preamble to IRIS Toxicological Reviews and cited at appropriate places in this assessment. The
findings of this assessment and related documents produced during its development are available
on the IRIS website (http://www.epa.gov/iris). Appendices containing information on assessments
by other health agencies, details of the literature search strategy, toxicokinetic information,
summaries of supplementary toxicity information, and dose-response modeling are provided as
Supplemental Information to this assessment (see Appendices A to D).
For additional information about this assessment or for general questions regarding IRIS,
please contact EPA's IRIS Hotline at 202-566-1676 (phone), or hotline.iris@epa.gov.
Uses and Environmental Occurrence
RDX is a military munitions explosive with limited civilian uses (Gadagbui etal.. 2012). In
the United States, RDX is produced at Army ammunition plants (AAPs) and is not produced
commercially. RDX production peaked in the 1960s; 180 million pounds per year were produced
from 1969 to 1971. Yearly total production dropped to 16 million pounds in 1984 fATSDR. 20121.
This document is a draft for review purposes only and does not constitute Agency policy.
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According to the U.S. EPA Chemical Data Access Tool f http://iava.epa.gov/oppt chemical search/1,
the aggregate national production volume in 2012 was approximately 6.3 million pounds per year.
RDX can be released into environmental media (air, water, soil) as a result of waste
generated during manufacture, packing, or disposal of the pure product, or use and disposal of RDX-
containing munitions fATSDR. 2012: Gadagbui etal.. 2012: ATSDR. 1999.1993.19921. RDX is
mobile in soil; leaching into groundwater has been reported in samples from military facilities (Best
etal.. 1999a: Godeiohann etal.. 1998: Bart etal.. 1997: Steuckart et al.. 1994: Spanggord etal..
1980a). RDX transport in soil is generally through dissolution by precipitation and subsequent
downward movement, including migration to groundwater aquifers, and not much via surface
runoff fU.S. EPA. 2012cl. An extensive discussion of RDX properties and fate and transport is
available in U.S. EPA (2012c). Detectable levels of RDX have been observed in plants irrigated or
grown with RDX-contaminated water fBestetal.. 1999b: Simini and Checkai. 1996: Harvey etal..
1991). RDX has also been detected in indoor air samples from military facilities where RDX is
produced (Bishop etal.. 1988).
Exposures to RDX among the general population are likely to be confined to individuals in
or around active or formerly-used military facilities where RDX is or was produced, stored, or used.
Oral, inhalation, and dermal routes of exposure may be relevant
As of 2015, RDX was detected in surface water, groundwater, sediment, or soil at 34 current
U.S. EPA National Priorities List (NPL) sites. The NPL serves as a list of sites with known or
threatened releases of hazardous substances, pollutants, or contaminants throughout the United
States and its territories. The NPL aids the Agency in identifying the most serious sites that may
warrant cleanup. The majority of the NPL sites where RDX was listed are associated with military
facilities. Based on Department of Defense records, Gadagbui etal. T20121 reported that RDX
contamination is present on 76 active military sites, 9 closed sites, and 15 sites under the Formerly
Used Defense Sites (FUDS) program. Not all sites under the FUDS program have been sampled, and
additional sites with RDX contamination in this program could be identified.
As of 2015, RDX was not regulated under the Safe Drinking Water Act (SDWA), although it
was included as a contaminant to be monitored under the Unregulated Contaminant Monitoring
(UCM) Rule by EPA's Office of Water from 2007 to 2011. Contaminants included in the UCM
program are suspected of being present in drinking water, but do not have existing health-based
standards set under the SDWA. RDX has also been included on the Office of Water's Drinking Water
Contaminant Candidate List (CCL) since the initial listing was published in 1998. The presence of a
chemical on the list suggests that it is known or anticipated to occur in public water systems.
Assessments by Other National and International Health Agencies
Toxicity information on RDX has been evaluated by the Agency for Toxic Substances and
Disease Registry (ATSDR), National Institute for Occupational Safety and Health (NIOSH),
Occupational Safety and Health Administration (OSHA), and Australian National Industrial
Chemicals Notification and Assessment Scheme (NICNAS). The results of these assessments (as of
This document is a draft for review purposes only and does not constitute Agency policy.
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1 2015) are presented in Appendix A of the Supplemental Information. It is important to recognize
2 that the assessments performed by other health agencies may have been prepared for different
3 purposes and may utilize different methods. In addition, newer studies may be included in the IRIS
4 assessment.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological Review of Hexahydro-l,3,5-trinitro-l,3,5-triazine
PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS
Note: The Preamble to IRIS assessments is bein
external peer reviewers and the public, and ba:
implementation of systematic review methods,
include the revised Preamble.
1. Scope of the IRIS Program
Soon after the EPA was established in
1970, it was at the forefront of developing
risk assessment as a science and applying it in
decisions to protect human health and the
environment. The Clean Air Act, for example,
mandates that the EPA provide "an ample
margin of safety to protect public health"; the
Safe Drinking Water Act, that "no adverse
effects on the health of persons may
reasonably be anticipated to occur, allowing
an adequate margin of safety." Accordingly,
the EPA uses information on the adverse
effects of chemicals and on exposure levels
below which these effects are not anticipated
to occur.
IRIS assessments critically review the
publicly available studies to identify adverse
health effects from exposure to chemicals and
to characterize exposure-response
relationships. In terms set forth by the
National Research Council fNRC. 19831. IRIS
assessments cover the hazard identification
and dose-response assessment steps of risk
assessment, not the exposure assessment or
risk characterization steps that are
conducted by the EPA's program and regional
offices and by other federal, state, and local
health agencies that evaluate risk in specific
populations and exposure scenarios. IRIS
assessments are distinct from and do not
address political, economic, and technical
considerations that influence the design and
selection of risk management alternatives.
An IRIS assessment may cover a single
chemical, a group of structurally or
7 revised based on comments received from
ed on IRIS Program experience with the
Subsequent drafts of the RDX assessment will
39 toxicologically related chemicals, or a
40 complex mixture. These agents may be found
41 in air, water, soil, or sediment. Exceptions are
42 chemicals currently used exclusively as
43 pesticides, ionizing and non-ionizing
44 radiation, and criteria air pollutants listed
45 under Section 108 of the Clean Air Act
46 (carbon monoxide, lead, nitrogen oxides,
47 ozone, particulate matter, and sulfur oxides).
48 Periodically, the IRIS Program asks other
49 EPA programs and regions, other federal
50 agencies, state health agencies, and the
51 general public to nominate chemicals and
52 mixtures for future assessment or
53 reassessment. Agents may be considered for
54 reassessment as significant new studies are
55 published. Selection is based on program and
56 regional office priorities and on availability of
57 adequate information to evaluate the
58 potential for adverse effects. Other agents
59 may also be assessed in response to an urgent
60 public health need.
61 2. Process for developing and peer-
62 reviewing IRIS assessments
63 The process for developing IRIS
64 assessments (revised in May 2009 and
65 enhanced in July 2013) involves critical
66 analysis of the pertinent studies,
67 opportunities for public input, and multiple
68 levels of scientific review. The EPA revises
69 draft assessments after each review, and
70 external drafts and comments become part of
71 the public record (U.S. EPA. 2009).
72 Before beginning an assessment, the IRIS
73 Program discusses the scope with other EPA
74 programs and regions to ensure that the
This document is a draft for review purposes only and does not constitute Agency policy.
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assessment will meet their needs. Then a
public meeting on problem formulation
invites discussion of the key issues and the
studies and analytical approaches that might
contribute to their resolution.
Step 1. Development of a draft
Toxicological Review. The draft
assessment considers all pertinent
publicly available studies and applies
consistent criteria to evaluate study
quality, identify health effects, identify
mechanistic events and pathways,
integrate the evidence of causation for
each effect, and derive toxicity values. A
public meeting prior to the integration of
evidence and derivation of toxicity values
promotes public discussion of the
literature search, evidence, and key
issues.
Step 2. Internal review by scientists in
EPA programs and regions. The draft
assessment is revised to address the
comments from within the EPA.
Step 3. Interagency science consultation
with other federal agencies and the
Executive Offices of the President. The
draft assessment is revised to address the
interagency comments. The science
consultation draft, interagency
comments, and the EPA's response to
major comments become part of the
public record.
Step 4. Public review and comment,
followed by external peer review. The
EPA releases the draft assessment for
public review and comment A public
meeting provides an opportunity to
discuss the assessment prior to peer
review. Then the EPA releases a draft for
external peer review. The peer review
meeting is open to the public and includes
time for oral public comments. The peer
reviewers assess whether the evidence
has been assembled and evaluated
according to guidelines and whether the
conclusions are justified by the evidence.
The peer review draft, written public
48 comments, and peer review report
49 become part of the public record.
50 Step 5. Revision of draft Toxicological
51 Review and development of draft IRIS
52 summary. The draft assessment is
53 revised to reflect the peer review
54 comments, public comments, and newly
55 published studies that are critical to the
56 conclusions of the assessment. The
57 disposition of peer review comments and
58 public comments becomes part of the
59 public record.
60 Step 6. Final EPA review and interagency
61 science discussion with other federal
62 agencies and the Executive Offices of
63 the President The draft assessment and
64 summary are revised to address the EPA
65 and interagency comments. The science
66 discussion draft, written interagency
67 comments, and EPA's response to major
68 comments become part of the public
69 record.
70 Step 7. Completion and posting. The
71 Toxicological Review and IRIS summary
72 are posted on the IRIS website
73 fhttp: / /www, epa. gov /iris /).
74 The remainder of this Preamble
75 addresses step 1, the development of a draft
76 Toxicological Review. IRIS assessments
77 follow standard practices of evidence
78 evaluation and peer review, many of which
79 are discussed in EPA guidelines fU.S. EPA.
80 2005a. b. 2000b. 1998.1996.1991.1986a. bl
81 and other methods (U.S. EPA. 2012a. b, 2011.
82 2006a. b, 2002. 19941. Transparent
83 application of scientific judgment is of
84 paramount importance. To provide a
85 harmonized approach across IRIS
86 assessments, this Preamble summarizes
87 concepts from these guidelines and
88 emphasizes principles of general
89 applicability.
90 3. Identifying and selecting
91 pertinent studies
92 3.1. Identifying studies
This document is a draft for review purposes only and does not constitute Agency policy.
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Before beginning an assessment, the EPA
conducts a comprehensive search of the
primary scientific literature. The literature
search follows standard practices and
includes the PubMed and ToxNet databases
of the National Library of Medicine, Web of
Science, and other databases listed in the
EPA's HERO system (Health and
Environmental Research Online,
http://her0.epa.g0v/l Searches for
information on mechanisms of toxicity are
inherently specialized and may include
studies on other agents that act through
related mechanisms.
Each assessment specifies the search
strategies, keywords, and cut-off dates of its
literature searches. The EPA posts the results
of the literature search on the IRIS website
and requests information from the public on
additional studies and ongoing research.
The EPA also considers studies received
through the IRIS Submission Desk and
studies (typically unpublished) submitted
under the Toxic Substances Control Actor the
Federal Insecticide, Fungicide, and
Rodenticide Act Material submitted as
Confidential Business Information is
considered only if it includes health and
safety data that can be publicly released. If a
study that may be critical to the conclusions
of the assessment has not been peer-
reviewed, the EPA will have it peer-reviewed.
The EPA also examines the toxicokinetics
of the agent to identify other chemicals (for
example, major metabolites of the agent) to
include in the assessment if adequate
information is available, in order to more
fully explain the toxicity of the agent and to
suggest dose metrics for subsequent
modeling.
In assessments of chemical mixtures,
mixture studies are preferred for their ability
to reflect interactions among components.
The literature search seeks, in
decreasing order of preference fU.S. EPA.
2000b. §2.2: 1986b. 52.111:
Studies of the mixture being assessed.
Studies of a sufficiently similar
mixture. In evaluating similarity, the
assessment considers the alteration
of mixtures in the environment
through partitioning and
transformation.
Studies of individual chemical
components of the mixture, if there
are not adequate studies of
sufficiently similar mixtures.
3.2. Selecting pertinent epidemiologic
studies
Study design is the key consideration for
selecting pertinent epidemiologic studies
from the results of the literature search.
Cohort studies, case-control studies,
and some population-based surveys
(for example, NHANES) provide the
strongest epidemiologic evidence,
especially if they collect information
about individual exposures and
effects.
Ecological studies (geographic
correlation studies) relate exposures
and effects by geographic area. They
can provide strong evidence if there
are large exposure contrasts between
geographic areas, relatively little
exposure variation within study
areas, and population migration is
limited.
Case reports of high or accidental
exposure lack definition of the
population at risk and the expected
number of cases. They can provide
information about a rare effect or
about the relevance of analogous
results in animals.
The assessment briefly reviews
ecological studies and case reports but
reports details only if they suggest effects not
identified by other studies.
3.3. Selecting pertinent experimental
studies
Exposure route is a key design
consideration for selecting pertinent
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experimental animal studies or human
clinical studies.
Studies of oral, inhalation, or dermal
exposure involve passage through an
absorption barrier and are
considered most pertinent to human
environmental exposure.
Injection or implantation studies are
often considered less pertinent but
may provide valuable toxicokinetic or
mechanistic information. They also
may be useful for identifying effects in
animals if deposition or absorption is
problematic (for example, for
particles and fibers).
Exposure duration is also a key design
consideration for selecting pertinent
experimental animal studies.
Studies of effects from chronic
exposure are most pertinent to
lifetime human exposure.
Studies of effects from less-than-
chronic exposure are pertinent but
less preferred for identifying effects
from lifetime human exposure. Such
studies may be indicative of effects
from less-than-lifetime human
exposure.
Short-duration studies involving animals
or humans may provide toxicokinetic or
mechanistic information.
For developmental toxicity and
reproductive toxicity, irreversible effects
may result from a brief exposure during a
critical period of development. Accordingly,
specialized study designs are used for these
effects (TJ.S. EPA. 2006b. 1998. 1996. 19911.
4. Evaluating the quality of
individual studies
After the subsets of pertinent
epidemiologic and experimental studies have
been selected from the literature searches,
the assessment evaluates the quality of each
individual study. This evaluation considers
the design, methods, conduct, and
44 documentation of each study, but not
45 whether the results are positive, negative, or
46 null. The objective is to identify the stronger,
47 more informative studies based on a uniform
48 evaluation of quality characteristics across
49 studies of similar design.
4.1. Evaluating the quality of
epidemiologic studies
50 The assessment evaluates design and
51 methodological aspects that can increase or
52 decrease the weight given to each
53 epidemiologic study in the overall evaluation
54 fU.S. EPA. 2005a. 1998. 1996.1994.19911:
55
Documentation of study design,
56
methods, population characteristics,
57
and results.
58
Definition and selection of the study
59
group and comparison group.
60
- Ascertainment of exposure to the
61
chemical or mixture.
62
- Ascertainment of disease or health
63
effect
64
Duration of exposure and follow-up
65
and adequacy for assessing the
66
occurrence of effects.
67
Characterization of exposure during
68
critical periods.
69
Sample size and statistical power to
70
detect anticipated effects.
71
Participation rates and potential for
72
selection bias as a result of the
73
achieved participation rates.
74
Measurement error (can lead to
75
misclassification of exposure, health
76
outcomes, and other factors) and
77
other types of information bias.
78
Potential confounding and other
79
sources of bias addressed in the study
80
design or in the analysis of results.
81
The basis for consideration of
82
confounding is a reasonable
83
expectation that the confounder is
84
related to both exposure and
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outcome and is sufficiently prevalent
to result in bias.
For developmental toxicity, reproductive
toxicity, neurotoxicity, and cancer there is
further guidance on the nuances of evaluating
epidemiologic studies of these effects (U.S.
EPA. 2005a. 1998.1996.19911.
4.2. Evaluating the quality of
experimental studies
The assessment evaluates design and
methodological aspects that can increase or
decrease the weight given to each
experimental animal study, in-vitro study, or
human clinical study fU.S. EPA. 2005a. 1998.
1996. 19911. Research involving human
subjects is considered only if conducted
according to ethical principles.
Documentation of study design,
animals or study population,
methods, basic data, and results.
Nature of the assay and validity for its
intended purpose.
Characterization of the nature and
extent of impurities and
contaminants of the administered
chemical or mixture.
Characterization of dose and dosing
regimen (including age at exposure)
and their adequacy to elicit adverse
effects, including latent effects.
Sample sizes and statistical power to
detect dose-related differences or
trends.
- Ascertainment of survival, vital signs,
disease or effects, and cause of death.
Control of other variables that could
influence the occurrence of effects.
The assessment uses statistical tests to
evaluate whether the observations may be
due to chance. The standard for determining
statistical significance of a response is a trend
test or comparison of outcomes in the
exposed groups against those of concurrent
controls. In some situations, examination of
43 historical control data from the same
44 laboratory within a few years of the study
45 may improve the analysis. For an uncommon
46 effect that is not statistically significant
47 compared with concurrent controls,
48 historical controls may show that the effect is
49 unlikely to be due to chance. For a response
50 that appears significant against a concurrent
51 control response that is unusual, historical
52 controls may offer a different interpretation
53 fU.S. EPA. 2005a. §2.2.2.1.31.
54 For developmental toxicity, reproductive
55 toxicity, neurotoxicity, and cancer there is
56 further guidance on the nuances of evaluating
57 experimental studies of these effects (U.S.
58 EPA. 2005a. 1998. 1996. 19911. In multi-
59 generation studies, agents that produce
60 developmental effects at doses that are not
61 toxic to the maternal animal are of special
62 concern. Effects that occur at doses
63 associated with mild maternal toxicity are not
64 assumed to result only from maternal
65 toxicity. Moreover, maternal effects may be
66 reversible, while effects on the offspring may
67 be permanent (U.S. EPA. 1998. §3.1.2.4.5.4:
68 1991. §3.1.1.41..
4.3. Reporting study results
69 The assessment uses evidence tables to
70 present the design and key results of
71 pertinent studies. There may be separate
72 tables for each site of toxicity or type of study.
73 If a large number of studies observe the
74 same effect, the assessment considers the
75 study quality characteristics in this section to
76 identify the strongest studies or types of
77 study. The tables present details from these
78 studies, and the assessment explains the
79 reasons for not reporting details of other
80 studies or groups of studies that do not add
81 new information. Supplemental information
82 provides references to all studies considered,
83 including those not summarized in the tables.
84 The assessment discusses strengths and
85 limitations that affect the interpretation of
86 each study. If the interpretation of a study in
87 the assessment differs from that of the study
88 authors, the assessment discusses the basis
89 for the difference.
This document is a draft for review purposes only and does not constitute Agency policy.
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As a check on the selection and evaluation
of pertinent studies, the EPA asks peer
reviewers to identify studies that were not
adequately considered.
5. Evaluating the overall evidence
of each effect
5.1. Concepts of causal inference
For each health effect, the assessment
evaluates the evidence as a whole to
determine whether it is reasonable to infer a
causal association between exposure to the
agent and the occurrence of the effect This
inference is based on information from
pertinent human studies, animal studies, and
mechanistic studies of adequate quality.
Positive, negative, and null results are given
weight according to study quality.
Causal inference involves scientific
judgment, and the considerations are
nuanced and complex. Several health
agencies have developed frameworks for
causal inference, among them the U.S.
Surgeon General fCDC. 2004: HEW. 19641.
the International Agency for Research on
Cancer (IARC. 20061. the Institute of Medicine
flOM. 20081. and the fU.S. EPA C20101. §1.6:
2005al. §2.51. Although developed for
different purposes, the frameworks are
similar in nature and provide an established
structure and language for causal inference.
Each considers aspects of an association that
suggest causation, discussed by Hill (19651
and elaborated by Roth man and Greenland
(19981. and (TJ.S. EPA C2005al. §2.2.1.7:
19941. Appendix CI.
Strength of association: The finding of a
large relative risk with narrow
confidence intervals strongly suggests
that an association is not due to chance,
bias, or other factors. Modest relative
risks, however, may reflect a small range
of exposures, an agent of low potency, an
increase in an effect that is common,
exposure misclassification, or other
sources of bias.
43 Consistency of association: An inference of
44 causation is strengthened if elevated
45 risks are observed in independent studies
46 of different populations and exposure
47 scenarios. Reproducibility of findings
48 constitutes one of the strongest
49 arguments for causation. Discordant
50 results sometimes reflect differences in
51 study design, exposure, or confounding
52 factors.
53 Specificity of association: As originally
54 intended, this refers to one cause
55 associated with one effect. Current
56 understanding that many agents cause
57 multiple effects and many effects have
58 multiple causes make this a less
59 informative aspect of causation, unless
60 the effect is rare or unlikely to have
61 multiple causes.
62 Temporal relationship: A causal
63 interpretation requires that exposure
64 precede development of the effect.
65 Biologic gradient (exposure-response
66 relationship): Exposure-response
67 relationships strongly suggest causation.
68 A monotonic increase is not the only
69 pattern consistent with causation. The
70 presence of an exposure-response
71 gradient also weighs against bias and
72 confounding as the source of an
73 association.
74 Biologic plausibility: An inference of
75 causation is strengthened by data
76 demonstrating plausible biologic
77 mechanisms, if available. Plausibility
78 may reflect subjective prior beliefs if
79 there is insufficient understanding of the
80 biologic process involved.
81 Coherence: An inference of causation is
82 strengthened by supportive results from
83 animal experiments, toxicokinetic
84 studies, and short-term tests. Coherence
85 may also be found in other lines of
86 evidence, such as changing disease
87 patterns in the population.
88 "Natural experiments": A change in
89 exposure that brings about a change in
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disease frequency provides strong
evidence, as it tests the hypothesis of
causation. An example would be an
intervention to reduce exposure in the
workplace or environment that is
followed by a reduction of an adverse
effect.
Analogy: Information on structural
analogues or on chemicals that induce
similar mechanistic events can provide
insight into causation.
These considerations are consistent with
guidelines for systematic reviews that
evaluate the quality and weight of evidence.
Confidence is increased if the magnitude of
effect is large, if there is evidence of an
exposure-response relationship, or if an
association was observed and the plausible
biases would tend to decrease the magnitude
of the reported effect. Confidence is
decreased for study limitations,
inconsistency of results, indirectness of
evidence, imprecision, or reporting bias
(Guvattetal.. 2008b: Guvatt et al.. 2008al
5.2. Evaluating evidence in humans
For each effect, the assessment evaluates
the evidence from the epidemiologic studies
as a whole. The objective is to determine
whether a credible association has been
observed and, if so, whether that association
is consistent with causation. In doing this, the
assessment explores alternative explanations
(such as chance, bias, and confounding) and
draws a conclusion about whether these
alternatives can satisfactorily explain any
observed association.
To make clear how much the
epidemiologic evidence contributes to the
overall weight of the evidence, the
assessment may select a standard descriptor
to characterize the epidemiologic evidence of
association between exposure to the agent
and occurrence of a health effect.
Sufficient epidemiologic evidence of an
association consistent with causation:
The evidence establishes a causal
association for which alternative
47 explanations such as chance, bias, and
48 confounding can be ruled out with
49 reasonable confidence.
50 Suggestive epidemiologic evidence of an
51 association consistent with causation:
52 The evidence suggests a causal
53 association but chance, bias, or
54 confounding cannot be ruled out as
55 explaining the association.
56 Inadequate epidemiologic evidence to infer
57 a causal association: The available
58 studies do not permit a conclusion
59 regarding the presence or absence of an
60 association.
61 Epidemiologic evidence consistent with no
62 causal association: Several adequate
63 studies covering the full range of human
64 exposures and considering susceptible
65 populations, and for which alternative
66 explanations such as bias and
67 confounding can be ruled out, are
68 mutually consistent in not finding an
69 association.
5.3. Evaluating evidence in animals
70 For each effect, the assessment evaluates
71 the evidence from the animal experiments as
72 a whole to determine the extent to which they
73 indicate a potential for effects in humans.
74 Consistent results across various species and
75 strains increase confidence that similar
76 results would occur in humans. Several
77 concepts discussed by Hill (1965) are
78 pertinent to the weight of experimental
79 results: consistency of response, dose-
80 response relationships, strength of response,
81 biologic plausibility, and coherence fU.S. EPA.
82 2005a. §2.2.1.7: 1994. Appendix CI
83 In weighing evidence from multiple
84 experiments, U.S. EPA (2005a). §2.5
85 distinguishes:
86 Conflicting evidence (that is, mixed positive
87 and negative results in the same sex and
88 strain using a similar study protocol)
89 from
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Differing results (that is, positive results and
negative results are in different sexes or
strains or use different study protocols).
Negative or null results do not invalidate
positive results in a different experimental
system. The EPA regards all as valid
observations and looks to explain differing
results using mechanistic information (for
example, physiologic or metabolic
differences across test systems) or
methodological differences (for example,
relative sensitivity of the tests, differences in
dose levels, insufficient sample size, or timing
of dosing or data collection).
It is well established that there are critical
periods for some developmental and
reproductive effects (U.S. EPA. 2006b. 2005a.
b, 1998. 1996. 1991). Accordingly, the
assessment determines whether critical
periods have been adequately investigated.
Similarly, the assessment determines
whether the database is adequate to evaluate
other critical sites and effects.
In evaluating evidence of genetic toxicity:
Demonstration of gene mutations,
chromosome aberrations, or
aneuploidy in humans or
experimental mammals [in vivo)
provides the strongest evidence.
- This is followed by positive results in
lower organisms or in cultured cells
[in vitro) or for other genetic events.
Negative results carry less weight,
partly because they cannot exclude
the possibility of effects in other
tissues flARC. 20061.
For germ-cell mutagenicity, the EPA has
defined categories of evidence, ranging from
positive results of human germ-cell
mutagenicity to negative results for all effects
of concern fU.S. EPA. 1986a. 52.31.
5.4. Evaluating mechanistic data
Mechanistic data can be useful in
answering several questions.
44 - The biologic plausibility of a causal
45 interpretation of human studies.
46 - The generalizability of animal studies
47 to humans.
48 - The susceptibility of particular
49 populations or lifestages.
50 The focus of the analysis is to describe, if
51 possible, mechanistic pathways that lead to a
52 health effect. These pathways encompass:
53 - Toxicokinetic processes of absorption,
54 distribution, metabolism, and
55 elimination that lead to the formation
56 of an active agent and its presence at
57 the site of initial biologic interaction.
58 - Toxicodynamic processes that lead to a
59 health effect at this or another site
60 (also known as a mode of action).
61 For each effect, the assessment discusses
62 the available information on its modes of
63 action and associated key events [key events
64 being empirically observable, necessary
65 precursor steps or biologic markers of such
66 steps; mode of action being a series of key
67 events involving interaction with cells,
68 operational and anatomic changes, and
69 resulting in disease). Pertinent information
70 may also come from studies of metabolites or
71 of compounds that are structurally similar or
72 that act through similar mechanisms.
73 Information on mode of action is not required
74 for a conclusion that the agent is causally
75 related to an effect fU.S. EPA. 2005a. 52.51.
76 The assessment addresses several
77 questions about each hypothesized mode of
78 action flJ.S. EPA. 2005a. §2.4.3.41.
79 1) Is the hypothesized mode of action
80 sufficiently supported in test animals?
81 Strong support for a key event being
82 necessary to a mode of action can come
83 from experimental challenge to the
84 hypothesized mode of action, in which
85 studies that suppress a key event observe
86 suppression of the effect Support for a
87 mode of action is meaningfully
88 strengthened by consistent results in
89 different experimental models, much
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more so than by replicate experiments in
the same model. The assessment may
consider various aspects of causation in
addressing this question.
2) Is the hypothesized mode of action
relevant to humans? The assessment
reviews the key events to identify critical
similarities and differences between the
test animals and humans. Site
concordance is not assumed between
animals and humans, though it may hold
for certain effects or modes of action.
Information suggesting quantitative
differences in doses where effects would
occur in animals or humans is considered
in the dose-response analysis. Current
levels of human exposure are not used to
rule out human relevance, as IRIS
assessments may be used in evaluating
new or unforeseen circumstances that
may entail higher exposures.
3) Which populations or lifestages can be
particularly susceptible to the
hypothesized mode of action? The
assessment reviews the key events to
identify populations and lifestages that
might be susceptible to their occurrence.
Quantitative differences may result in
separate toxicity values for susceptible
populations or lifestages.
The assessment discusses the likelihood
that an agent operates through multiple
modes of action. An uneven level of support
for different modes of action can reflect
disproportionate resources spent
investigating them (U.S. EPA.
2005a. §2.4.3.31. It should be noted that in
clinical reviews, the credibility of a series of
studies is reduced if evidence is limited to
studies funded by one interested sector
fCiiivattetal.. 2008al.
For cancer, the assessment evaluates
evidence of a mutagenic mode of action to
guide extrapolation to lower doses and
consideration of susceptible lifestages. Key
data include the ability of the agent or a
metabolite to react with or bind to DNA,
positive results in multiple test systems, or
49 similar properties and structure-activity
50 relationships to mutagenic carcinogens (U.S.
51 EPA. 2005a .52.3.51
5.5. Characterizing the overall weight
of the evidence
52 After evaluating the human, animal, and
53 mechanistic evidence pertinent to an effect,
54 the assessment answers the question: Does
55 the agent cause the adverse effect? fNRC.
56 2009. 19831. In doing this, the assessment
57 develops a narrative that integrates the
58 evidence pertinent to causation. To provide
59 clarity and consistency, the narrative
60 includes a standard hazard descriptor. For
61 example, the following standard descriptors
62 combine epidemiologic, experimental, and
63 mechanistic evidence of carcinogenicity (U.S.
64 EPA. 2005a. 52.51.
65 Carcinogenic to humans: There is
66 convincing epidemiologic evidence of a
67 causal association (that is, there is
68 reasonable confidence that the
69 association cannot be fully explained by
70 chance, bias, or confounding); or there is
71 strong human evidence of cancer or its
72 precursors, extensive animal evidence,
73 identification of key precursor events in
74 animals, and strong evidence that they
75 are anticipated to occur in humans.
76 Likely to be carcinogenic to humans: The
77 evidence demonstrates a potential
78 hazard to humans but does not meet the
79 criteria for carcinogenic. There may be a
80 plausible association in humans, multiple
81 positive results in animals, or a
82 combination of human, animal, or other
83 experimental evidence.
84 Suggestive evidence of carcinogenic
85 potential: The evidence raises concern
86 for effects in humans but is not sufficient
87 for a stronger conclusion. This descriptor
88 covers a range of evidence, from a
89 positive result in the only available study
90 to a single positive result in an extensive
91 database that includes negative results in
92 other species.
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Inadequate information to assess
carcinogenic potential: No other
descriptors apply. Conflicting evidence
can be classified as inadequate
information if all positive results are
opposed by negative studies of equal
quality in the same sex and strain.
Differing results, however, can be
classified as suggestive evidence or as
likely to be carcinogenic.
Not likely to be carcinogenic to humans:
There is robust evidence for concluding
that there is no basis for concern. There
may be no effects in both sexes of at least
two appropriate animal species; positive
animal results and strong, consistent
evidence that each mode of action in
animals does not operate in humans; or
convincing evidence that effects are not
likely by a particular exposure route or
below a defined dose.
Multiple descriptors may be used if there
is evidence that carcinogenic effects differ by
dose range or exposure route (U.S. EPA.
2005a. §2.51
Another example of standard descriptors
comes from the EPA's Integrated Science
Assessments, which evaluate causation for
the effects of the criteria pollutants in
ambient air fU.S. EPA. 20101.
Causal relationship: Sufficient evidence to
conclude that there is a causal
relationship. Observational studies
cannot be explained by plausible
alternatives, or they are supported by
other lines of evidence, for example,
animal studies or mechanistic
information.
Likely to be a causal relationship: Sufficient
evidence that a causal relationship is
likely, but important uncertainties
remain. For example, observational
studies show an association but co-
exposures are difficult to address or other
lines of evidence are limited or
inconsistent; or multiple animal studies
from different laboratories demonstrate
effects and there are limited or no human
data.
Suggestive of a causal relationship: At least
one high-quality epidemiologic study
shows an association but other studies
are inconsistent
Inadequate to infer a causal relationship:
The studies do not permit a conclusion
regarding the presence or absence of an
association.
Not likely to be a causal relationship:
Several adequate studies, covering the
full range of human exposure and
considering susceptible populations, are
mutually consistent in not showing an
effect at any level of exposure.
The EPA is investigating and may on a
trial basis use these or other standard
descriptors to characterize the overall weight
of the evidence for effects other than cancer.
6. Selecting studies for derivation
of toxicity values
For each effect where there is credible
evidence of an association with the agent, the
assessment derives toxicity values if there
are suitable epidemiologic or experimental
data. The decision to derive toxicity values
may be linked to the hazard descriptor.
Dose-response analysis requires
quantitative measures of dose and response.
Then, other factors being equal:
Epidemiologic studies are preferred
over animal studies, if quantitative
measures of exposure are available
and effects can be attributed to the
agent
- Among experimental animal models,
those that respond most like humans
are preferred, if the comparability of
response can be determined.
Studies by a route of human
environmental exposure are
preferred, although a validated
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toxicokinetic model can be used to
extrapolate across exposure routes.
Studies of longer exposure duration
and follow-up are preferred, to
minimize uncertainty about whether
effects are representative of lifetime
exposure.
Studies with multiple exposure levels
are preferred for their ability to
provide information about the shape
of the exposure-response curve.
Studies with adequate power to
detect effects at lower exposure
levels are preferred, to minimize the
extent of extrapolation to levels found
in the environment
Studies with non-monotonic exposure-
response relationships are not necessarily
excluded from the analysis. A diminished
effect at higher exposure levels may be
satisfactorily explained by factors such as
competing toxicity, saturation of absorption
or metabolism, exposure misclassification, or
selection bias.
If a large number of studies are suitable
for dose-response analysis, the assessment
considers the study characteristics in this
section to focus on the most informative data.
The assessment explains the reasons for not
analyzing other groups of studies. As a check
on the selection of studies for dose-response
analysis, the EPA asks peer reviewers to
identify studies that were not adequately
considered.
7. Deriving toxicity values
7.1. General framework for dose-
response analysis
The EPA uses a two-step approach that
distinguishes analysis of the observed dose-
response data from inferences about lower
doses fU.S. EPA. 2005a. §3").
Within the observed range, the preferred
approach is to use modeling to incorporate a
wide range of data into the analysis. The
modeling yields a point of departure (an
exposure level near the lower end of the
observed range, without significant
extrapolation to lower doses) (Sections 7.2-
7.3).
Extrapolation to lower doses considers
what is known about the modes of action for
each effect (Sections 7.4-7.5). If response
estimates at lower doses are not required, an
alternative is to derive reference values,
which are calculated by applying factors to
the point of departure in order to account for
sources of uncertainty and variability
(Section 7.6).
For a group of agents that induce an effect
through a common mode of action, the dose-
response analysis may derive a relative
potency factor for each agent A full dose-
response analysis is conducted for one well-
studied index chemical in the group, then the
potencies of other members are expressed in
relative terms based on relative toxic effects,
relative absorption or metabolic rates,
quantitative structure-activity relationships,
or receptor binding characteristics (U.S. EPA.
2005a. §3.2.6: 2000b. 54.41.
Increasingly, the EPA is basing toxicity
values on combined analyses of multiple data
sets or multiple responses. The EPA also
considers multiple dose-response
approaches if they can be supported by
robust data.
7.2. Modeling dose to sites of biologic
effects
The preferred approach for analysis of
dose is toxicokinetic modeling because of its
ability to incorporate a wide range of data.
The preferred dose metric would refer to the
active agent at the site of its biologic effect or
to a close, reliable surrogate measure. The
active agent may be the administered
chemical or a metabolite. Confidence in the
use of a toxicokinetic model depends on the
robustness of its validation process and on
the results of sensitivity analyses fU.S. EPA.
2006a: 2005a. 53.1: 1994. 54.31.
Because toxicokinetic modeling can
require many parameters and more data than
are typically available, the EPA has developed
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standard approaches that can be applied to
typical data sets. These standard approaches
also facilitate comparison across exposure
patterns and species.
Intermittent study exposures are
standardized to a daily average over
the duration of exposure. For chronic
effects, daily exposures are averaged
over the lifespan. Exposures during a
critical period, however, are not
averaged over a longer duration fU.S.
EPA. 2005a. 33.1.1: 1991. 53.21.
Doses are standardized to equivalent
human terms to facilitate comparison
of results from different species.
Oral doses are scaled allometrically
using mg/kg3/4-day as the equivalent
dose metric across species.
Allometric scaling pertains to
equivalence across species, not
across lifestages, and is not used to
scale doses from adult humans or
mature animals to infants or children
fU.S. EPA. 2011: 2005a. 33.1.31.
Inhalation exposures are scaled using
dosimetry models that apply species-
specific physiologic and anatomic
factors and consider whether the
effect occurs at the site of first contact
or after systemic circulation (U.S.
EPA. 2012a: 1994.331.
It can be informative to convert doses
across exposure routes. If this is done, the
assessment describes the underlying data,
algorithms, and assumptions (U.S. EPA.
2005a. 33.1.41.
In the absence of study-specific data on,
for example, intake rates or body weight, the
EPA has developed recommended values for
use in dose-response analysis fU.S. EPA.
19881.
7.3. Modeling response in the range
of observation
Toxicodynamic ("biologically based")
modeling can incorporate data on biologic
processes leading to an effect. Such models
require sufficient data to ascertain a mode of
action and to quantitatively support model
parameters associated with its key events.
Because different models may provide
equivalent fits to the observed data but
diverge substantially at lower doses, critical
biologic parameters should be measured
from laboratory studies, not by model fitting.
Confidence in the use of a toxicodynamic
model depends on the robustness of its
validation process and on the results of
sensitivity analyses. Peer review of the
scientific basis and performance of a model is
essential (U.S. EPA. 2005a. 33.2.21.
Because toxicodynamic modeling can
require many parameters and more
knowledge and data than are typically
available, the EPA has developed a standard
set of empirical ("curve-fitting") models
(http://www.epa.gOv/ncea/bmds/l that can
be applied to typical data sets, including those
that are nonlinear. The EPA has also
developed guidance on modeling dose-
response data, assessing model fit, selecting
suitable models, and reporting modeling
results (U.S. EPA. 2012b). Additional
judgment or alternative analyses are used if
the procedure fails to yield reliable results,
for example, if the fit is poor, modeling may
be restricted to the lower doses, especially if
there is competing toxicity at higher doses
fU.S. EPA. 2005a. 33.2.31.
Modeling is used to derive a point of
departure ("U.S. EPA. 2012b: 2005a. 33.2.41.
(See Section 7.6 for alternatives if a point of
departure cannot be derived by modeling.):
If linear extrapolation is used,
selection of a response level
corresponding to the point of
departure is not highly influential, so
standard values near the low end of
the observable range are generally
used (for example, 10% extra risk for
cancer bioassay data, 1% for
epidemiologic data, lower for rare
cancers).
For nonlinear approaches, both
statistical and biologic considerations
are taken into account.
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For dichotomous data, a response
level of 10% extra risk is generally
used for minimally adverse effects,
5% or lower for more severe effects.
For continuous data, a response level
is ideally based on an established
definition of biologic significance. In
the absence of such definition, one
control standard deviation from the
control mean is often used for
minimally adverse effects, one-half
standard deviation for more severe
effects.
The point of departure is the 95% lower
bound on the dose associated with the
selected response level.
7.4. Extrapolating to lower doses and
response levels
The purpose of extrapolating to lower
doses is to estimate responses at exposures
below the observed data. Low-dose
extrapolation, typically used for cancer data,
considers what is known about modes of
action fU.S. EPA. 2005a. S3.3.1 and S3.3.21.
1) If a biologically based model has been
developed and validated for the agent,
extrapolation may use the fitted model
below the observed range if significant
model uncertainty can be ruled out with
reasonable confidence.
2) Linear extrapolation is used if the dose-
response curve is expected to have a
linear component below the point of
departure. This includes:
- Agents or their metabolites that are
DNA-reactive and have direct
mutagenic activity.
- Agents or their metabolites for which
human exposures or body burdens
are near doses associated with key
events leading to an effect
Linear extrapolation is also used when
data are insufficient to establish mode of
action and when scientifically plausible.
The result of linear extrapolation is
described by an oral slope factor or an
inhalation unit risk, which is the slope of
the dose-response curve at lower doses
or concentrations, respectively.
3) Nonlinear models are used for
extrapolation if there are sufficient data
to ascertain the mode of action and to
conclude that it is not linear at lower
doses, and the agent does not
demonstrate mutagenic or other activity
consistent with linearity at lower doses.
Nonlinear approaches generally should
not be used in cases where mode of action
has not ascertained. If nonlinear
extrapolation is appropriate but no
model is developed, an alternative is to
calculate reference values.
4) Both linear and nonlinear approaches
may be used if there a multiple modes of
action. For example, modeling to a low
response level can be useful for
estimating the response at doses where a
high-dose mode of action would be less
important
If linear extrapolation is used, the
assessment develops a candidate slope factor
or unit risk for each suitable data set. These
results are arrayed, using common dose
metrics, to show the distribution of relative
potency across various effects and
experimental systems. The assessment then
derives or selects an overall slope factor and
an overall unit risk for the agent, considering
the various dose-response analyses, the
study preferences discussed in Section 6, and
the possibility of basing a more robust result
on multiple data sets.
7.5. Considering susceptible
populations and lifestages
The assessment analyzes the available
information on populations and lifestages
that may be particularly susceptible to each
effect. A tiered approach is used fU.S. EPA.
2005a. $3.51
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1) If an epidemiologic or experimental study
reports quantitative results for a
susceptible population or lifestage, these
data are analyzed to derive separate
toxicity values for susceptible
individuals.
2) If data on risk-related parameters allow
comparison of the general population and
susceptible individuals, these data are
used to adjust the general-population
toxicity values for application to
susceptible individuals.
3) In the absence of chemical-specific data,
the EPA has developed age-dependent
adjustment factors for early-life exposure
to potential carcinogens that have a
mutagenic mode of action. There is
evidence of early-life susceptibility to
various carcinogenic agents, but most
epidemiologic studies and cancer
bioassays do not include early-life
exposure. To address the potential for
early-life susceptibility, the EPA
recommends (U.S. EPA. 2005b. §51:
10-fold adjustment for exposures
before age 2 years.
3-fold adjustment for exposures
between ages 2 and 16 years.
7.6. Reference values and uncertainty
factors
An oral reference dose or an inhalation
reference concentration is an estimate of an
exposure (including in susceptible
subgroups) that is likely to be without an
appreciable risk of adverse health effects
over a lifetime fU.S. EPA. 2002. §4.21
Reference values are typically calculated for
effects other than cancer and for suspected
carcinogens if a well characterized mode of
action indicates that a necessary key event
does not occur below a specific dose.
Reference values provide no information
about risks at higher exposure levels.
The assessment characterizes effects that
form the basis for reference values as
adverse, considered to be adverse, or a
precursor to an adverse effect For
developmental toxicity, reproductive toxicity,
and neurotoxicity there is guidance on
adverse effects and their biologic markers
(TJ.S. EPA. 1998.1996. 19911.
To account for uncertainty and variability
in the derivation of a lifetime human
exposure where adverse effects are not
anticipated to occur, reference values are
calculated by applying a series of uncertainty
factors to the point of departure. If a point of
departure cannot be derived by modeling, a
no-observed-adverse-effect level or a lowest-
observed-adverse-effect level is used instead.
The assessment discusses scientific
considerations involving several areas of
variability or uncertainty.
Human variation. The assessment accounts
for variation in susceptibility across the
human population and the possibility
that the available data may not be
representative of individuals who are
most susceptible to the effect A factor of
10 is generally used to account for this
variation. This factor is reduced only if
the point of departure is derived or
adjusted specifically for susceptible
individuals (not for a general population
that includes both susceptible and non-
susceptible individuals) (U.S. EPA.
2002. §4.4.5: 1998. §4.2: 1996. §4:
1994. §4.3.9.1: 1991. §3.41.
Animal-to-human extrapolation. If animal
results are used to make inferences about
humans, the assessment adjusts for
cross-species differences. These may
arise from differences in toxicokinetics or
toxicodynamics. Accordingly, if the point
of departure is standardized to
equivalent human terms or is based on
toxicokinetic or dosimetry modeling, a
factor of 101/2 (rounded to 3) is applied to
account for the remaining uncertainty
involving toxicokinetic and
toxicodynamic differences. If a
biologically based model adjusts fully for
toxicokinetic and toxicodynamic
differences across species, this factor is
not used. In most other cases, a factor of
10 is applied ("U.S. EPA. 2011:
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2002. §4.4.5: 1998. §4.2: 1996. §4:
1994. §4.3.9.1: 1991. §3.41.
Adverse-effect level to no-observed-
adverse-effect level. If a point of
departure is based on a lowest-observed-
adverse-effect level, the assessment must
infer a dose where such effects are not
expected. This can be a matter of great
uncertainty, especially if there is no
evidence available at lower doses. A
factor of 10 is applied to account for the
uncertainty in making this inference. A
factor other than 10 may be used,
depending on the magnitude and nature
of the response and the shape of the dose-
response curve (U.S. EPA. 2002. §4.4.5:
1998. §4.2: 1996. §4: 1994. §4.3.9.1:
1991. §3.41.
Subchronic-to-chronic exposure. If a point
of departure is based on subchronic
studies, the assessment considers
whether lifetime exposure could have
effects at lower levels of exposure. A
factor of 10 is applied to account for the
uncertainty in using subchronic studies
to make inferences about lifetime
exposure. This factor may also be applied
for developmental or reproductive effects
if exposure covered less than the full
critical period. A factor other than 10
may be used, depending on the duration
of the studies and the nature of the
response fU.S. EPA. 2002. §4.4.5: 1998.
§4.2: 1994. §4.3.9.11.
Incomplete database. If an incomplete
database raises concern that further
studies might identify a more sensitive
effect, organ system, or lifestage, the
assessment may apply a database
uncertainty factor fU.S. EPA. 2002. §4.4.5:
1998. §4.2: 1996. §4: 1994. §4.3.9.1:
1991. §3.41. The size of the factor
depends on the nature of the database
deficiency. For example, the EPA
typically follows the suggestion that a
factor of 10 be applied if both a prenatal
toxicity study and a two-generation
reproduction study are missing and a
factor of 101/2 if either is missing (U.S.
EPA. 2002. §4.4.51.
In this way, the assessment derives
candidate values for each suitable data set
and effect that is credibly associated with the
agent. These results are arrayed, using
common dose metrics, to show where effects
occur across a range of exposures (U.S. EPA.
1994. §4.3.91.
The assessment derives or selects an
organ- or system-specific reference value for
each organ or system affected by the agent.
The assessment explains the rationale for
each organ/system-specific reference value
(based on, for example, the highest quality
studies, the most sensitive outcome, or a
clustering of values). By providing these
organ/system-specific reference values, IRIS
assessments facilitate subsequent cumulative
risk assessments that consider the combined
effect of multiple agents acting at a common
site or through common mechanisms fNRC.
20091.
The assessment then selects an overall
reference dose and an overall reference
concentration for the agent to represent
lifetime human exposure levels where effects
are not anticipated to occur. This is generally
the most sensitive organ/system-specific
reference value, though consideration of
study quality and confidence in each value
may lead to a different selection.
7.7. Confidence and uncertainty in the
reference values
The assessment selects a standard
descriptor to characterize the level of
confidence in each reference value, based on
the likelihood that the value would change
with further testing. Confidence in reference
values is based on quality of the studies used
and completeness of the database, with more
weight given to the latter. The level of
confidence is increased for reference values
based on human data supported by animal
data fU.S. EPA. 1994. §4.3.9.21.
High confidence: The reference value is not
likely to change with further testing,
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except for mechanistic studies that might
affect the interpretation of prior test
results.
Medium confidence: This is a matter of
judgment, between high and low
confidence.
Low confidence: The reference value is
especially vulnerable to change with
further testing.
These criteria are consistent with
guidelines for systematic reviews that
evaluate the quality of evidence. These also
focus on whether further research would be
likely to change confidence in the estimate of
effect fGuvatt etal.. 2008b).
All assessments discuss the significant
uncertainties encountered in the analysis.
The EPA provides guidance on
characterization of uncertainty (U.S. EPA.
2005a. §3.61 For example, the discussion
distinguishes model uncertainty (lack of
knowledge about the most appropriate
experimental or analytic model) and
parameter uncertainty (lack of knowledge
about the parameters of a model).
Assessments also discuss human variation
(interpersonal differences in biologic
susceptibility or in exposures that modify the
effects of the agent).
August 2013
This document is a draft for review purposes only and does not constitute Agency policy.
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EXECUTIVE SUMMARY
Summary of Occurrence and Health Effects
Hexahydro-l,3,5-trinitro-l,3,5-triazine (RDX) is a synthetic chemical used
primarily as a military explosive. RDX releases have been reported in air, water, and
soil. Exposure to RDX is likely limited to individuals in or around military facilities
where RDX is or was produced, used, or stored. Oral exposure may occur from
drinking contaminated groundwater or ingesting crops irrigated with contaminated
water. Inhalation or dermal exposures are more likely in occupational settings.
Epidemiological studies provide only limited information on worker
populations exposed to RDX; several case reports describe effects primarily in the
nervous system following acute exposure to RDX. Animal studies of ingested RDX
demonstrate toxicity, including nervous system effects, kidney and other urogenital
effects, and male reproductive effects.
Results from animal studies provide suggestive evidence of carcinogenic
potential for RDX based on evidence of positive trends in liver and lung tumor
incidence in experimental animals. There are no data on the carcinogenicity of RDX
in humans.
Effects Other Than Cancer Observed Following Oral Exposure
Nervous system effects are a human hazard of RDX exposure. Several human case reports
and animal studies provide consistent evidence of an association between RDX exposure and effects
on the nervous system, including seizures or convulsions, tremors, hyperirritability, hyper-
reactivity, and behavioral changes. Mechanistic data support the hypothesis that RDX-induced
hyperactivity and seizures likely result from inhibition of GABAergic signaling in the limbic system.
Kidney and other urogenital effects are a potential human hazard of RDX exposure based on
observations in 2-year oral toxicity studies of increased relative kidney weights in male and female
mice and histopathological changes in the urogenital system of male rats exposed to RDX. An
increased incidence of suppurative prostatitis was identified, and is considered a marker for RDX-
related urogenital effects. There is no established mode of action (MOA) for RDX-related effects on
the urogenital system.
There is suggestive evidence of male reproductive effects associated with RDX exposure
based on the finding of testicular degeneration in male mice exposed to RDX in the diet for 2 years,
in the only mouse study conducted of that duration. There is no known MOA for male reproductive
effects of RDX exposure. Evidence for effects on other organs/systems, including the liver and
developmental effects, was more limited than for the endpoints summarized above.
This document is a draft for review purposes only and does not constitute Agency policy.
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Oral Reference Dose (RfD) for Effects Other Than Cancer
Organ-specific RfDs were derived for hazards associated with RDX exposure (see
Table ES-1). These organ- or system-specific reference values may be useful for subsequent
cumulative risk assessments that consider the combined effect of multiple agents acting at a
common site.
Table ES-1. Organ/system-specific RfDs and overall RfD for RDX
Effect
Basis
RfD (mg/kg-d)
Study exposure
description
Confidence
Nervous system
Convulsions
3 x 10"3
Subchronic
Medium
Kidney/urogenital
Suppurative prostatitis
8 x 10"3
Chronic
Low
Male reproductive
Testicular degeneration
8 x 10"2
Chronic
Low
Overall RfD
Nervous system effects
3 x 10"3
Subchronic
Medium
The overall RfD (see Table ES-2) is derived to be protective of all types of hazards
associated with RDX exposure. The effect of RDX on the nervous system was chosen as the basis for
the overall RfD because nervous system effects were observed most consistently across studies,
species, and exposure durations, and because they represent the most sensitive human hazard of
RDX exposure. Incidence of seizures or convulsions as reported in a subchronic gavage study
(Grouse etal.. 20061 was selected for derivation of the overall RfD as this endpointwas measured in
a study that was well-conducted, utilized a test material of higher purity than other studies, and had
five closely-spaced dose groups that allowed characterization of the dose-response curve.
Benchmark dose (BMD) modeling was utilized to derive the point of departure (POD) for RfD
derivation (expressed as the BMDLoi). A 1% response level was chosen because of the severity of
the endpoint. Further, the doses associated with nervous system effects in the Grouse etal. (2006)
study also caused increased mortality in the animals. Experimental animal studies provide some
evidence of an association between RDX-induced convulsions and mortality. In three studies in rats
f Grouse etal.. 2006: Levine etal.. 1983b: Cholakis etal.. 19801. investigators noted that early deaths
were frequently preceded by neurotoxic signs such as tremors and convulsions; however, the
recorded data from Grouse etal. (2006) do not show as clear a correspondence between
convulsions (and other neurotoxic signs) and mortality (Section 1.2.1).
A physiologically-based pharmacokinetic (PBPK) model was used to extrapolate the
BMDLoi derived from a rat study to a human equivalent dose (HED) based on RDX arterial blood
concentration, which was then used for RfD derivation.
The overall RfD was calculated by dividing the BMDLoi-hed for nervous system effects by a
composite uncertainty factor (UF) of 100 to account for extrapolation from animals to humans (3),
interindividual differences in human susceptibility (10), and uncertainty in the database (3).
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1 Table ES-2. Summary of reference dose (RfD) derivation
Critical effect
Point of departure3
UF
Chronic RfD
Nervous system effects (convulsions)
90-d F344 rat study
Crouse et al. (2006)
BMDLoi-hed: 0.28 mg/kg-d
100
3 x 10"3 mg/kg-d
2
3 aA benchmark response (BMR) of 1% was used to derive the BMD and BMDL given the severity of the endpoint.
4 The resulting POD was converted to a BMDLoi-hed using a PBPK model based on modeled arterial blood
5 concentration. The concentration was derived from the area under the curve (AUC) of modeled RDX
6 concentration in arterial blood, which reflects the average blood RDX concentration for the exposure duration
7 normalized to 24 hours.
Effects Other Than Cancer Observed Following Inhalation Exposure
8 No studies were identified that provided useful information on the effects observed
9 following inhalation exposure to RDX. Of the available human epidemiological studies of RDX, none
10 provided data that could be used for dose-response analysis of inhalation exposures. The single
11 experimental animal study involving inhalation exposure is not publicly available, and was
12 excluded from consideration due to significant study limitations, including small numbers of
13 animals tested, lack of controls, and incomplete reporting of exposure levels. Therefore, the
14 available health effects literature does not support the identification of hazards following inhalation
15 exposure to RDX.
Inhalation Reference Concentration (RfC) for Effects Other Than Cancer
16 An RfC for RDX could not be derived based on the available health effects data. While
17 inhalation absorption of RDX particulates is a plausible route of exposure, there are no toxicokinetic
18 studies of RDX inhalation absorption to support an inhalation model. Therefore, a PBPK model for
19 inhaled RDX was not developed to support route-to-route extrapolation of an RfC from the RfD.
20 Evidence for Human Carcinogenicity
21 Under EPA's cancer guidelines fU.S. EPA. 2005al. there is suggestive evidence of carcinogenic
22 potential for RDX. RDX induced benign and malignant tumors in the liver and lungs of mice (Parker
23 etal.. 2006: Lish etal.. 19841 or rats (Levine etal.. 1983b) following long-term administration in the
24 diet The potential for carcinogenicity applies to all routes of human exposure.
25 Quantitative Estimate of Carcinogenic Risk from Oral Exposure
26 A quantitative estimate of carcinogenic risk from oral exposure to RDX was based on the
27 increased incidence of hepatocellular adenomas or carcinomas and alveolar/bronchiolar adenomas
28 or carcinomas in female B6C3Fi mice observed in the carcinogenicity bioassay in mice (Lish etal..
29 19841. This 2-year dietary study included four dose groups and a control group, adequate numbers
30 of animals per dose group (85/sex/group, with interim sacrifices of 10/sex/group at 6 and
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12 months), and detailed reporting of methods and results (including individual animal data). The
initial high dose (175 mg/kg-day) was reduced to 100 mg/kg-day at week 11 due to high mortality.
Considering these data along with the uncertainty associated with the suggestive nature of
the weight of the evidence for RDX carcinogenicity, quantitative analysis of the mouse tumor data
may be useful for providing a sense of the magnitude of potential carcinogenic risk.
An oral slope factor (OSF) that considered the combination of female mouse liver and lung
tumors was derived from BMD and BMDL estimates that correspond to a 10% extra risk (ER) of
either tumor. The BMDLio so derived was extrapolated to the HED using BW3/4 scaling, and an OSF
was derived by linear extrapolation from the BMDLio hed. The OSF is 0.04 per mg/kg-day, based on
the liver and lung tumor response in female mice fLish etal.. 19841.
Quantitative Estimate of Carcinogenic Risk from Inhalation Exposure
An inhalation unit risk (IUR) value was not calculated because inhalation carcinogenicity
data for RDX are not available. While inhalation absorption of RDX particulates is a plausible route
of exposure, there are no toxicokinetic studies of RDX inhalation absorption to support an
inhalation model. Therefore, a PBPK model for inhaled RDX was not developed to support route-to-
route extrapolation of an IUR from the OSF. Thus, a quantitative cancer assessment was not
conducted.
Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes
Little information is available on populations that may be especially vulnerable to the toxic
effects of RDX. Lifestage, and in particular childhood, susceptibility has not been observed in
human or animal studies of RDX toxicity. In rats, transfer of RDX from the dam to the fetus during
gestation and to pups via maternal milk has been reported; however, reproductive and
developmental toxicity studies did not identify effects in offspring at doses below those that also
caused maternal toxicity. Data to suggest that males may be more susceptible than females to
noncancer toxicity associated with RDX exposure are limited. Specifically, urogenital effects have
been noted at lower doses in males than in females. Data on the incidence of convulsions and
mortality provide some indication that pregnant animals may be a susceptible population, although
the evidence is unclear. Some evidence suggests that cytochrome P450 (CYP450) enzymes may be
involved in the metabolism of RDX, indicating a potential for genetic polymorphisms in these
metabolic enzymes to affect susceptibility to RDX. Similarly, individuals with epilepsy or other
seizure syndromes that have their basis in genetic mutation to GABAa receptors may represent
another group that may be susceptible to RDX exposure; however, there is no information to
indicate how genetic polymorphisms may affect susceptibility to RDX.
Key Issues Addressed in Assessment
Selection of a 1 % benchmark response (BMRJfor convulsions. In most instances, the
spectrum of effects associated with chemical exposure will range in severity, with relatively less
This document is a draft for review purposes only and does not constitute Agency policy.
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severe effects generally occurring at doses lower than those associated with more severe or "frank"
toxicity. Convulsions in rats were selected as the basis for derivation of the RDX RfD; less severe
nervous system effects were generally not observed at lower doses. U.S. EPA f2012b) recommends
considering the statistical and biological characteristics of the dataset when selecting a BMR,
including the severity of the effect. For convulsions, a BMR level of 1% ER was selected for
modeling, balancing the quantitative limitations of the available animal bioassays and the severity
of this effect Modeling convulsion incidence from Grouse etal. (2006) using this BMR resulted in a
moderate extrapolation of the BMD (3.0 mg/kg-day) below the range of experimental data (dose
range from Grouse etal. (2006): 4-15 mg/kg-day).
Influence of the method of oral dosing (diet and gavage). Some uncertainty in the RfD is
also associated with the influence of the method of oral dosing on the magnitude of dose required
to induce nervous system effects. Findings from animal studies suggest that gavage administration
generally induced convulsions in experimental animals at lower doses than did dietary
administration, possibly due to the bolus dose received from gavage administration resulting in a
comparatively faster absorption and higher peak blood concentrations of RDX (see Section 1.2.1).
The difference in neurotoxic response associated with gavage versus dietary administration is in
part reflected in the 14-fold difference in the candidate PODhed values derived from the Grouse et al.
(2006) (gavage administration) and Levine etal. (1983b) (dietaryadministration) studies (see
Table 2-2). A more rigorous examination of the effect of oral dosing method cannot be performed
because of the differences across studies in test materials and experimental designs (e.g., test article
purity and particle size, number and spacing of dose groups, exposure duration, frequency of
clinical observations, and thoroughness of the reporting of observations) that could also have
contributed to differences in response. As dietary administration is more representative of
potential human exposures to RDX, the use of toxicity data from a gavage (bolus dosing) study may
introduce uncertainty in the RfD.
Suppurative prostatitis as a marker for kidney and other urogenital effects. The
candidate RfD for kidney and other urogenital effects is based on a dose-related increase in the
incidence of suppurative prostatitis from a 2-year feeding study in male F344 rats (Levine etal..
1983b). This study is the only 2-year study in rats that examined the prostate. Some reports have
hypothesized that the observed suppurative prostatitis is a secondary effect from a bacterial
infection unrelated to RDX toxicity fATSDR. 2012: Sweeney etal.. 2012a: Grouse etal.. 20061. While
an opportunistic bacterial infection may have been the proximal cause of the suppurative
prostatitis, the infection was considered secondary to urogenital effects associated with RDX
exposure. Histopathological findings for the bladder are not definitive because the design of the
principal study called for histopathological examination of the bladder only if gross abnormalities
were observed. Although the pathogenesis of kidney and urogenital effects is unclear, suppurative
prostatitis was considered to be a marker for the broader array of kidney and other urogenital
effects observed by Levine etal. Q983bl
This document is a draft for review purposes only and does not constitute Agency policy.
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LITERATURE SEARCH STRATEGY |
STUDY SELECTION AND EVALUATION
Literature Search and Screening Strategy
A literature search and screening strategy was applied to identify literature related to
characterizing the health effects of hexadydro-l,3,5-trinitro-l,3,5-triazine (RDX). This strategy
consisted of a search of online scientific databases and other sources, casting a wide net in order to
identify all potentially pertinent studies. In subsequent steps, references were screened to exclude
papers not pertinent to an assessment of the chronic health effects of RDX, and the remaining
references were sorted into categories for further evaluation.
The literature search for RDX was conducted in four online scientific databases—PubMed,
Toxline, Toxcenter, and Toxic Substances Control Act Test Submissions (TSCATS). The initial
search was performed in April 2012, and literature search updates were conducted in February
2013, January 2014, and January 2015. Searches of TSCATS were performed in February 2013 and
January 2015 only. The detailed search approach for these databases, including the query strings,
and the numbers of citations identified per database are provided in Appendix B, Table B-l. The
Department of Defense has conducted several unpublished toxicological studies on RDX; to ensure
that all such studies were located, the Defense Technical Information Center (DTIC) database, a
central online repository of defense-related scientific and technical information within the
Department of Defense, was also searched. A separate strategy was applied in searching DTIC
because of limitations in the classification and distribution of materials in DTIC; the detailed search
strategy is described in Appendix B, Table B-2. Searches of the five online databases identified
1,143 citations (after electronically eliminating duplicates). The computerized database searches
were supplemented by reviewing online regulatory sources, performing "forward" and "backward"
searches of Web of Science (see Appendix B, Table B-3), and adding additional references that were
identified during the development of the Toxicological Review (including submissions from the
Department of Defense); 34 citations were obtained using these additional search strategies. In
total, 1,177 citations were identified using online scientific databases and additional search
strategies.
The U.S. Environmental Protection Agency (EPA) requested public submissions of
additional information in 2010 (75 FR 76982; December 10, 2010). EPA also issued a request to
the public for additional information in a Federal Register Notice in 2013 (78 FR 48674; August 9,
2013), and established a docket for public comment (EPA-HQ-ORD-2013-0430; available at
www.regulations.govl maintained through the development of the assessment No submissions
were received in response to these calls for data.
This document is a draft for review purposes only and does not constitute Agency policy.
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The citations identified using the search strategy described above were screened using the
title, abstract, and in limited instances, full text for pertinence to examining the health effects of
chronic RDX exposure. The process for screening the literature is described below and is shown
graphically in Figure LS-1.1 The objective of this manual screen was to identify sources of primary
human health effects data and sources of primary data that inform the assessment of RDX health
effects (i.e., the bottom three boxes in Figure LS-1). Inclusion and exclusion criteria used to
manually screen the references in order to identify health effect studies (i.e., the green boxes in
Figure LS-1) are provided in Table LS-1. Specific inclusion criteria were not applied in identifying
sources of mechanistic and toxicokinetic data. The number of such studies for RDX is not large, and
therefore, all studies that provided data on adsorption, distribution, metabolism, or elimination,
physiologically-based pharmacokinetic (PBPK) models, or relevant RDX mode of action (MOA)
were considered. Studies that met one or more of the exclusion criteria in Table LS-1 were binned
as "Excluded/Not on Topic" and were not further considered in this assessment A final group of
studies consisted of reviews and other sources of RDX information (e.g., exposure, ecosystem
effects) that did not meet the inclusion criteria in Table LS-1. These studies were binned into a
category called "Secondary Literature and Sources of Other RDX Information," and were considered
as appropriate during development of this assessment.
The results of this literature screening are described below and graphically in Figure LS-1:
• 25 references (including both human and animal studies) were identified as sources of
health effects data and were considered for data extraction to evidence tables and
exposure-response arrays.2
• 25 references were identified as sources of supplementary health effects data, including
experimental animal studies involving acute or short-term exposures or dermal
exposure, and human case reports. Studies investigating the effects of acute/short-term
and dermal exposures and case reports are generally less pertinent for characterizing
health hazards associated with chronic oral and inhalation exposure. Therefore,
information from these studies was not extracted into evidence tables. Nevertheless,
these studies were still considered as possible sources of supplementary health effects
information.
'Studies were assigned (or "tagged") to a given category in Health and Environmental Research Online
(HERO) that best reflected the primary content of the study. In general, studies were not assigned multiple
tags in order to simplify the tracking of references. Nevertheless, the inclusion of a citation in a given
category (or tag) did not preclude its use in one or more other categories. For example, Woody et al. (1986). a
case report of accidental ingestion of RDX by a child, was tagged to the human case reports under
Supplementary Studies in Figure LS-1. This case report also provides pharmacokinetic data and was a
pertinent source of information on RDX toxicokinetics, but was not assigned a second tag for toxicokinetics.
2 On the HERO project page, 27 records are associated with "Sources of Health Effects Data," rather than the
25 identified in Figure LS-1. Two of the records in HERO are not unique studies or reports. Rather, these two
records provide links to the multi-volume laboratory reports for the 2-year studies in rats by Levine et al.
fl983bl and mice by Lish etal. f 19841
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47 references were identified as sources of mechanistic and toxicokinetic data; these
included 19 studies describing PBPK models and other toxicokinetic information,
11 studies providing genotoxicity information, and 17 studies pertaining to other
mechanistic information. Information from these studies was not extracted into
evidence tables; however, these studies supplemented the assessment of RDX health
effects (e.g., evaluation of MOA and extrapolation of experimental animal findings to
humans).
280 references were identified as secondary literature (e.g., reviews and other agency
assessments) or as studies providing potentially useful information on RDX (e.g., studies
providing information on exposure levels or effects on nonmammalian species); these
references were kept as additional resources for development of the Toxicological
Review.
801 references were identified as not being pertinent (or not on topic) to an evaluation
of the chronic health effects of RDX and were excluded from further consideration (see
Figure LS-1 and Table LS-1 for exclusion criteria). Retrieving a large number of
references that are not on topic is a consequence of applying an initial search strategy
designed to cast a wide net and to minimize the possibility of missing potentially
relevant health effects data.
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<---
Supplementary Studies
Combined Dataset
n=l,177
Other Sources of Supplementary
Health Effects Data (n=26)
9 Acute/short-term animal studies
16 Human case reports
Sources of Health Effects Data
(n=25)
4 Human health effects studies
21 Animal toxicology studies
Additional Search Strategies
(see Table B-3 for methods and results)
n=34
Sources of Mechanistic and
Toxicokinetic Data (n=47)
11 Genotoxicity studies
17 Other mechanistic studies
19 Toxicokinetic studies
Manual Screening For Pertinence
(Title/Abstract/Full Text)
Secondary Literature and Sources of
Other RDX Information (n=280)
91 Ecosystem effects
113 Exposure levels
8 Mixtures only
15 Regulatory documents
51 Reviews; risk assessments; editorials
2 Other
Excluded/Not on Topic (n=801)
32 Abstract only; inadequate reporting in
abstract; no abstract
231 Treatment/remediation
126 Chemical, physical or explosive
properties
235 Laboratory methods
155
Not chemical specific
22 Other
Pubmed
n=591
Toxline
n=434
Database Searches
(see Tables B-l and B-2 for keywords and limits)
(After duplicates removed electronically)
n=l,143
Toxcenter
n=32
DTIC
n=85
TSCATS l/2/8e
n=2
The numbers on this figure match the HERO project page as of 1/12/2016; subsequent changes may not be
reflected, A limited number of references were assigned more than one tag; therefore, the sum of the references
in boxes below "Manual Screening for Pertinence" does not match exactly the total number of references in the
"Combined Dataset."
Figure LS-1. Summary of literature search and screening process for RDX.
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Table LS-1. Inclusion-exclusion criteria for health effect studies
Inclusion criteria
Exclusion criteria
Population
• Humans
• Standard mammalian animal models,
including rat, mouse, rabbit, guinea pig,
monkey, dog
• Ecological species3
• Nonmammalian species3
Exposure
• Exposure is to RDX
• Exposure is measured in an
environmental medium (e.g., air, water,
diet)
• Exposure via oral or inhalation routes
• Study population is not exposed to RDX
• Exposure to a mixture only
• Exposure via injection (e.g., intravenous [i.v.])
Outcome
• Study includes a measure of one or
more health effect endpoints, including
effects on the nervous,
kidney/urogenital, musculoskeletal,
cardiovascular, immune, and
gastrointestinal systems, reproduction,
development, liver, eyes, and cancer
Other
Not on topic, including:
• Abstract only, inadequately reported abstract, or no
abstract, and not considered further because study
was not potentially relevant
• Bioremediation, biodegradation, or chemical or
physical treatment of RDX and other munitions,
including evaluation of wastewater treatment
technologies and methods for remediation of
contaminated water and soil
• Chemical, physical, or explosive properties, including
studies of RDX crystal quality, energetics
characteristics, sublimation kinetics, isotope ratios,
and thermal decomposition and other explosive
properties
• Analytical methods for measuring/detecting/
remotely sensing RDX in environmental media, and
use in sample preparations and assays
• Not chemical specific (studies that do not involve
testing of RDX)
• Other studies not informative for evaluating RDX
health effects and not captured by other exclusion
criteria, including:
- Superfund site records of decision that describe
remedial action plans for waste sites
- characterization of waste sites contaminated by
explosives
-- foreign language studies where translation was not
warranted because, based on title or abstract, the
added value to the evaluation of RDX health effects
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Inclusion criteria
Exclusion criteria
was considered small (e.g., Chinese paper of case
reports of RDX poisonings)
-- duplicate studies not previously identified
aStudies that met this exclusion criterion were not considered a source of health effects or supplementary health
effects data, but were considered as other sources of information potentially useful in assessing the health effects
of RDX.
The documentation and results for the literature search and screen can be found on the
Health and Environmental Research Online (HERO) website on the RDX project page at:
(http://hero.epa.gov/index.cfm/proiect/page/proiect id/22161.
Selection of Critical Studies for Presentation in Evidence Tables
Selection of Critical Studies
In order to systematically summarize the important information from the primary health
effects studies in the RDX database, evidence tables were constructed in a standardized tabular
format as recommended by the NRC (2011). Of the studies that were retained after the literature
search and screen, 25 were categorized as "Sources of Health Effects Data" (Figure LS-1, Table LS-1)
and were considered for extraction into evidence tables for hazard identification in Chapter 1.
A study was not presented in the evidence tables if flaws in its design, conduct, or reporting
were so great that the results would not be considered credible (e.g., studies where concurrent
control information is lacking). Such study design flaws are discussed in a number of EPA's
guidelines (see http: //www.epa.gov/iris/backgrd.htmn or summarized in the Preamble. For RDX,
four studies were considered uninformative and were removed from further consideration in the
assessment because of fundamental issues with study design, conduct, or reporting. The specific
studies and basis for considering the studies to be uninformative are summarized in Table LS-2.
This document is a draft for review purposes only and does not constitute Agency policy.
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Table LS-2. Studies determined not to be informative because of significant
issues with design, conduct, or reporting
Reference
Rationale for exclusion
Haskell Laboratories (1942);
14-wk study in dogs
Incomplete information on exposure levels; breed of dog was not
reported; inadequate reporting of results; sections of document
were illegible.
von Oettingen et al. (1949);
10-wk oral study in rats
No control group; strain of rat was not reported.
ATSDR (1996);
Disease prevalence study in residential
population
Study of a population residing in two neighborhoods where RDX
had been detected in well water. The study was conducted 7 yrs
after residents were provided the opportunity to connect to a
municipal water supply. Only one target-area household reported
using private well water for bathing and cooking at the time of the
health study. The study was not considered informative because
the design was not able to adequately define the exposed
population.
Unpublished report (dated 1944) from the
DTIC database;
Human and animal data
One section of the report describes a human case series with no
referent group. Issues with the inhalation experimental animal
studies included lack of control groups, incomplete information on
exposure levels, and inadequate reporting of results. [Because this
report is classified as a limited distribution document in the DTIC
database, it was not added to the HERO project page for RDX.]
The health effects literature for RDX is not extensive. With the exception of the studies
listed in Table LS-2 (i.e., those determined to be uninformative), all human and experimental animal
studies of RDX involving repeated exposure were considered in assessing the evidence for health
effects associated with chronic exposure to RDX.
Studies that contain pertinent information for the toxicological review and augment hazard
identification conclusions, such as genotoxicity and other mechanistic studies, studies describing
the toxicokinetics of RDX, human case reports, and experimental animal studies involving
exposures of acute/short-term duration or routes of exposure other than oral and inhalation, were
not included in evidence tables. Nevertheless, these studies were considered, where relevant, in the
evaluation of RDX health hazards.
This document is a draft for review purposes only and does not constitute Agency policy.
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Study Evaluation
For this assessment, primary sources of health effects data consisted of three human
studies3 and 21 reports4 presenting results of experimental animal studies. These studies were
evaluated using the study quality considerations outlined in the Preamble, considering aspects of
design, conduct, or reporting that could affect the interpretation of results, overall contribution to
the synthesis of evidence, and determination of hazard potential as noted in various EPA guidance
documents (U.S. EPA. 2005a. 2002.19941. The objective was to identify the stronger, more
informative studies based on a uniform evaluation of quality characteristics across studies of
similar design.
Additionally, a number of general questions, presented in Table LS-3, were considered in
evaluating the animal studies. Much of the key information for conducting this evaluation can be
determined based on study methods and how the study results were reported. Importantly, the
evaluation at this stage does not consider the direction or magnitude of any reported effects.
Table LS-3. Considerations and relevant experimental information for
evaluation of experimental animal studies
Methodological
feature
Considerations
(relevant information extracted into evidence tables)
Test animal
Suitability of the species, strain, sex, and source of the test animals
Experimental design
Suitability of animal age/lifestage at exposure and endpoint testing; periodicity and
duration of exposure (e.g., hrs/d, d/wk); timing of endpoint evaluations; and sample size
and experimental unit (e.g., animals, dams, litters)
Exposure
Characterization of test article source, composition, purity, and stability; suitability of the
control (e.g., vehicle control); documentation of exposure techniques (e.g., route,
chamber type, gavage volume); verification of exposure levels (e.g., consideration of
homogeneity, stability, analytical methods)
Endpoint evaluation
Suitability of specific methods for assessing the endpoint(s) of interest
Results presentation
Data presentation for endpoint(s) of interest (including measures of variability) and for
other relevant endpoints needed for results interpretation (e.g., maternal toxicity,
decrements in body weight in relation to organ weight)
3Two reports with human data were determined not to be informative; see Table LS-2. The study by ATS PR
(19961 was included in HERO and in Figure LS-1. The unpublished report from the DTIC database was not
included in either HERO or Figure LS-1 because this report is classified as a limited distribution document in
DTIC. This accounts for the three human studies being reviewed for study evaluation rather than the four
identified in the literature search (see Figure LS-11.
4The number of reports of experimental animal studies does not equal the number of studies. The results of
some studies were documented in multiple reports (e.g., a 2-year study in F344 rats by Levine et al. f 1983bl
was published in three volumesl. The Cholakis et al. (19801 study included, in a single report, subchronic
studies in rats and mice, a 2-generation reproductive toxicity study in rats, and developmental toxicity
studies in rats and rabbits.
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Information on study features related to this evaluation is reported in evidence tables and
was considered in the synthesis of evidence. Discussion of study strengths and limitations (that
ultimately supported preferences for the studies and data relied upon) were included in the text
where relevant If EPA's interpretation of a study differed from that of the study authors, the
assessment discusses the basis for the difference.
The general findings of this evaluation are presented in the remainder of this section. Study
evaluation considerations that are outcome specific are discussed in the relevant health effect
sections in Section 1.2.
Human Studies
The body of literature on RDX includes three studies of populations occupationally exposed
to RDX (one case-control and two cross-sectional studies) fWest and Stafford. 1997: Hathaway and
Buck. 1977). To varying degrees, these epidemiology studies are limited in their ability to assess
the relationship between RDX exposure and the incidence of human health effects. Some studies
lacked information related to study design, such as a precise definition of the study population,
while others did not include a comprehensive exposure assessment or details regarding potential
confounders. All three studies had small sample sizes (60-69 exposed workers in the cross-
sectional studies and 32 cases in the case-control study), which limits their statistical power when
comparing exposed workers or cases and unexposed or control participants.
The study by Ma and Li (1993) of Chinese industrial workers provided limited information
on participant recruitment, selection, and participation rate; the available information was not
adequate to evaluate the potential for selection bias. Also, no information on adjustment for co-
exposure to trinitrotoluene (TNT) or other neurological risk factors (e.g., alcohol consumption) was
provided. The study by Hathaway and Buck (1977) included details on exposure assessment, but
did not provide information on length of employment or other metrics that could be used to
ascertain duration of exposure. In the case-control study by West and Stafford T1997I RDX was
identified as one of the many chemicals that workers may have been exposed to in the ordnance
factory. Thus, there is a potential for co-exposure to other chemicals that may elicit the observed
effects. The methodological limitations in these three studies were considered in the synthesis of
evidence for each of the health effects and in reaching determinations of hazard (see Section 1.2).
In addition to the aforementioned studies, the human health effects literature includes
16 case reports that describe effects following acute exposure to RDX. Case reports can suggest
organ systems and health outcomes that might be related to RDX exposure but are often anecdotal,
and typically describe unusual or extreme exposure situations; thus, they provide little information
that would be useful for characterizing chronic health effects. Therefore, RDX case reports were
only briefly reviewed; a critical evaluation was not undertaken. A summary of these case reports is
provided in Appendix C, Section C.2.
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Experimental Animal Studies
The oral toxicity database for RDX includes three chronic studies in rats and mice, eight
subchronic studies in rats, mice, dogs, and monkeys, two shorter-term studies in dogs and rats, one
two-generation reproductive toxicity study in the rat, four developmental toxicity studies in rats
and rabbits, and a single-exposure study of audiogenic seizures in rats (Table LS-4).
Table LS-4. Summary of experimental animal database
Study category
Study duration, species/strain, and oral administration method
Chronic
2-Yr study in B6C3Fi mice (diet) (Lish et al., 1984)
2-Yr study in Sprague-Dawlev rats (diet) (Hart, 1976)
2-Yr study in F344 rats (diet) (Levine et al., 1983b)
Subchronic
13-Wk studv in B6C3Fi mice, experiment 1 (diet) (Cholakis et al., 1980)
13-Wk studv in B6C3Fi mice, experiment 2 (diet) (Cholakis et al., 1980)
13-Wk studv in F344 rats (diet) (Cholakis et al., 1980)
13-Wk studv in F344 rats (diet) (Levine et al., 1990; Levine et al., 1981a, b)
13-Wk studv in F344 rats (gavage) (Crouse et al., 2006)
13-Wk studv in rats, strain not specified (diet) (von Oettingen et al., 1949)
13-Wk studv in beagle dogs (diet) (Hart, 1974)
13-Wk studv in monkevs (gavage) (Martin and Hart, 1974)
6-Wk studv in dogs, breed not specified (diet) (von Oettingen et al., 1949)
30-D studv in Sprague-Dawlev rats (gavage) (MacPhail et al., 1985)
Reproductive
2-Generation reproductive toxicity studv in CD rats (diet) (Cholakis et al., 1980)
Developmental
Developmental studv (gestational davs fGDsl 6-19) in F344 rats (gavage) (Cholakis et al., 1980)
Developmental studv (GDs 6-15) in Sprague-Dawlev rats, range-finding (gavage) (Angerhofer et
al., 1986)
Developmental studv (GDs 6-15) in Sprague-Dawlev rats (gavage) (Angerhofer et al., 1986)
Developmental studv (GDs 7-29) in New Zealand White (NZW) rabbits (gavage) (Cholakis et al.,
1980)
Nervous system
8-Hr studv of audiogenic seizures in Long Evans rats (gavage) (Burdette et al., 1988)
With the exception of two studies (Levine etal.. 1990: von Oettingen et al.. 19491. these
toxicity studies are available only as unpublished contract laboratory reports. Peer reviews of
three unpublished studies identified as most informative to the assessment of the health effects of
RDX—the 2-year bioassays by Levine etal. Q983bl and Lish etal. f 19841 and the subchronic
toxicity study by Grouse etal. (2006)—were conducted by Versar, Inc. for EPA in 2012. The report
of the peer reviews (U.S. EPA. 2012dl is available at https://epa.gov/hero. The peer reviewers
generally concluded that the 2-year bioassay reports provided useful information on the toxicity of
RDX, noting that there were limitations in interpretation due to aspects of the histopathological
analysis and the statistical approaches employed. The peer reviewers similarly determined that the
report by Grouse etal. f2 0061 provided useful information on RDX toxicity, including an array of
endpoints for neurotoxicity and immunotoxicity fU.S. EPA. 2012dl.
Only one unpublished inhalation study of RDX (dated 1944) was identified. As discussed in
Appendix B and Table LS-2, this inhalation study was considered uninformative and was excluded
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from consideration in the development of the Toxicological Review because of study design issues
(including lack of a control group, incomplete information on exposure levels, and inadequate
reporting). Therefore, evaluation of the experimental animal database for RDX is limited to studies
of oral toxicity. An evaluation of the oral toxicity literature, organized by general methodological
features, is provided in the remainder of this section.
Test animal
The RDX database consists of health effect studies conducted in multiple strains of rats
(F344, Sprague-Dawley, CD), mice (B6C3Fi), dogs (beagle), and monkeys. The species and strains
of animals used are consistent with those typically used in laboratory studies. All of these species
or strains were considered relevant to assessing the potential human health effects of RDX. Several
studies in the RDX database provided inadequate information on test animals. The strain of
monkey (Rhesus or Cynomolgus) used in the study by Martin and Hart Q9741 was not clearly
specified. In one study, the breed of dog and strain of rat were unreported fvon Oettingen etal..
1949). The species, strain, and sex of the animals used are recorded in the evidence tables.
Other studies of RDX were identified that used nonstandard species, including deer mice
[Peromyscus maniculatus), western fence lizards (Sceloporus occidentalis), prairie voles (Microtus
ochrogaster), and northern bobwhite quail (Colinus virginianus). These studies provide information
relevant to RDX toxicokinetics and mechanism of action on the nervous system, but not health
effects data. Therefore, these studies are not included in evidence tables, but are discussed where
relevant in the assessment
Experimental setup
General aspects of study design and experimental setup were evaluated for all studies that
included health effects data to determine if they were appropriate for evaluation of specific
endpoints. Key features of the experimental setup, including the periodicity and duration of
exposure, timing of exposure (e.g., gestational days for developmental studies), experimental group
sample sizes, and interim sacrifices are summarized in the evidence tables. Note that sample size
was not a basis for excluding a study from consideration, as studies with a small number of animals
can still inform the consistency of effects observed for a specific endpoint Nevertheless, the
informativeness of studies with small sample sizes, e.g., three animals/sex/group in the case of Hart
(1974) and Martin and Hart (1974). was reduced. Elements of the experimental setup that could
influence interpretation of study findings are discussed in the relevant hazard identification
sections of the assessment
Exposure
Properties of the test material were also considered in determining whether the exposures
were sufficiently specific to the compound of interest Two properties of the RDX test materials
that varied across experimental animal studies and that were taken into consideration in evaluating
the evidence for RDX hazards are the particle size and purity of the test material. The purity of RDX
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used in health effects studies varied from 84 to 99.99%. The major contaminants were octahydro-
l,3,5,7-tetranitro-l,3,5,7-tetrazocine (HMX) and water, which are the primary contaminants of RDX
produced during manufacturing. The majority of studies used RDX with ~10% impurities; only
Grouse etal. (2006) used 99.99% pure RDX as a test material in their study. The toxicity of HMX
was assessed by the Integrated Risk Information System (IRIS) Program in 1988
fhttp://cfpub.epa.gov/ncea/iris2/chemicalLanding.cfm7siibstance nmbr=3111: histopathological
changes in the liver in male F344 rats and in the kidney in female rats were reported in a 13-week
feeding study. No chronic studies were available to evaluate the carcinogenicity of HMX. The
presence of the impurities introduces some uncertainty in attribution of toxicity to RDX. However,
consistency in the doses at which some toxic effects were seen across studies suggests that the
uncertainty associated with the use of less pure test materials may be relatively small. Evidence of
neurotoxic effects in the study with 99.99% pure RDX occurred at doses of 8-15 mg/kg-day;
studies with less pure RDX reported similar symptoms at doses >20 mg/kg-day. It should be noted
that the test materials employed in these studies (i.e., with ~10% impurities) are consistent with
the purity of RDX that would be released into the environment.
Differences in milling procedures used to generate the test material resulted in the use of
RDX of varying particle sizes across studies. Some studies utilized a test material with a relatively
fine particle size (majority of particles <66 [im in size), while others used a test material with
comparatively coarse particle size (~200 [im particle size). Differences in particle size across
studies could result in different rates of absorption of RDX into the blood stream, which could
account for differences in response observed across studies, including neurotoxicity. Information
on test material purity and particle size, as provided by study authors, is reported in the evidence
tables, and was considered in evaluating the toxicity of RDX. The lack of characterization of the test
material in the studies by Hart Q974I Hart Q976I and Martin and Hart f19741 was considered a
deficiency.
Endpoint evaluation procedures
Some methodological considerations used to evaluate studies of RDX toxicity are outcome
specific—in particular, effects on the nervous system and development. Outcome-specific
methodological considerations are discussed in the relevant health effect sections in Section 1.2.
For example, many of the studies that noted neurotoxicity in the form of seizures or convulsions
were not designed to assess that specific endpoint and reported the number of animals with
seizures as part of clinical observations that, in general, were recorded only once daily. This
frequency of observations could have missed neurobehavioral events. While these studies can
provide qualitative evidence of neurotoxicity, they may have underestimated the true incidence of
seizures or convulsions because they were not designed to systematically evaluate neurotoxic
outcomes.
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Outcomes and data reporting
In evaluating studies, consideration was given to whether data were reported for all
endpoints specified in the methods section and for all study groups, and whether any data were
excluded from presentation or analysis. For example, it was noted where histopathological analysis
was limited to control and high-dose groups, a study reporting feature that limited the ability to
identify dose-related trends. In limited cases, EPA performed additional statistical analysis to
identify trends or refine analyses consistent with EPA guidance (e.g., analyzing developmental data
sets on a per litter basis rather than by individual fetus). Study results have been extracted and
presented in evidence tables.
Notable features of the RDX database
Three 2-year toxicity bioassays of RDX are available as unpublished laboratory studies (Lish
etal.. 1984: Levine etal.. 1983b: Hart. 19761. The bioassays by Levine etal. (1983b) in the rat and
by Lish etal. (1984) in the mouse were conducted in accordance with Food and Drug
Administration (FDA) Good Laboratory Practices (GLPs) in place at the time of the studies. Both
studies included interim sacrifices (at 6 and 12 months). Complete histopathological examinations
were performed on all animals in the control and high-dose groups; however, only a subset of
tissues was examined in the mid-dose groups, limiting the ability to identify dose-related trends for
tissues with incomplete histopathology. Additionally, in the mouse bioassay by Lish etal. Q9841.
the initial high dose (175 mg/kg-day) was reduced to 100 mg/kg-day at week 11 because of high
mortality, thereby reducing the number of high-dose animals on study for the full 2 years of dosing
(see Table LS-5).
An earlier unpublished 2-year study in rats by Hart T19761 used a dose range that was
lower than the Levine etal. Q983bl and Lish etal. T19841 bioassays. Histopathology findings were
limited by the lack of pathology examinations in the mid-dose groups and lack of individual time of
death, which impacts the ability to interpret the histopathology data. In addition, a heating system
malfunction on days 75-76 of the study resulted in the death of 59 rats from the control and
treatment groups, thereby reducing the number of animals in the study (see Table LS-5).
Experimental animal toxicity studies of RDX involving less-than-lifetime exposure (Grouse
etal.. 2006: Angerhofer etal.. 1986: MacPhail etal.. 1985: Levine etal.. 1981a: Cholakis etal.. 1980:
Hart. 1974: Martin and Hart. 1974: von Oettingen etal.. 1949) were published or reported between
the years 1949 and 2006, and differences in robustness of study design, conduct, and reporting
reflect that time span. All but two of the eight short-term and subchronic toxicity studies of RDX
are available as unpublished laboratory studies; published studies include von Oettingen et al.
(1949) and Levine etal. (1981a). a laboratory report of a 13-week study of RDX in F344 rats with
subsets of the data subsequently published as Levine etal. f!981b) and Levine etal. f 19901 The
majority of studies conducted histopathological examinations on only some of the experimental
groups (e.g., control and high dose). One subchronic study Grouse etal. (2006) was peer-reviewed
by Versar, Inc. for EPA in 2012. The peer reviewers determined that the report provided useful
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1 information on the toxicity of RDX, including an array of endpoints for neurotoxicity and
2 immunotoxicity (U.S. EPA. 2012dl The assessment of neurotoxicity in the study could have been
3 improved with more histological evaluation as well as additional behavioral assessment
4 Some of the more important limitations in study design, conduct, and reporting of
5 experimental animal toxicity studies of RDX are summarized in Table LS-5. Limitations of these
6 studies as well as the study evaluation consideration described in this section were taken into
7 consideration in evaluating and synthesizing the evidence for each of the health effects in
8 Section 1.2.
9 Table LS-5. Experimental animal studies considered less informative because
10 of certain study design, conduct, or reporting limitations
References
Study design, conduct, and reporting limitations
Lish etal. (1984)
2-yr mouse study
The initial high dose (175 mg/kg-d) was reduced to 100 mg/kg-d at wk 11
due to high mortality. Mortality of surviving mice was similar to controls
after dose reduction.
Hart (1976)
2-yr rat study
A heating system malfunction on d 75-76 of the study resulted in the
deaths of 59 rats from the control and treatment groups. Dead animals
were subsequently eliminated from the analysis. Interpretation of the
histopathology findings was limited by the lack of pathology examinations
in the mid-dose groups and lack of individual time of death. Test material
poorly characterized; purity was not reported.
Cholakis et al. (1980)
13-wk mouse study (Experiment 1)
The dose range was too low to produce effects in mice. Histopathological
examinations were not performed.
Cholakis et al. (1980)
13-wk mouse study (Experiment 2)
Nonstandard dosing regimen followed: 0, 40, 60, or 80 mg/kg-d for 2 wks.
For the next 11 wks, the dosing was inverted, so that the 40 mg/kg-d
group received 320 mg/kg-d, the 60 mg/kg-d group received 160 mg/kg-d,
and the 80 mg/kg-d group continued to receive the same dose. The
rationale for this dosing regimen was not provided in the study report.
Levine et al. (1981a)
13-wk rat study
Analysis of one lot of rodent feed showed measurable levels of
contaminants, including chlorinated pesticides (dieldrin, heptachlor
epoxide, beta-hexachlorocyclohexane [BHC], and
dichlorodiphenyltrichloroethane [DDT]), polychlorinated biphenyls (PCBs),
and organophosphates (methyl parathion, carbophenothion, and
disulfeton).
Martin and Hart (1974)
13-wk monkey study
The species of monkey is unclear (either Cynomolgus or Rhesus). Some
test subjects may have had variable dosing due to emesis. Small sample
size per dose group (n = 3). Test material poorly characterized; purity was
not reported.
von Oettingen et al. (1949)
12-wk rat study
The strain of rat was not reported. Only gross observations were made at
autopsy.
von Oettingen et al. (1949)
6-wk dog study
The breed of dog was not reported. Only gross observations were made
at autopsy.
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Toxicological Review of Hexahydro-l,3,5-trinitro-l,3,5-triazine
1.HAZARD IDENTIFICATION
1.1. Overview of Chemical Properties and Toxicokinetics
1.1.1. Chemical Properties
1 Hexahydro-l,3,5-trinitro-l,3,5-triazine (RDX) is a member of the nitramine class of organic
2 nitrate explosives fBoileau etal.. 2003: Bingham etal.. 20011 and is not found naturally in the
3 environment It has low solubility in water fYalkowskv and He. 20031 and slowly volatilizes from
4 water or moist soil fATSDR. 20121. The normalized soil organic carbon/water partition coefficient
5 (Koc) values for RDX indicate a potential for RDX to be mobile in soil (Spanggord et al.. 1980al. The
6 vapor pressure suggests that RDX will exist as particulate matter in air and be removed by both wet
7 and dry deposition (Spanggord et al.. 1980al. RDX degrades in the environment and can be subject
8 to both photolysis fSikkaetal.. 1980: Spanggord etal.. 1980al and biodegradation fFunk etal..
9 1993: McCormicketal.. 19811 fTable 1-11.
10 Table 1-1. Chemical identity and physicochemical properties of RDX
Characteristic or property
Value
Reference
Chemical structure
+-°~
N
rS
N+ N
II 1
0 Cr
NLM (2011)
CASRN
121-82-4
Synonyms
Hexahydro-l,3,5-trinitro-s-triazine;
l,3,5-trinitro-l,3,5-triazacyclohexane;
1,3,5-trinitrohexahydro-s-triazine; cyclonite;
cyclotrimethylenetrinitramine; hexogen;
cyclotrimethylenenitramine; Research
Department Explosive; Royal Demolition
explosive; RDX
Color/form
White, crystalline solid
Bingham et al. (2001)
Molecular formula
C3H6N6O6
NLM (2011)
Molecular weight
222.12
Lide (2005)
Density (g/cm3 at 20°C)
1.82
Lide (2005)
Boiling point (°C)
276-280
Bingham et al. (2001)
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Characteristic or property
Value
Reference
Melting point (°C)
205.5
Lide (2005)
Heat of formation (kJ/g)
-0.277
Rvon et al. (1984)
Log Kow
0.87-0.90
Hansch et al. (1995)
Koc
42-167
Soanggord et al. (1980b)
Henry's law constant
(atm-m3/mole at 25°C)
2.0 x 10"11
ATSDR (2012)
Vapor pressure (mm Hg at 20°C)
4.10 x 10"9
Spanggord et al. (1980a)
Solubility in water (mg/L at
25°C)
59.7
Yalkowskv and He (2003)
1.1.2. Toxicokinetics
RDX is absorbed following exposure by oral and inhalation routes (see Appendix C,
Section C.l.l). Studies in experimental animals indicate that oral absorption rates can range from
approximately 50 to 90% (Krishnan et al.. 2009: Guo etal.. 1985: Schneider et al.. 1978.19771. with
the rate and extent of absorption dependent on the physical form of RDX (i.e., the increased surface
area associated with finely powdered RDX allows for increased absorption) and the dosing
preparation or matrix (Bannon etal.. 2009a: Krishnan et al.. 2 0 0 9: Grouse etal.. 2008: Bannon.
2006: Guo etal.. 1985: MacPhail etal.. 1985: Schneider et al.. 19771. Dermal absorption of RDX has
been demonstrated in in vitro studies using human and pig skin fReddv etal.. 2008: Reifenrath et
al.. 20081.
RDX is systemically distributed, including to the brain (i.e., RDX can cross the blood:brain
barrier), heart, kidney, liver, and fat (Musick etal.. 2010: Bannon etal.. 2006: MacPhail etal.. 1985:
Schneider et al.. 19771. In rats, RDX can be transferred from dam to fetus across the placentahblood
barrier, and has been identified in maternal milk fHess-Ruth et al.. 20071.
The metabolism of RDX in humans has not been investigated. Studies in experimental
animals indicate that metabolism of RDX is extensive and includes denitration, ring cleavage, and
generation of C02 possibly through cytochrome P450 (CYP450) (Musick et al.. 2010: Major etal..
2007: Fellows etal.. 2006: Bhushan etal.. 2003: Schneider et al.. 1978.19771.
RDX and metabolites are eliminated primarily via urinary excretion and exhalation of C02
fSweenev etal.. 2012a: Musick et al.. 2010: Krishnan etal.. 2009: Maior etal.. 2007: Schneider etal..
19771. Estimated elimination half-lives (ti/2) indicate that RDX is more rapidly metabolized in mice
than in rats and humans; estimated ti/2 values were 1.2 hours for mice, 5-10 hours for rats, and
15-29 hours for humans fSweenev etal.. 2012b: Krishnan et al.. 2009: Ozhan etal.. 2003: Woodv et
al.. 1986: Schneider etal.. 19771.
A more detailed summary of RDX toxicokinetics is provided in Appendix C, Section C.l.
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1.1.3. Description of Toxicokinetic Models
A physiologically based pharmacokinetic (PBPK) model to simulate the pharmacokinetics of
RDX in rats was first developed by Krishnan et al. f20091 and revised to extend the model to
humans and mice (Sweeney etal.. 2012a: Sweeney et al.. 2012b). The Sweeney etal. (2012a) model
consists of six main compartments: blood, brain, fat, liver, and lumped compartments for rapidly
perfused tissues and slowly perfused tissues, and can simulate RDX exposures via the intravenous
(i.v.) or oral route. This model assumes that the distribution of RDX to tissues is flow-limited, and
represents oral absorption as first-order uptake from the gastrointestinal (GI) tract into the liver,
with 100% of the dose absorbed. RDX is assumed to be cleared by first-order metabolism in the
liver. The model does not represent the kinetics of any RDX metabolites. The Sweeney et al.
(2012a) and Sweeney etal. (2 012b) PBPK models were evaluated and subsequently modified by
the U.S. Environmental Protection Agency (EPA) for use in dose-response modeling in this
assessment (see Appendix C, Section C.1.5).
1.2. PRESENTATION AND SYNTHESIS OF EVIDENCE BY ORGAN/SYSTEM
In experimental animal studies, RDX test material administered in toxicology studies
included formulations that ranged in purity (from 84 to 99.99%) and in particle size (from <66 to
~200 [im particle size). Differences in test material purity and particle size were taken into
consideration while evaluating RDX toxicity findings; this is discussed in the literature search
section and incorporated in the synthesis of evidence.
Mortality has been reported in the animal toxicology studies conducted for RDX. Due to the
serious nature associated with a frank effect such as mortality, EPA specifically evaluated the
database with respect to mortality (see Appendix C, Section C.3.1). In brief, mortality was observed
following exposure to a range of doses in chronic-duration studies, in studies up to 6 months in
duration, and during gestation. In further analyzing the available evidence, mortality occurred at
lower doses in rats compared with mice and following gavage administration compared with
dietary administration. Additionally, mortality occurred to a greater extent with administration of
RDX in the form of relatively finer particle sizes, potentially due to the reduced ability of larger
particles of RDX to enter the bloodstream. Some investigators attributed the mortality to RDX-
related cancer or noncancer effects (e.g., kidney or nervous system effects); others identified no
cause for the animal deaths. Typically, evidence related to various hazards is presented and
synthesized in distinct organ- or system-specific sections. However, in this case, the assessment
does not present mortality in a hazard section by itself due to the likelihood that events leading to
mortality fall under other specific hazards. Mortality evidence is considered in discussions of the
evidence for organ/system-specific hazards where applicable.
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1.2.1. Nervous System Effects
In humans, nervous system effects following RDX exposure have been observed in multiple
case reports, and the association between RDX exposure and neurobehavioral effects has been
examined in a single cross-sectional occupational epidemiology study. Information relevant to an
examination of the association between RDX exposure and nervous system effects also comes from
experimental animal studies involving chronic, subchronic, and gestational exposure to ingested
RDX. A summary of nervous system effects associated with RDX exposure is presented in
Tables 1-2 and 1-3 and Figure 1-1. Experimental animal studies are ordered in the evidence table
and exposure-response array by duration of exposure and then species.
Observational Studies in Humans
In a cross-sectional study by Ma and Li T1993I neurobehavioral effects were evaluated in
Chinese workers occupationally exposed to RDX. Memory retention and block design scores5 were
significantly lower among exposed workers (mean concentrations of RDX in two exposed groups:
0.407 and 0.672 mg/m3) compared to unexposed workers from the same plant However, no
significant differences were observed between the groups on other neurobehavioral tests (e.g.,
simple and choice reaction times, and letter cancellation test) (Table 1-2). This study did not
consider potential confounders such as alcohol consumption or co-exposure to trinitrotoluene
(TNT), and there was limited information characterizing exposure to RDX.
Case reports suggest an association between RDX exposure (via ingestion, inhalation, and
possibly dermal exposure) and neurological effects (see Appendix C, Section C.2). Severe
neurological disturbances include tonic-clonic seizures (formerly known as grand mal seizures) in
factory workers fTestud etal.. 1996a: Testud etal.. 1996b: Kaplan etal.. 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 etal.. 1969: Merrill. 19681.
seizures, dizziness, headache, and nausea following nonwartime/nonoccupational exposures
(Kasuske etal.. 2009: Davies etal.. 2007: Kiiciikardali et al.. 2003: Hettand Fichtner. 2002: Harrell-
B ruder and Hutchins. 1995: Goldberg etal.. 1992). and seizures in a child following ingestion of
plasticized RDX from the mother's clothing fWoodv et al.. 1986).
Studies in Experimental Animals
Nervous system effects in experimental animals include a wide array of behavioral changes
consistent with the induction of seizures by RDX exposure, and have been observed in the majority
of chronic, subchronic, and developmental studies examining oral exposure to RDX (see Table 1-3
5The memory quotient index measured short-term hearing memory, visual memory, combined hearing and
visual memory, and learning ability. The block design index measured visual perception and design
replication, and the ability to analyze spatial relationships.
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and Figure 1-1). Although study authors interchangeably used the terms seizures and convulsions,
seizures, which result from abnormal electrical activity in the brain, can outwardly manifest in a
variety of ways. While seizures can be detected in the form of convulsions, they can also manifest
as facial twitches, tremors, or increased irritability, or they may go unnoticed. While behavioral
methods exist to capture a spectrum of responses known to occur as a result of this aberrant
neuronal activity, the most reliable detection methods are electrophysiological fRacine. 19721.
Convulsions have been reported in studies with different animal species and experimental
designs. In every study that reported convulsions, the incidence of convulsions increased with
dose. In 2-year dietary studies in rats (F344 and Sprague-Dawley) and mice (B6C3Fi), convulsions
were observed beginning at doses of 35-40 mg/kg-day, but not at lower doses fLish etal.. 1984:
Levine etal.. 1983b: Hart. 19761.6 Subchronic dietary exposure to RDX was also associated with
convulsions in the rat, although doses reported to increase convulsive activity were inconsistent
across studies. Convulsions were reported in RDX-exposed rats at subchronic doses as low as 8 and
25 mg/kg-day (Crouse etal.. 2006: von Oettingen etal.. 19491. In three other rat studies involving
exposure durations of 30-90 days, no evidence of seizures, convulsions, or tremors was reported at
doses ranging from 1 to 50 mg/kg-day (MacPhail etal.. 1985: Cholakis etal.. 19801 (both
unpublished technical reports). Levine etal. f 19901 reported convulsions in rats following
subchronic exposure only at a dose of 600 mg/kg-day (a dose associated with 100% mortality);
however, the unpublished technical report of this study f Levine etal.. 1981 al inconsistently
reported convulsions at 600 and >30 mg/kg-day, thereby reducing confidence in the identification
of the dose level at which nervous system effects were observed in this study. RDX exposure (by
gavage) during gestation in the rat was associated with induction of seizures or convulsions in the
dams at doses ranging from 2 to 40 mg/kg-day (Angerhofer etal.. 1986: Cholakis etal.. 19801
(unpublished technical reports), demonstrating that effects on the nervous system can be observed
following exposure durations as short as 10-14 days. Convulsions were also reported in dogs
exposed to 50 mg/kg-day RDX for 6 weels fvon Oettingen et al.. 19491. but not 10 mg/kg-day for
the 13 weeks (Hart. 19741 (unpublished technical report), and in five of six monkeys following a
gavage dose of 10 mg/kg-day (Martin and Hart. 19741 (unpublished technical report).
1 In the only study addressing susceptibility to seizures, Burdette etal. (19881 found that
2 seizure occurrence was more frequent in Long Evans rats exposed to a single dose of 50 or
3 60 mg/kg RDX by gavage when challenged with an audiogenic stimulus 8 and 16 hours after
4 treatment. However, no audiogenic seizures were observed at the earlier 2- and 4-hour post-
"The 2-year dietary studies in F344 rats by Levine et al. f1983bl and B6C3Fi mice by Lish et al. f 19841 were
available only as a laboratory reports. An external peer review was conducted by EPA in July 2012 to
evaluate the accuracy of experimental procedures, results, and interpretation and discussion of the findings
presented. A report of this peer review performed by Versar, Inc. is available through the EPA's IRIS Hotline
at (202) 566-1676 (phone) or hotline.iris@epa.gov (e-mail address), and on the Health and Environmental
Research Online (HERO) database (U.S. EPA. 2012dl. The 2-year dietary study in Sprague-Dawley rats by
Hart T19761 is available as an unpublished technical report.
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dosing test periods even though RDX plasma concentrations were elevated throughout the testing
period. In a complementary experiment, Long Evans rats treated daily with 6 mg/kg-day RDX for
up to 18 days required fewer stimulation trials to exhibit amygdaloid kindled seizures compared to
controls. Neither the purity nor the specific particle size of the RDX used in the experiments by
Burdette etal. f 19881 was reported.
The majority of animal studies reported convulsions and/or seizures as clinical
observations; interpretation of these observations is limited to some extent because the nature and
severity of convulsions and seizures were not more fully characterized. The 90-day study by
Grouse etal. (2006)7 was one of the few studies that collected and reported incidence data for
convulsions and tremors, and demonstrated a clear dose-related increase in convulsions and
tremors in male and female F344 rats associated with RDX exposure via gavage (see Table 1-3).
Tremors were reported following administration of >12 mg/kg-day, persisting throughout the
90-day study. Convulsions were observed at >8 mg/kg-day in male and female rats; information on
duration and onset was not reported (Grouse etal.. 2006).
In general, gavage dosing induced convulsions at lower doses than did dietary
administration. For example, in the gavage studies by Grouse etal. (2006) and Cholakis etal.
Q9801. convulsions were observed in 1-3 rats/group at doses of 2-8 mg/kg-day; at doses of
15-20 mg/kg-day, convulsions were observed in approximately 60-70% of the animals. In
contrast, in a 2-year dietary study by Levine etal. f!983bl. convulsions were reported only at a
dose of 40 mg/kg-day; no convulsions were observed at lower doses (<8 mg/kg-day). The
difference in response between gavage and dietary administration may be due to the bolus dosing
resulting from gavage administration and the comparatively faster absorption and higher peak
blood concentrations of RDX.
Several experimental animal studies documented that unscheduled deaths were frequently
preceded by convulsions or seizures. In a 2-year study in rats, Levine etal. (1983b) noted that
tremors and/or convulsions were often seen in high-dose animals prior to their death. In a rat
developmental toxicity study fCholakis etal.. 1980). investigators concluded that early deaths in
dams were preceded by convulsions based on the observation of convulsions in one rat prior to
death, and a similar appearance (e.g., dried blood around the mouth and nose) in other dams that
died during the study. Convulsions preceding death were also observed in pregnant Sprague-
Dawley rats exposed to RDX during gestation fAngerhofer et al.. 19861.
7The 13-week gavage study in F344 rats by Grouse etal. (2006) was available only as a laboratory report. An
external peer review was conducted by Versar, Inc. in July 2012 to evaluate the accuracy of experimental
procedures, results, and interpretation and discussion of the findings presented. The U.S. EPA (2012d) report
of this peer review is available through the EPA's IRIS Hotline at (202) 566-1676 (phone) or
hotline.iris@epa.gov (e-mail address), and on the HERO database.
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The 90-day Grouse etal. (2006) study provides the most detailed information on the
relationship between convulsions and mortality (see Appendix C, Table C-10 for additional
information on evidence of mortality associated with RDX exposure). Convulsions (3/20) and pre-
term deaths (2/20)8 were observed in male and female rats exposed to 8 mg/kg-day RDX; the
incidences of both convulsions and pre-term deaths were higher in dose groups with greater
exposures. Investigators stated that nearly all observed pre-term deaths in rats exposed to the
three higher doses (10,12, and 15 mg/kg-day RDX) for 90 days were preceded by neurotoxic signs
such as rearing behavior, tremors and convulsions; however, pre-term death did not occur in all
animals that convulsed. Some uncertainty exists in that convulsions were not typically observed
during a functional observational battery (FOB) test conducted after exposure, likely due to the
time needed to complete exposures prior to beginning behavioral testing. Of the 100 RDX-treated
rats in the Grouse etal. (2006) study, convulsions were documented in 34 male and female rats
across the five dose groups (with convulsions initially observed anywhere from day 7 to 87); based
on additional information provided as a memorandum by study investigators flohnson. 2015a). 26
of these 34 rats (76%) survived to the end of the 90-day study. In general, higher doses of RDX
were associated with fewer days of exposure before the first convulsion was observed. Of the eight
rats that exhibited convulsions prior to pre-term death, convulsions were documented anywhere
from the same day that the animal died to 8 weeks prior to death. Of the 26 rats that seized and
survived to day 90, the first seizures were observed as early as day 10 and as late as day 87. Thus,
while an increase in mortality was observed in the Grouse etal. (2006) study at the same dose as
convulsions, the additional information provided by lohnson (2015a) do not show as clear a
correspondence between convulsions (and other neurotoxic signs) and mortality. Analysis of these
data is limited to the extent that convulsions may have occurred at times when animals were not
observed and are therefore undercounted in the individual animal data; however, lohnson (2015a)
noted that it is unlikely that seizure observations were missed, since seizures generally occurred
soon after dosing.
A few studies reported mortality that was not specifically or directly associated with
neurological effects (see Appendix C, Table C-10) (Angerhofer etal.. 1986: Levine etal.. 1981a: von
Oettingen etal.. 1949): however, in these studies, animals may not have been monitored for clinical
observations with sufficient frequency to have observed convulsive activity prior to death. In case
reports of convulsions and other nervous system effects in workers exposed to RDX during
manufacture and in individuals exposed acutely as a result of accidental or intentional ingestion,
there were no reports of mortality subsequent to the convulsions (see Appendix C, Section C.2).
Additional neurobehavioral effects associated with RDX exposure in rats included increased
hyperactivity, hyper-reactivity to approach, fighting, and irritability at doses similar to those that
8At the 8 mg/kg-day dose level, the three rats that convulsed survived to the end of the study; no convulsions
were observed in the two rats that died before study termination.
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induced tremors, convulsions, and seizures (20-100 mg/kg-day) (Levine etal.. 1990: Angerhofer et
al.. 1986: Levine etal.. 1983b: Levine etal.. 1981a. b; von Oettingen et al.. 19491. Hyperactivity and
nervousness were also reported in male mice that received a subchronic exposure to 320 mg/kg-
day RDX fCholakis etal.. 19801. No changes in motor activity, flavor aversion, scheduled-controlled
behavior, or acoustic startle response were observed in a 30-day gavage study in rats (changes in
acoustic startle response in acute exposures at higher doses [12.5-50 mg/kg] were noted), but
doses were relatively low (<10 mg/kg-day) (MacPhail etal.. 19851. and no significant changes in
behavioral or neuromuscular activity were observed in rats following exposure to <15 mg/kg-day
for 90 days (Grouse etal.. 20061. Grouse etal. (20061 observed that stained haircoats and increased
barbering in female F344 rats receiving 15 mg/kg-day may have been caused by the oral dosing
procedure (gavage) alone.
Changes in absolute and relative brain weight were mixed across studies. Elevated absolute
brain weights were reported in subchronic assays in B6C3Fi mice and F344 rats (Grouse etal..
2006: Levine etal.. 1990: Levine etal.. 1981a. b; Cholakis etal.. 19801: however, the changes were
not consistently observed across studies. Relative brain weights in some studies showed
correspondingly greater increases compared to absolute brain weight (Grouse etal.. 2006: Levine
etal.. 1983b: Cholakis etal.. 19801. but these changes were likely a result of changes in body weight
in the study, and were not a useful measure of effects of RDX on brain weights. In 2-year oral
studies, a decrease in absolute brain weight of female B6C3Fi mice (3-4% relative to control) was
reported at doses >35 mg/kg-day (Lish etal.. 19841. where as an increase in absolute brain weight
(2% relative to control) was observed in F344 rats at a dose of 40 mg/kg-day (Levine etal.. 1983b).
Less weight is placed on evidence of organ weight changes from chronic (2-year) studies because
normal physiological changes associated with aging and intercurrent disease may contribute to
inter-animal variability that could confound organ weight interpretation f Sellers etal.. 20071.
In some studies, seizures appeared soon after dosing, suggesting that seizure induction was
more strongly correlated with dose level than with duration of exposure. Consistent with this
observation are the findings of Williams etal. (20111. who demonstrated that RDX is rapidly
absorbed and crosses the blood:brain barrier following oral administration in rats, and that
distribution of RDX (8 [ig/g wet weight) to the brain correlated with seizure onset
While a dose-response relationship was observed consistently within studies, a dose that
induced convulsions in animals in one study did not necessarily induce convulsions at the same
dose in another study. This lack of consistency may be attributed, at least in part, to differences in
the purity or particle size of the test material across studies. Assuming that increased particle size
(and the corresponding reduction in available surface area compared with smaller particle sizes)
results in slowed absorption and distribution to the brain, studies that used a larger particle size
may be expected to produce less neurotoxicity in test animals. The mouse study by Cholakis et al.
f 19801 used a relatively large RDX particle size (200 [im) compared to the rat study by Levine et al.
f 1983b 1 that used a smaller (<66 [im) particle size. This could contribute to why the Cholakis et al.
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1 (1980) subchronic dietary study in the mouse (doses up to 320 mg/kg-day RDX) and rat (doses up
2 to 40 mg/kg-day) failed to report seizures or convulsions. Finally, differences in study design may
3 have contributed to differences in reported neurological responses in subchronic and chronic
4 duration studies. In particular, the protocols for observation for clinical signs (e.g., observations
5 performed once daily in the morning in Levine etal. f l983bll may not have been sufficiently
6 frequent to accurately measure the incidence of seizures or other nervous system effects.
7 Table 1-2. Evidence pertaining to nervous system effects in humans
Reference and study design
Results
Ma and Li (1993) (China)
Cross-sectional study, 60 workers from
Neurobehavioral function tests, scaled scores (mean, standard
deviation)
the same plant exposed to RDX (30 in
Group A [26 males; 4 females]; 30 in
Group B [24 males; 6 females]),
compared to 32 workers with similar
age, education level, and length of
employment from same plant with no
exposure to RDX (27 males; 5 females).
Exposure measures: Details of
exposure measurement were not
provided; two groups of workers
exposed to the following mean RDX
concentrations in air (basis for dividing
workers into two exposure groups was
Test
Control
Group A
Group B
Memory retention*
Simple reaction time
(milliseconds)
Choice reaction time
(milliseconds)
Block design*
(elapsed time)
Letter cancellation
(quality per unit time)
111.3 (9.3)
493 (199)
763 (180)
18.0 (5.4)
1,487 (343)
96.9 (9.6)
539 (183)
775 (161)
16.0 (4.3)
1,449 (331)
91.1 (10.3)
578 (280)
770 (193)
13.5(6.7)
1,484 (443)
not provided).
Concentration (mg/m3) (mean
± standard deviation):
Group A 0.407 (± 0.332)
Group B 0.672 (± 0.556)
*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)
Effect measures3: Five neurobehavioral
function tests and five additional
memory subtests.
Analysis: Variance (F-test); unadjusted
linear regression, multiple regression,
and correlation analysis.
Subtest
Control
Group A
Group B
Directional memory*
Associative learning*
Image free recall*
23.5 (3.6)
24.9 (5.1)
24.1 (3.8)
17.2 (4.9)
20.0 (4.3)
20.9 (4.1)
18.1 (5.7)
18.5 (4.6)
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).
8
9 aSymptom data were not included in evidence table because of incomplete reporting.
10
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Table 1-3. Evidence pertaining to nervous system effects in animals
Reference and study design
Results
Convulsions and neurobehavioral effects
Lish etal. (1984)
Mice, B6C3Fi, 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 urn
0,1.5, 7.0, 35, or 175/100 mg/kg-d (high
dose reduced to 100 mg/kg-d in wk 11
due to excessive mortality)
Diet
2 yrs
One male in the 35 mg/kg-d dose group and one female in the
175/100 mg/kg-d group convulsed near the end of the study.
Hart (1976)
Rats, Sprague-Dawley, 100/sex/group
Purity and particle size not specified
0,1.0, 3.1, or 10 mg/kg-d
Diet
2 yrs
No nervous system effects, as evidenced by clinical signs or changes
in appearance or behavior, were reported.
Levine et al. (1983b)
Rats, F344, 75/sex/group; interim
sacrifices (10/sex/group) at 6 and 12 mo
89.2-98.7% pure, with 3-10% HMX as
contaminant; 83-89% of particles <66 nm
0, 0.3,1.5, 8.0, or 40 mg/kg-d
Diet
2 yrs
Tremors, convulsions, and hyper-responsiveness to stimuli were
noted in males and females at 40 mg/kg-d; no incidence data were
reported.
Cholakis et al. (1980)
Mice, B6C3Fi, 10-12/sex/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 nm particle size
0, 40, 60, or 80 mg/kg-d for 2 wks
followed by 0, 320,160, or 80 mg/kg-d
(TWA doses of 0, 79.6, 147.8, or
256.7 mg/kg-d for males and 0, 82.4,
136.3, or 276.4 mg/kg-d for females)3
Diet
13 wks
Hyperactivity and/or nervousness observed in 50% of the high-dose
males; no signs observed in females'5; no incidence data were
reported.
Cholakis et al. (1980)
Rats, F344,10/sex/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 nm particle size
0,10,14, 20, 28, or 40 mg/kg-d
Diet
13 wks
No nervous system effects, as evidenced by clinical signs or changes
in appearance or behavior, were reported.
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Reference and study design
Results
Cholakis et al. (1980)
Rats, CD, two-generation study;
FO: 22/sex/group; Fl: 26/sex/group;
F2: 10/sex/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 urn particle size
FO and Fl parental animals: 0, 5,16, or
50 mg/kg-d
Diet
FO exposure: 13 wks pre-mating, and
during mating, gestation, and lactation of
Fl; Fl exposure: 13 wks after weaning,
and during mating, gestation, and
lactation of F2; F2 exposure: until weaning
No nervous system effects were reported.
Crouse et al. (2006)
Rats, F344,10/sex/group
99.99% pure
0, 4, 8,10,12, or 15 mg/kg-d
Gavage
13 wks
Doses
0 4 8b 10 12 15
Convulsions (incidence)
M
F
0/10 0/10 1/10 3/10 8/10 7/10
0/10 0/10 2/10 3/10 5/10 5/10
Tremors (incidence)
M
F
0/10 0/10 0/10 0/10 2/10 3/10
0/10 0/10 0/10 0/10 0/10 1/10
(Levine et al. (1990): Levine et al. (1981a),
1981b))c
Rats, F344,10/sex/group; 30/sex for
control
84.7 ± 4.7% purity, ~10% HMX, median
particle diameter 20 nm, ~90% of particles
<66 nm
0,10, 30,100, 300, or 600 mg/kg-d
Diet
13 wks
Hyper-reactivity to approach was observed in rats (sex not specified)
receiving >100 mg/kg-d; no incidence data were reported.
Tremors and convulsions were observed prior to death in one female
and two male rats receiving 600 mg/kg-d.d (600 mg/kg-d was lethal
to all rats.)
von Oettingen et al. (1949)
Rats, sex/strain not specified, 20/group
90-97% pure, with 3-10% HMX; particle
size not specified
0,15, 25, or 50 mg/kg-d
Diet
13 wks
Hyperirritability and convulsions were observed in the 25 and
50 mg/kg-d groups'5; no incidence data were reported.
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Reference and study design
Results
Hart (1974)
Dogs, Beagle, 3/sex/group
Pre-mix with ground dog chow containing
20 mg RDX/g-chow, 60 g dog food; purity
and particle size not specified
0, 0.1,1, or 10 mg/kg-d
Diet
13 wks
No nervous system effects, as evidenced by clinical signs or changes
in appearance or behavior, were reported.
Martin and Hart (1974)
Monkeys, Cynomolgus or Rhesus6,
3/sex/group
Purity of test material not specified
0, 0.1,1, or 10 mg/kg-d
Gavage
13 wks
Doses
_Q
o
1
1
1
o
o
CNS effects characterized as depression, trembling, shaking, jerking,
or convulsions (incidence)
M
F
0/3 0/3 0/3 2/3
0/3 0/3 0/3 3/3
von Oettingen et al. (1949)
Dogs, breed not specified,
5 females/group (control);
7 females/group (exposed)
90-97% pure, with 3-10% HMX; particle
size not specified
0 or 50 mg/kg-d
Diet
6 d/wk for 6 wks
Treated dogs exhibited convulsions, excitability, ataxia, and
hyperactive reflexesb; no incidence data were reported.
MacPhail et al. (1985)
Rats, Sprague-Dawley derived CD,
8-10 males or females/group
Purity 84 ± 4.7%; <66 nm particle size
0,1, 3, or 10 mg/kg-d
Gavage
30 d
No changes in motor activity, flavor aversion, scheduled-controlled
response, or acoustic startle-response were reported.
Cholakis et al. (1980)
Rats, F344, 24-25 females/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants
0, 0.2, 2.0, or 20 mg/kg-d
Gavage
GDs 6-19
Doses
0 0.2 2.0 20
Convulsions (incidence)
F
0/24 0/24 1/24 18/25
Angerhofer et al. (1986) (range-finding
study)
Rats, Sprague-Dawley, 6 pregnant
females/group
Purity 90%; 10% HMX and 0.3% acetic acid
occurred as contaminants
0,10, 20, 40, 80, or 120 mg/kg-d
Gavage
GDs 6-15
Convulsions preceding death were observed at >40 mg/kg-d;
no incidence data were reported.
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Reference and study design
Results
Angerhofer et al. (1986)
Rats, Sprague-Dawley, 39-51 mated
females/group
Purity 90%; 10% HMX and 0.3% acetic acid
occurred as contaminants
0, 2, 6, or 20 mg/kg-d
Gavage
GDs 6-15
Convulsions and hyperactivity15 were observed at 20 mg/kg-d;
no incidence data were reported.
Burdette et al. (1988)
Rats, Long Evans, 10-21 males/group
0,10,12.5, 20, 25, 50, or 60 mg/kg-d
Study conducted as two experiments with
the same study design, each with a
control group
Gavage
8 hrs (single exposure)
After an 8-hr exposure, rats placed in
observation chamber; 0-64 kHz, 95 dB
ultrasonic cleaner turned on for 1 min or
until seizure initiated with uncontrolled
running (whichever occurred first)
Doses
0 10 12.5 20 25 50 60
Prevalence of audiogenic seizures (%f
M
0 9 0 29 40 82* 78*
fValues estimated from graph using Grab It! Software. Statistical
significance indicated by study authors; p < 0.017.
Brain weight
Lish et al. (1984)
Mice, B6C3Fi, 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 nm
0,1.5, 7.0, 35, or 175/100 mg/kg-d (high
dose reduced to 100 mg/kg-d in wk 11
due to excessive mortality)
Diet
2 yrs
Doses
0 1.5 7 35 175/100
Absolute brain weight (percent change compared to control)
M
F
0% -0.2% 0.61% 0.81% -1%
0% -2% -2% -4%* -3%*
Relative brain weight (percent change compared to control)
M
F
0% 4% 2% 2% 5%
0% -4% -1% -3% 18%*
Levine et al. (1983b)
Rats, F344, 75/sex/group; interim
sacrifices (10/sex/group) at 6 and 12 mo
89.2-98.7% pure, with 3-10% HMX as
contaminant; 83-89% of particles <66 nm
0, 0.3,1.5, 8.0, or 40 mg/kg-d
Diet
2 yrs
Doses
o
00
LO
m
o
o
Absolute brain weight (percent change compared to control)
M
F
0% 2% -1% 2% 2%
0% -0.3% -0.4% 1% 2%*
Relative brain weight (percent change compared to control)
M
F
0% 0% 8% 2% 22%*
0% -1% 3% 4% 20%*
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Reference and study design
Results
Cholakis et al. (1980)
Mice, B6C3Fi, 10-12/sex/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 urn particle size
Experiment 1: 0,10,14, 20, 28, or
40 mg/kg-d
Diet
13 wks
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)3
Diet
13 wks
Doses
0 10 14 20 28 40
Absolute brain weight (percent change compared to control)
M
F
0% - - - 2% 2%
0% - 4% 2%
Relative brain weight (percent change compared to control)
M
F
0% - 6% 2%
0% - - - 0% 3%
Doses
0 80 160 320
Absolute brain weight (percent change compared to control)
M
F
0% 0% 2% 10%
0% 0% 4% 2%
Relative brain weight (percent change compared to control)
M
F
0% -3% 1% 8%
0% 0% 3% -4%
Cholakis et al. (1980)
Rats, F344,10/sex/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 nm particle size
0,10,14, 20, 28, or 40 mg/kg-d
Diet
13 wks
Doses
0 10 14 20 28 40
Absolute brain weight (percent change compared to control)
M
F
0% - - - 3% 0%
0% - - - 0% 0%
Relative brain weight (percent change compared to control)
M
F
0% - 7%* 10%*
0% - - 5% 6%
Crouse et al. (2006)
Rats, F344,10/sex/group
99.99% pure
0, 4, 8,10,12, or 15 mg/kg-d
Gavage
13 wks
Doses
0 4 8 10 12 15
Absolute brain weight (percent change compared to control)
M
F
0% -1% -0.3% 2% 5%* 7%*
0% -2% 6% 1% 4% 6%
Relative brain weight (percent change compared to control)
M
F
0% 6% 10% 5% 3% 4%
0% -2% -2% -12%* -12%* -15%*
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Reference and study design
Results
(Levine et al. (1990); Levine et al. (1981a),
1981b))c
Rats, F344,10/sex/group; 30/sex for
control
84.7 ± 4.7% purity, ~10% HMX, median
particle diameter 20 nm, ~90% of particles
<66 nm
0,10, 30,100, 300, or 600 mg/kg-d
Diet
13 wks
Doses
0 10 30 100 300 600
Absolute brain weight (percent change compared to control)
M
F
0% 1% 0.53% -6%
0% -1% 1% 2%
Relative brain weight (percent change compared to control)
M
F
0% 4% 7% 14%
0% 0.3% 2% 5%
^Statistically significant (p < 0.05) based on analysis by study authors.
aDoses were calculated by the study authors.
bMortality was reported in some RDX-treated groups in this study.
cLevine et al. (1981a) is a laboratory report of a 13-week study of RDX in F344 rats; two subsequently published
papers (Levine et al., 1990; Levine et al., 1981b) present subsets of the data provided in the full laboratory report.
discrepancies in the doses at which convulsions occurred were identified in the technical report. The nervous
system effects reported in this table and in the corresponding exposure-response array are those provided in the
results section of the technical report (Levine et al., 1981a) and in the published paper (Levine et al., 1990). In
other sections of the technical report, the authors reported that hyperactivity to approach and convulsions were
observed in rats receiving >30 mg/kg-day (abstract and executive summary), or that mortality was observed in
rats receiving 100 mg/kg-day and that hyperactivity to approach, tremors, and convulsions were observed in
animals exposed to lethal doses (discussion).
eThe species of monkey used in this study was inconsistently reported in the study as either Cynomolgus (in the
methods section) or Rhesus (in the summary).
CNS = central nervous system; GD = gestational day; HMX = octahydro-l,3,5,7-tetranitro-l,3,5,7-tetrazocine;
TWA = time-weighted average
Note: A dash ("-") indicates that the study authors did not measure or report a value for that dose group.
This document is a draft for review purposes only and does not constitute Agency policy.
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1000
100
-3 10
s
• significantly changed
o not significantly changed
6 o
O 6
O 6
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Mechanistic Evidence
Studies that have explored the mode of action (MOA) of RDX on the central nervous system
(CNS) have focused on the potential impacts on neurotransmission. These studies implicate a MOA
for RDX-induced seizures and convulsions involving distribution to the brain (across the
blood:brain barrier) and subsequent effects on neurotransmitters, including gamma-amino butyric
acid (GABA) and glutamate. There is significant evidence from the scientific literature to suggest
that RDX neurotoxicity results from interactions of RDX with the GABAa receptor. GABA is a major
inhibitory neurotransmitter in the brain, and the GABAa receptor has been implicated in
susceptibility to seizures fGalanopoulou. 20081.
In research conducted by the U.S. Army Center for Health Promotion and Preventative
Medicine, Williams etal. f20111 and Bannonetal. f2009al showed a correlation between blood and
brain concentrations of RDX in rats that received a single oral dose of RDX (>98-99.5% purity) by
gavage, which closely correlated with the time of seizure onset. RDX (75 mg/kg) was distributed to
the brain in direct proportion to levels found in the blood, while time to seizure onset was reduced
as RDX brain levels increased fWilliams etal.. 20111. Similarly, oral exposure to RDX (via a gel
capsule: 3 or 18 mg/kg) resulted in quick absorption followed by transport to the brain and
subsequent alterations in neurotransmission fBannon et al.. 2009al.
In receptor binding studies, RDX has only showed affinity for GABAa receptors fWilliams et
al.. 2011: Williams and Bannon. 20091. Specifically, RDX showed a significant affinity for the
picrotoxin convulsant site of the GABA channel. RDX treatment in brain slices from the basolateral
amygdala inhibited GABAA-mediated inhibitory postsynaptic currents and initiated seizure-like
neuronal discharges. RDX exposure may reduce the inhibitory effects of GABAergic neurons,
resulting in enhanced excitability that could lead to seizures fWilliams etal.. 2011: Williams and
Bannon. 20091. 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 (leffervs etal.. 2012: Gilbert. 19941. Burdette etal. (19881 hypothesized that the limbic
system was involved in seizures caused by RDX exposure, given that amygdaloid kindled rats (rats
subjected to patterns of electrical stimulation to promote the development of seizures) exhibited
pro-convulsant activity at a dose that was approximately half of the dose necessary for RDX to
induce spontaneous seizures (rats treated with RDX also required fewer electrical stimulations to
trigger kindled seizures). Potential limbic system involvement is also suggested given its role in
integrating emotional and behavioral responses (including aggression) and the anecdotal
observations of hyperactivity, hyper-responsiveness to approach, and irritability noted across
several studies of RDX toxicity (Levine etal.. 1990: Levine etal.. 1983b: Levine etal.. 1981a. b;
Cholakis etal.. 1980: von Oettingen et al.. 19491.
RDX binding at the picrotoxin convulsant site of the GABA channel may also inform the
relationship between exposure to the chemical and the time when a seizure was observed. In
general across the RDX database, induction of convulsions and seizures appears to be more strongly
This document is a draft for review purposes only and does not constitute Agency policy.
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correlated with dose than duration of exposure. However, Gerkin etal. (2010) demonstrated that
young C57/B16 mice injected intraperitoneally (i.p.) with picrotoxin to induce seizures had a
significantly increased frequency of elevated neuronal activity ("Up state"), and firing rates were
significantly increased in neocortical neurons up to 24 hours after exposure. It is possible that this
period of elevated neuronal activity could increase the likelihood that a subsequent stimulus could
trigger a seizure. While the study authors did not look at longer durations post exposure, it is
possible in a chronic exposure scenario that repeated exposure to RDX binding at the same site as
picrotoxin, through a general increase in brain tissue with elevated neuronal activity, could
increase the likelihood of seizures developing over time, or have other longer-term effects on
normal brain function.
It is possible to construct a hypothetical MOA for RDX-induced neurotoxicity based on the
evidence summarized above. Following absorption and transport to the brain:
1) Parent RDX acts as a receptor antagonist (supported by Schneider etal. (1977) and
Williams etal. (2011)). binding noncompetitively to the picrotoxin convulsant site of the
GABAa receptor (supported by Williams and Bannon (2009) and Williams etal. (2011)).
2) RDX binding to the GABAa receptor results in decreased conduction of chloride through
the ion channel.
3) Reduced chloride conduction results in depolarization of the neuronal membrane,
thereby reducing spontaneous inhibitory postsynaptic currents (sIPSCs). Williams et al.
(2011) observed a reduction in the amplitude and frequency of sIPSCs in whole-cell in
vitro recordings of neurons in brain slices from the ratbasolateral amygdala after
exposure to RDX. In addition, RDX treatment of slices inhibited GABA-induced currents.
4) Reduction in sIPSCs results in an overall reduction in inhibitory inputs to the nervous
system. Williams etal. (2011) observed a pattern of seizure-like neuronal discharges in
vitro from slices of the basolateral amydgala in rats after adding RDX and noted that the
effects were not reversible after 40 minutes of washout.
The steps above provide a biologically plausible sequence of mechanistic events that result
in the generation of seizure-like neuronal activity. Reduction of the inhibitory GABAergic signaling
is common to many convulsants, as summarized in Kalueff f20071. Some organochlorine
insecticides, including alpha-endosulfan, dieldrin, and lindane, also exert neurotoxic effects through
interaction with the GABAa receptor, and can produce a range of hyperexcitability effects (including
convulsions) in mammals (Vale etal.. 2003: Bloomquist. 1992: Sunol etal.. 1989). The interaction
of RDX with the GABAa receptor is directly supported by receptor-binding assays fWilliams etal..
2011). Although these binding assays were performed on rat receptors, it is plausible that the
results are relevant to human neurotoxicity. Seizures have been observed in many species,
including humans, rats, mice, dogs, lizards, and birds at varying dosages and durations of exposure
This document is a draft for review purposes only and does not constitute Agency policy.
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fOuinn etal.. 2013: McFarland etal.. 2009: Johnson et al.. 2007: Bruchim etal.. 2005: Kiiciikardali et
al.. 2003: Woody etal.. 1986: Lish etal.. 1984: Berry etal.. 1983: Levine etal.. 1983b). A more
recent meta-analysis of toxicogenomic data across a phylogenetically diverse set of organisms (rat,
quail, fathead minnow, earthworm, and coral) demonstrated that neurotoxic responses are
conserved in more highly-related species and that binding to the GABAa receptor is a common
molecular initiating event f Garcia-Revero etal.. 20111. While these lines of evidence do not
preclude a role of other receptors as yet unscreened for RDX binding affinity, they support
involvement of the GABAergic pathway described above in the development of RDX neurotoxicity.
The GABAa receptor is also a target of many anticonvulsant therapies (e.g., benzodiazepines,
propofol, barbiturates) fMeldrum and Rogawski. 2007: Mohler. 20061. Additional support for the
involvement of GABAergic signaling in the neurotoxicity of RDX comes from human case reports. In
multiple case reports, medical intervention included treatment with benzodiazepines (commonly
diazepam or lorazepam) to treat seizing patients (Kasuske etal.. 2009: Davies etal.. 2007:
Kiiciikardali etal.. 2003: Hettand Fichtner. 2002: Woody etal.. 19861. Benzodiazepines act in large
part by enhancing the effects of GABA at the GABAa receptor by increasing chloride conductance,
resulting in anticonvulsant and relaxant effects (Goodman et al.. 19961.
Some other pro-convulsant agents with minimal direct toxicity to nerve cells, such as sarin
and some organophosphate pesticides, are known to act through inhibition of acetylcholinesterase
(AChE) activity fMcDonough and Shih. 19971. Some of the clinical signs observed following RDX
exposure are similar to the clinical signs associated with organophosphate pesticides and nerve
agents (Grouse etal.. 2006: Burdette etal.. 1988: Barsotti and Crotti. 19491. However, the limited
data available for RDX do not support AChE inhibition as a primary mechanism because: (1) blood
and brain levels of AChE are unaffected by RDX (Williams et al.. 2011: Williams and Bannon. 20091:
and (2) in vitro neurotransmitter receptor binding studies do not reveal any affinity of RDX for
acetylcholine receptors fWilliams etal.. 2011: Williams and Bannon. 20091. Additionally, common
AChE-induced symptoms (salivation and lacrimation) have not routinely been observed fWilliams
etal.. 20111. RDX showed no affinity for other receptors that are known targets of convulsants,
including the glutamate family of receptors, nicotinic receptors, glycine receptors, and several
monoamine receptors (Williams etal.. 2011: Williams and Bannon. 20091.
In a microarray experiment, Bannon et al. (2009a) found that RDX caused a down
regulation of an abundance of genes in the cerebral cortex related to neurotransmission, including
those encoding proteins involved in synaptic transmission and vesicle transport Genes encoding
proteins involved in the glutamate pathway were also underexpressed, indicating a possible
mechanism of action for RDX via excessive glutamate stimulation. The authors speculated that this
depression of the major excitatory neurotransmitter system could be a negative response to the
increase in seizure likelihood from RDX influx into the brain. Molecular changes in response to RDX
have been described by Zhang and Pan f2009bl. who observed significant changes in micro-RNA
(miRNA) expression in the brains of B6C3Fi mice fed 5 mg RDX/kg diet festimated dose: 0.75-1.5
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mg/kg-dav: Bannon etal.. 2009b) for 28 days. One miRNA, miR-206, was upregulated 26-fold in
RDX-exposed brains; brain-derived neurotrophic factor (BDNF) was identified as a downstream
gene target of this miRNA, along with two other miRNAs that were upregulated in RDX-exposed
brains (miR-30a and miR-195) f Zhang and Pan. 2009a. b). BDNF is a member of the neurotrophin
family of growth factors, and promotes the survival and differentiation of existing and new neurons.
Deng etal. f 20141 conducted miRNA and mRNA profiling in rats to identify targets up or
downregulated after 48-hour exposure to RDX, finding that many of the gene targets of these
miRNAs were associated with nervous system function, and may contribute to the neurotoxicity of
RDX. However, while effects of RDX on BDNF expression or other downstream targets may play a
role in RDX neurotoxicity, the utility of miRNAs as predictors of toxicity has not been demonstrated
and downstream targets of miRNA require verification fBannon etal.. 2009b). Thus, the
contribution, if any, of aberrant expression of a suite of miRNAs to the MOA for RDX neurotoxicity is
unknown.
Recent research has provided greater insight to inform a mechanistic basis of RDX
neurotoxicity. While other possible MOA(s) may contribute to the overall neurotoxicity of RDX, the
demonstrated affinity of RDX for the GABAa receptor, evidence of supportive electrophysiological
changes with direct application of RDX, and toxicokinetic evidence of distribution of RDX to the
brain provides a mechanistic basis for the association of seizures with exposure to RDX. The
available information supports that RDX-induced hyperactivity and seizures likely result from
inhibition of GABAergic signaling in the limbic system.
Integration of Nervous System Effects
Evidence for nervous system effects associated with exposure to RDX comes from studies in
both humans and animals. One occupational study reported memory impairment and decrements
in certain neurobehavioral tests in workers exposed to RDX compared to controls fMa and Li.
19931. and human case reports provide other evidence of an association between acute RDX
exposure and neurological effects. There was consistent evidence of neurotoxicity associated with
exposure to RDX; 11 of 16 repeat-dose animal studies (of varying design) reported neurological
effects (some severe), including seizures, convulsions, tremors, hyperirritability, hyper-reactivity,
and behavioral changes, associated with RDX exposure (Grouse etal.. 2006: Angerhofer etal.. 1986:
Levine etal.. 1983b: Levine etal.. 1981b: Cholakis etal.. 1980: von Oettingen et al.. 1949). In most
of these studies, the occurrence of neurological effects was dose-related. In those studies that
found no evidence of RDX-associated neurotoxicity fMacPhail etal.. 1985: Cholakis etal.. 1980:
Hart. 1976.1974). differences in dosing, particle size, and purity of the RDX administered could
possibly account for the lack of effect. Seizures resulting from RDX exposure likely result from
inhibition of GABAergic signaling due to the interaction of RDX with the GABAa receptor.
Convulsant receptor binding leading to a decreased seizure threshold, considered with kindling
studies, suggests that the effect is specific to CNS toxicity.
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Together, toxicological information in animals and humans, supported by toxicokinetic and
mechanistic information, provides a coherent identification of nervous system effects as a human
hazard of RDX exposure.
1.2.2. Kidney and Other Urogenital System Effects
The association between RDX exposure and effects on clinical measures of kidney function
was examined in one occupational epidemiology study. Case reports, involving accidental exposure
to ingested or inhaled RDX, offer some information on the potential for acute exposure to RDX to
affect the kidney in humans. Organ weight and histopathology findings from experimental animal
studies involving subchronic and chronic exposure to ingested RDX also provide data relevant to an
examination of the association between RDX exposure and kidney and other urogenital system
effects. A summary of these effects associated with RDX exposure is presented in Tables 1-4 to 1-8
and Figure 1-2. Experimental animal studies are ordered in the evidence table and exposure-
response array by duration of exposure and then by species.
Human case reports of individuals accidently exposed to unknown amounts of RDX by
ingestion or inhalation provide some evidence that RDX may affect the kidney and the urogenital
system. Reported symptoms included decreased urine output fKetel and Hughes. 1972: Knepshield
and Stone. 1972: Hollander and Colbach. 1969: Merrill. 19681. blood in urine fKasuske etal.. 2009:
Knepshield and Stone. 1972: Hollander and Colbach. 1969: Merrill. 19681. proteinuria fKasuske et
al.. 2009: Kiiciikardali etal.. 2003: Ketel and Hughes. 1972: Hollander and Colbach. 1969: Merrill.
19681. glucosuria fKiiciikardali et al.. 20031. elevated blood urea nitrogen (BUN) levels (Hollander
and Colbach. 1969: Merrill. 19681. and one case of acute renal failure requiring hemodialysis
following accidental inhalation of RDX fKetel and Hughes. 19721. In many of these case reports,
renal parameters returned to normal within a few days following exposure. No changes in renal
parameters were reported in other individuals exposed to unknown amounts of RDX fStone etal..
1969: Kaplan etal.. 19651. In a cross-sectional epidemiologic study of workers from five U.S. Army
munitions plants (69 exposed to RDX alone and 24 exposed to RDX and octahydro-
l,3,5,7-tetranitro-l,3,5,7-tetrazocine (HMX); RDX exposure range: undetectable [<0.01 mg/m3] to
1.6 mg/m3), no statistically significant differences in BUN or total serum protein between
nonexposed and RDX-exposed groups were observed fHathawav and Buck. 19771 (Table 1-4). As it
is a cross-sectional study, no information was provided on the length of employment or other
proxies that could be used to indicate cumulative exposure concentrations.
Studies in experimental animals provide some evidence that RDX exposure is associated
with kidney and other urogenital system effects (Table 1-5 and Figure 1-2). Histopathological
changes in the urogenital system were associated with exposure to RDX in a 2-year bioassay.
Specifically, increased incidences of kidney medullary papillary necrosis and pyelitis, uremic
mineralization, bladder distention and/or cystitis, and suppurative prostatitis were observed in
high-dose (40 mg/kg-day) male rats that died spontaneously or were sacrificed in moribund
condition fLevine etal.. 1983bl. These renal effects were considered the principal cause of
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treatment-related morbidity and mortality in these high-dose males. Similar kidney lesions were
not observed in female rats in this study. An increased incidence of tubular nephrosis was
observed in male B6C3Fi mice exposed to 320 mg/kg-day RDX in feed for 90 days, but not in female
mice in this study fCholakis etal.. 19801. In other chronic and subchronic oral studies in rats and
mice, no histopathological changes in the kidney were associated with RDX exposure fCrouse etal..
2006: Levine etal.. 1990: Lish etal.. 1984: Levine etal.. 1981a. b; Cholakis etal.. 1980: Hart. 19761.
Increased incidence of minimal to mild mineralization of the medulla was observed in male and
female monkeys exposed to 10 mg/kg-day RDX for 90 days by gavage f Martin and Hart. 19741. but
the study authors did not identify this as treatment related. No dose-related histopathological
changes were reported in a subchronic study in dogs fHart. 19741. and no histological alterations
were noted in the kidneys of rabbits exposed dermally to a cumulative dose of 165 mg/kg RDX in
dimethylsulfoxide (DMSO) received over a 4-week period (5 days/week) fMcNamara etal.. 19741.
Measurement of serum chemistry parameters that may indicate effects on renal function, including
BUN and uric acid, in studies of RDX in mice, rats, dogs, and monkeys fCrouse etal.. 2008: Levine et
al.. 1990: Lish etal.. 1984: Levine etal.. 1981a. b; Cholakis etal.. 1980: Hart. 1976.1974: Martin and
Hart. 19741 revealed variations (increases or decreases) from the respective control groups that
were not dose-related.
The findings of suppurative prostatitis provide the strongest evidence of urogenital toxicity.
A significant, dose-related increase in the total incidence of suppurative prostatitis was reported in
male F344 rats exposed to >1.5 mg/kg-day RDX in the diet for 2 years (Levine etal.. 1983b).
Suppurative prostatitis was not observed in 90-day studies in the rat involving oral (dietary or
gavage) exposure to RDX (Grouse etal.. 2006: Levine etal.. 1990: Levine etal.. 1981a. b). Similarly,
prostate effects were not observed in a 2-year dietary study in mice (Lish etal.. 19841. The Levine
etal. Q983bl report is the only 2-year study that reported examination of the prostate in rats.
Some reports have hypothesized that the observation of prostate inflammation in Levine et al.
f 1983b 1 is secondary to a bacterial infection unrelated to RDX toxicity fATSDR. 2012: Sweeney et
al.. 2012a: Grouse etal.. 20061. For example, in describing the results from the 2-year dietary study
in rats, Grouse etal. (20061 observed that the inflammation reflects a common condition in rodents,
noting that since 85% of the incidence occurred in rats found at spontaneous death or moribund
sacrifice (SDMS), it was most likely that the condition was a result of an incidental bacterial
infection. However, Levine etal. Q983bl distinguished between nonsuppurative and suppurative
inflammation (the latter being characterized by the formation of pus and a high concentration of
neutrophils). Although the proportion of suppurative prostatitis was higher in SDMS rats, there
was an increasing trend with dose in both the scheduled sacrifice (SS) and SDMS groups; the
incidence of suppurative prostatitis in the control group was 4% when the SS and SDMS groups
were combined. Additionally, the dose-related nature of the increased incidence suggests that the
primary cause (potentially leading to bacterial infection) was treatment-related, as a more uniform
distribution of rats with suppurative prostatitis would be expected with a spontaneous or age-
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related lesion. The dose-responsiveness could be explained if the infections were secondary to
treatment-related immunotoxicity, but there is no evidence from Levine etal. f!983bl to support
this possibility. A more thorough analysis of immune endpoints in a 90-day gavage exposure of
F344 rats did not identify any immunotoxic effects associated with RDX (Grouse etal.. 20061.
As noted above, Levine etal. Q983bl documented an array of kidney and other urogenital
lesions in their 2-year dietary exposure of F344 rats to RDX. However, the sequence by which those
effects may have occurred is unclear. Renal medullary necrosis, bladder distension, and cystitis
were observed mainly in the male rats exposed to 40 mg/kg-day RDX for 24 months, although one
rat in the 0.3 mg/kg-day dose group also exhibited these lesions. Treatment-related effects on the
kidney (necrosis) and bladder (distension/obstruction and hemorrhagic cystitis) were also
identified in the 12-month pathology report (see Tables 1-6 to 1-8). The absence of these
observations in the 6-month interim pathology report suggests that an exposure duration
>6 months may be required before RDX-induced effects on the urogenital system are observed.
Suppurative prostatitis was observed with increasing incidence in each dose group in the study at
24 months. Considered as a group, treatment-related kidney and urogenital lesions may have led to
a blockage that resulted in urinary stasis. Reduced urinary flow and/or retrograde flow may have
contributed to an environment that allowed bacterial infection of the prostate. Thus, while an
opportunistic bacterial infection could be the proximal cause of the suppurative prostatitis (ATSDR.
2012: Sweeney et al.. 2012a: Grouse etal.. 20061. it may have been secondary to the effects of RDX
on the urogenital system. This hypothesis is consistent with the observed dose-related increase in
incidence of suppurative prostatitis.
Although the ultimate sequence of effects in the urogenital system is unclear, even from
review of the scheduled sacrifices at 6 or 12 months on study, it is plausible that suppurative
prostatitis would occur after other kidney or bladder lesions that resulted in the initial blockage
and urinary stasis. The incidence of suppurative prostatitis reported in Levine etal. (1983b) was
increased at doses lower than the doses associated with an increased incidence of other urogenital
lesions. However, the incidence of bladder lesions may have been underreported, as the bladders
were only examined following observation of a gross abnormality. Bladder distension was
reported sporadically among the lower dose groups (0.3,1.5, or 8.0 mg/kg-day), but the bladder
was not routinely examined in these groups (Levine etal.. 1983bl. Although the pathogenesis of
kidney and urogenital system effects cannot be established, the available evidence is consistent
with suppurative prostatitis as an indirect effect of RDX exposure and as a marker for the broader
array of kidney and urogenital system effects observed by Levine etal. Q983bl.
Changes in kidney weights in subchronic oral toxicity studies in rats, dogs, and monkeys did
not show a clear pattern of increase or decrease associated with RDX exposure. Kidney weight
changes were either not dose-related or were inconsistent across sexes when absolute and relative
weights were compared (see Table 1-5). Less weight is placed on evidence of organ weight changes
from chronic (2-year) studies fLish etal.. 1984: Hart. 19761 because normal physiological changes
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associated with aging and intercurrent disease may contribute to inter-animal variability that could
confound organ weight interpretation (Sellers etal.. 20071.
Exposure to HMX, the major contaminant in many of the available RDX studies, was
associated with histopathological changes in the kidney and alterations in renal function in female,
but not male, rats fed doses >450 mg/kg-day HMX for 13 weeks (see the Integrated Risk
Information System [IRIS] assessment of octahydro-1,3,5,7-tetranitro-1,3,5,7-tetrazocine [HMX] at
http://www.epa.gov/irisl. No effects were observed at doses <115 mg/kg-day. Because the
percentage of HMX as an impurity ranged from 3 to 10%, resulting in HMX exposures
<60 mg/kg-day in the studies of RDX toxicity, the contribution of HMX to the observed kidney
toxicity in studies of RDX is expected to be negligible. Further, differences in the pattern of toxicity
(i.e., kidney effects observed only in RDX-exposed males and HMX-exposed females) also suggest
that HMX contaminants were not responsible for kidney effects in rats exposed to RDX.
Table 1-4. Evidence pertaining to kidney effects in humans
Reference and study design
Results
Hathawav and Buck (1977)
Renal function tests: mean (standard deviation not reported)
Cross-sectional study, 2,022 workers,
RDX exposed males*
1,491 participated (74% response rate).
Referent
Undetected (0.01 mg/m3
Analysis group: limited to whites;
Test
(n = 237)
(n = 22) (n = 45)
69 workers exposed to RDX alone and
BUN
15.5
15.6 16.4
24 workers exposed to RDX and HMX,
Total protein
7.2
7.2 7.3
compared to 338 workers not exposed to
RDX exposed females*
RDX, HMX, or TNT.
Exposure measures: Exposure
Referent
(n = 101)
Undetected (0.01 mg/m3
(n = 1) (n = 25)
determination based on job title and
industrial hygiene evaluation; exposed
BUN
Total protein
13.2
7.3
8 12.6
7.6 7.2
subjects assigned to two groups:
undetected (0.01 mg/m3
(mean for employees with exposures
includes both workers exposed to RDX alone and RDX and HMX.
No differences were statistically significant in men or women.
>LOD: 0.28 mg/m3).
Effect measures: Renal function tests
(blood)
Analysis: Types of statistical tests were
not reported (assumed to be t-tests for
comparison of means and x2 tests for
comparison of proportions).
LOD = limit of detection
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Table 1-5. Evidence pertaining to kidney and other urogenital system effects
in animals
Reference and study design
Results
Histopathological lesions
Lish etal. (1984)
Mice, B6C3Fi, 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 urn
0,1.5, 7.0, 35, or 175/100 mg/kg-d (high
dose reduced to 100 mg/kg-d in wk 11
due to excessive mortality)
Diet
2 yrs
The incidence of cytoplasmic vacuolization of renal tubules was
greater for RDX-treated males than the control group males after
6 mo of treatment. However, at 12 and 24 mo of treatment, this
lesion was observed as frequently in controls as males treated with
RDX. There was no increase in incidence of this lesion in females at
any time point.
Hart (1976)
Rats, Sprague-Dawley, 100/sex/group
Purity and particle size not specified
0,1.0, 3.1, or 10 mg/kg-d
Diet
2 yrs
Histopathological examination of kidney did not reveal any significant
differences compared to controls; lesions observed were not
attributed to RDX treatment; incidence data were reported only for
control and 10 mg/kg-d groups.
Levine et al. (1983b)
Rats, F344, 75/sex/group; interim
sacrifices (10/sex/group) at 6 and 12 mo
89.2-98.7% pure, with 3-10% HMX as
contaminant; 83-89% of particles <66 nm
0, 0.3,1.5, 8.0, or 40 mg/kg-d
Diet
2 yrs
Note: More detailed histopathological
results, including interim sacrifice data at
6 and 12 mo, are provided in Tables 1-6 to
1-8.
Data for male rats sacrificed on schedule (SS) and those that died
spontaneously or were sacrificed moribund (SDMS) (summarized
below) were analyzed separately; incidence data were not reported
for females.
Doses
0
0.3
1.5
8.0
40
Kidney, medullary papillary necrosis; 24 mo (incidence)
(SS)
0/38
0/36
0/25
0/29
0/4
(SDMS)
0/17
1/19
0/27
0/26
18/27*
(Sum)
0/55
1/55
0/52
0/55
18/31*
Kidney, suppurative pyelitis; 24 mo (incidence)
(SS)
0/38
0/36
0/25
0/29
0/4
(SDMS)
0/17
1/19
0/27
1/26
5/27*
(Sum)
0/55
1/55
0/52
1/55
5/31*
Kidney, uremic mineralization; 24 mo (incidence)
(SS)
1/38
0/36
0/25
0/29
0/4
(SDMS)
0/17
1/19
2/27
0/26
13/27
(Sum)
1/55
1/55
2/52
0/55
13/31
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Reference and study design
Results
Urinary bladder, luminal distention; 24 mo (incidence)
(SS)
(SDMS)
(Sum)
0/38
0/36
0/25
0/29
1/4*
0/16
2/19
1/27
3/22
24/28
0/54
2/55
1/52
3/51
25/32
Urinary bladder, cystitis hemorrhagic/suppurative; 24 mo
(incidence)
(SS)
(SDMS)
(Sum)
0/38
0/36
0/25
1/29
0/4
0/16
2/19
1/27
0/22
18/27
0/54
2/55
1/52
1/51
18/31
Prostate, suppurative inflammation (prostatitis); 24 mo (incidence)
SS 0/38 1/36 2/25* 4/29* 0/4
SDMS 2/16 3/19 7/27* 8/26 19/27*
(Sum) 2/54 4/55 9/52* 12/55* 19/31*
Cholakis et al. (1980)
Mice, B6C3Fi, 10-12/sex/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 nm 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)3
Diet
13 wks
Doses
0
80
160
320
Tubular nephrosis (incidence)
M
F
0/10
0/11
4/9*
1/11
Cholakis et al. (1980)
Rats, F344,10/sex/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 nm 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.
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Reference and study design
Results
Cholakis et al. (1980)
Rats, CD, two-generation study;
FO: 22/sex/group; Fl: 26/sex/group;
F2: 10/sex/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 urn particle size
FO and Fl parental animals: 0, 5,16, or
50 mg/kg-d
Diet
FO exposure: 13 wks pre-mating, and
during mating, gestation, and lactation of
Fl; Fl exposure: 13 wks after weaning,
and during mating, gestation, and
lactation of F2; F2 exposure: until weaning
Data were reported only for F2 generation controls and 5 and
16 mg/kg-d groups.
Doses
0 5 16 50
Cortical cysts (incidence)
M
F
4/10 4/10 8/10
3/10 4/10 8/10
Crouse et al. (2006)
Rats, F344,10/sex/group
99.99% pure
0, 4, 8,10,12, or 15 mg/kg-d
Gavage
13 wks
Doses
0 4 8 10 12 15
Prostate, mild subacute inflammation (incidence)
M
0/10 - - 1/8
Histopathological examination of kidney did not reveal any significant
differences compared to controls; incidence data were reported only
for control and 15 mg/kg-d groups.
(Levine et al. (1990): Levine et al. (1981a),
1981b))b
Rats, F344,10/sex/group; 30/sex for
control
84.7 ± 4.7% purity, ~10% HMX, median
particle diameter 20 nm, ~90% of particles
<66 nm
0,10, 30,100, 300, or 600 mg/kg-d
Diet
13 wks
Histopathological examination of kidney did not reveal any significant
differences compared to controls. No histopathology findings
available for the 300 or 600 mg/kg-d dose groups because all rats in
these groups died before the 13-wk necropsy.
Hart (1974)
Dogs, Beagle, 3/sex/group
Pre-mix with ground dog chow containing
20 mg RDX/g-chow, 60 g dog food; purity
and particle size not specified
0, 0.1,1, or 10 mg/kg-d
Diet
13 wks
Histopathological examination of kidney did not reveal any significant
differences compared to controls; incidences were reported only for
control and 10 mg/kg-d groups.
Martin and Hart (1974)
Monkeys, Cynomolgus or Rhesus0,
3/sex/group
Purity of test material not specified
0, 0.1,1, or 10 mg/kg-d
Gavage
13 wks
Doses
o
1
1
1
o
o
Medulla; mineralization, minimal to mild (incidence)
M + F
0/6 1/6 0/6 4/6
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Reference and study design
Results
Kidney weight?
Lish etal. (1984)
Mice, B6C3Fi, 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 urn
0,1.5, 7.0, 35, or 175/100 mg/kg-d (high
dose reduced to 100 mg/kg-d in wk 11
due to excessive mortality)
Diet
2 yrs
Doses
0 1.5 7.0 35 175/100
Absolute kidney weight at 104 wks (percent change compared to
control)
M
F
0% -1% 4% 9%* 19%*
0% 3% 1% 1% -2%
Relative kidney weight at 104 wks (percent change compared to
control)
M
F
0% 3% 6% 11%* 27%*
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
o
1
1
rn
o
O
Absolute kidney weight (percent change compared to control)
M
F
0% -3% -7% 2%
0% 14% -4% 8%
Relative kidney weight (percent change compared to control)
M
F
0% -1% -4% 4%
0% 22% 3% 18%
Levine et al. (1983b)
Rats, F344, 75/sex/group; interim
sacrifices (10/sex/group) at 6 and 12 mo
89.2-98.7% pure, with 3-10% HMX as
contaminant; 83-89% of particles <66 nm
0, 0.3,1.5, 8.0, or 40 mg/kg-d
Diet
2 yrs
Doses
o
o
00
LO
m
o
o
Absolute kidney weight at 105 wks (percent change compared to
control)
M
F
0% 2% -7% 1% 0%
0% 3% 3% 2% 2%
Relative kidney weight at 105 wks (percent change compared to
control)
M
F
0% 1% 0% 2% 20%*
0% 3% 6% 5% 21%*
Cholakis et al. (1980)
Mice, B6C3Fi, 10-12/sex/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 nm 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
F
0% - 18% 2%
0% - - - -8% -5%
Relative kidney weight (percent change compared to control)
M
F
0% - 29% 0%
0% - - - -8% -3%
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Reference and study design
Results
Experiment 2: 0, 40, 60, or 80 mg/kg-d for
2 wks followed by 0, 320,160, or
80 mg/kg-d (TWA doses of 0, 79.6,147.8,
or 256.7 mg/kg-d for males and 0, 82.4,
136.3, or 276.4 mg/kg-d for females)3
Diet
13 wks
Doses
0 80 160 320
Absolute kidney weight (percent change compared to control)
M
F
0% 8% 11% 13%
0% -5% -3% 0%
Relative kidney weight (percent change compared to control)
M
F
0% 5% 9% 10%
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 nm particle size
0,10,14, 20, 28, or 40 mg/kg-d
Diet
13 wks
Doses
0 10 14 20 28 40
Absolute kidney weight (percent change compared to control)
M
F
0% - - - -2% -5%
0% - 1% 0%
Relative kidney weight (percent change compared to control)
M
F
0% - - - 1% 5%
0% - 6% 6%
Cholakis et al. (1980)
Rats, CD, two-generation study;
F0: 22/sex/group; Fl: 26/sex/group;
F2: 10/sex/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 nm particle size
F0 and Fl parental animals: 0, 5,16, or
50 mg/kg-d
Diet
F0 exposure: 13 wks pre-mating, and
during mating, gestation, and lactation of
Fl; Fl exposure: 13 wks after weaning,
and during mating, gestation, and
lactation of F2; F2 exposure: until weaning
Doses
0 5 16 50
Absolute kidney weight (percent change compared to control)
M
F
0% 6% -12%
0% -4% -21%*
Crouse et al. (2006)
Rats, F344,10/sex/group
99.99% pure
0, 4, 8,10,12, or 15 mg/kg-d
Gavage
13 wks
Doses
0 4 8 10 12 15
Absolute kidney weight (percent change compared to control)
M
F
0% -3% -4% -1% 3% 5%
0% 2% 5% 13%* 10% 15%*
Relative kidney weight (percent change compared to control)
M
F
0% 3% 6% 2% 1% 3%
0% 1% -3% -1% -6% -7%*
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Reference and study design
Results
(Levine et al. (1990); Levine et al. (1981a),
1981b))b
Rats, F344,10/sex/group; 30/sex for
control
84.7 ± 4.7% purity, ~10% HMX, median
particle diameter 20 nm, ~90% of particles
<66 nm
0,10, 30,100, 300, or 600 mg/kg-d
Diet
13 wks
Doses
0 10 30 100 300 600
Absolute kidney weight (percent change compared to control)
M
F
0% 1% 1% -9%
0% 1% 3% -1%
Relative kidney weight (percent change compared to control)
M
F
0% 5% 7% 10%
0% 3% 5% 2%
Hart (1974)e
Dogs, Beagle, 3/sex/group
Pre-mix with ground dog chow containing
20 mg RDX/g-chow, 60 g dog food; purity
and particle size not specified
0, 0.1,1, or 10 mg/kg-d
Diet
13 wks
Doses 0 0.1 1 10
Absolute kidney weight (percent change compared to control)
M
F
0% - - 38%
0% - - -18%
Martin and Hart (1974)e
Monkeys, Cynomolgus or Rhesus6,
3/sex/group
Purity of test material not specified
0, 0.1, 1, or 10 mg/kg-d
Gavage
13 wks
Doses 0 0.1 1 10
Absolute kidney weight (percent change compared to control)
M + F
0% -2% -3% 4%
^Statistically significant (p < 0.05) based on analysis by study authors.
aDoses were calculated by the study authors.
bLevine et al. (1981a) is a laboratory report of a 13-week study of RDX in F344 rats; two subsequently published
papers (Levine et al., 1990; Levine et al., 1981b) present subsets of the data provided in the full laboratory report.
cThe species of monkey used in this study was inconsistently reported in the study as either Cynomolgus (in the
methods section) or Rhesus (in the summary).
dAn analysis by Craig et al. (2014) found a statistically significant correlation between absolute, but not relative,
kidney weights and renal histopathology. Therefore, only absolute kidney weight data from RDX studies are
presented in Figure 1-2.
eKidney weight data from the Hart (1974) and Martin and Hart (1974) studies were considered less informative
than other studies. Hart (1974) reported organ weight data for high-dose dogs (3/sex/group) only, and the kidney
weights from Martin and Hart (1974) were highly variable across monkeys (e.g., kidney weights for the control
animals ranged from 4.9 to 13.1 g). Therefore, kidney weight data from these two studies were not presented in
the exposure-response array for kidney and other urogenital system effects (Figure 1-2).
Note: A dash ("-") indicates that the study authors did not measure or report a value for that dose group.
SDMS - spontaneous death or moribund sacrifice; SS -- scheduled sacrifice
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1 Table 1-6. Six-, 12-, and 24-month incidence of kidney endpoints in male F344
2 rats reported for statistical evaluation in Levine et al. (1983b)
Doses (mg/kg-d)
0
0.3
1.5
8.0
40
Medullary papillary necrosis (incidence)
6 mo
SS
0/10
0/10
0/10
0/10
0/10
SDMS
-
-
-
-
0/5
Sum
0/10
0/10
0/10
0/10
0/15
12 mo
SS
0/10
0/10
0/10
0/10
0/10
SDMS
-
-
0/3
-
15/19*
Sum
0/10
0/10
0/13
0/10
15/29*
24 mo
SS
0/38
0/36
0/25
0/29
0/4
SDMS
0/17
1/19
0/27
0/26
18/27*
Sum
0/55
1/55
0/52
0/55
18/31*
Pyelitis (incidence)
6 mo
SS
0/10
0/10
0/10
0/10
0/10
SDMS
-
-
-
-
0/5
Sum
0/10
0/10
0/10
0/10
0/15
12 mo
SS
0/10
0/10
0/10
0/10
0/10
SDMS
-
-
0/3
-
1/19
Sum
0/10
0/10
0/13
0/10
1/29
24 mo
SS
0/38
0/36
0/25
0/29
0/4
SDMS
0/17
1/19
0/27
1/26
5/27*
Sum
0/55
1/55
0/52
1/55
5/31*
Pyelonephritis (incidence)
6 mo
SS
0/10
0/10
0/10
0/10
0/10
SDMS
-
-
-
-
0/5
Sum
0/10
0/10
0/10
0/10
0/15
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Doses (mg/kg-d)
0
0.3
1.5
8.0
40
12 mo
SS
0/10
0/10
0/10
0/10
0/10
SDMS
-
-
0/3
-
1/19
Sum
0/10
0/10
0/13
0/10
1/29
24 mo
SS
0/38
0/36
0/25
1/29
0/4
SDMS
0/17
0/19
2/27
1/26
1/27
Sum
0/55
0/55
2/52
2/55
1/31
1 ^Statistically significant (p < 0.05) based on analysis by study authors.
2
3 Note: A dash indicates that the study authors did not measure or report a value for that dose group.
4 SDMS - spontaneous death or moribund sacrifice; SS -- scheduled sacrifice
5
6 Source: Levine et al. (1983b).
7 Table 1-7. Six-, 12-, and 24-month incidence of urinary bladder endpoints in
8 male F344 rats reported for statistical evaluation in Levine et al. f1983b)
Doses (mg/kg-d)
0
0.3
1.5
8.0
40
Luminal distention (incidence)
6 mo
SS
0/10
0/10
0/10
0/10
0/10
SDMS
-
-
-
-
0/5
Sum
0/10
0/10
0/10
0/10
0/15
12 mo
SS
0/10
0/10
0/10
0/10
0/10
SDMS
-
-
0/3
-
18/19*
Sum
0/10
0/10
0/13
0/10
18/29
24 mo
SS
0/38
0/36
0/25
0/29
1/4*
SDMS
0/16
2/19
1/27
3/22
24/28*
Sum
0/54
2/55
1/52
3/51
25/32*
Cystitis, hemorrhagic/suppurative (incidence)
6 mo
SS
0/10
0/10
0/10
0/10
0/10
SDMS
-
-
-
-
0/5
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Doses (mg/kg-d)
0
0.3
1.5
8.0
40
Sum
0/10
0/10
0/10
0/10
0/15
12 mo
SS
0/10
0/10
0/10
0/10
0/10
SDMS
-
-
0/3
-
17/19*
Sum
0/10
0/10
0/13
0/10
17/29
24 mo
SS
0/38
0/36
0/25
1/29
0/4
SDMS
0/16
2/19
1/27
0/22
18/27*
Sum
0/54
2/55
1/52
1/51
18/31*
1
2 ^Statistically significant (p < 0.05) based on analysis by study authors.
3
4 Note: A dash indicates that the study authors did not measure or report a value for that dose group.
5 SDMS - spontaneous death or moribund sacrifice; SS -- scheduled sacrifice
6
7 Source: Levine et al. (1983b).
8 Table 1-8. Six-, 12-, and 24-month incidence of prostate endpoints in male
9 F344 rats reported for statistical evaluation in Levine et al. f1983b)
Doses (mg/kg-d)
0
0.3
1.5
8.0
40
Spermatic granuloma (incidence)
6 mo
SS
0/10
2/10
2/10
1/10
6/10*
SDMS
-
-
-
-
2/5
Sum
0/10
2/10
2/10
1/10
8/15*
12 mo
SS
0/10
0/10
1/10
1/10
0/10
SDMS
-
-
0/3
-
0/19
Sum
0/10
0/10
1/13
1/10
0/29
24 mo
SS
0/38
0/36
0/25
0/29
0/4
SDMS
0/16
0/19
0/27
0/26
0/27
Sum
0/54
0/55
0/52
0/55
0/31
Suppurative inflammation (incidence)
6 mo
SS
0/10
0/10
0/10
0/10
0/10
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Doses (mg/kg-d)
0
0.3
1.5
8.0
40
SDMS
-
-
-
-
0/5
Sum
0/10
0/10
0/10
0/10
0/15
12 mo
SS
0/10
0/10
0/10
0/10
0/10
SDMS
-
-
0/3
-
0/19
Sum
0/10
0/10
0/13
0/10
0/29
24 mo
SS
0/38
1/36
2/25*
4/29*
0/4
SDMS
2/16
3/19
7/27*
8/26
19/27*
Sum
2/54
4/55
9/52*
12/55*
19/31*
1
2 ^Statistically significant (p < 0.05) based on analysis by study authors.
3
4 Note: A dash indicates that the study authors did not measure or report a value for that dose group.
5 SDMS - spontaneous death or moribund sacrifice; SS -- scheduled sacrifice
6
7 Source: Levine et al. (1983b).
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1000
100
>
(0
T3
I
OD
JxL
OD
E
o
Q
10
0.1
• o
• o
o o °
o
o o
o
o
o
o
o
o
o •
o o
o
o
o
o
o • • • •
o o o o •
o o o o
• significantly changed
O not significantly changed
O O
o o o
«nT
00
cn
Chronic
o
o
00
cn
o
o
00
cn
o
.c
u
o
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cn
—
<£>
o
o
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cn
cn
o
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o
.c
u
Subchronic
«n|-
00
cn
Chronic
o
00
cn
o
.c
u
<£>
o
o
o
u
o
cn
cn
x
o2
Subchronic
Histopathological lesions
Absolute kidney weight
Note: Filled circle indicates that response was statistically significantly different from the control.
(1) Statistical significance determined from incidence at time of scheduled sacrfice. (2) Statistical significance determined from incidence at spontaneous death.
Figure 1-2. Exposure-response array of kidney and other urogenital system effects.
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Mechanistic Evidence
No MOA information is available for RDX-induced kidney and other urogenital effects.
However, mechanistic information underlying the neurotoxicity observed with RDX exposure, and
the specific affinity of RDX to the GABAa receptor-convulsant site fWilliams etal.. 2011: Williams
and Bannon. 20091. suggests a biologically plausible role for the GABAa receptor in RDX-related
effects on the urogenital system, and provides some potential MOA hypotheses for the effects
reported in Levine etal. (1983b).
One hypothesis is that urogenital effects of RDX are caused by interactions with GABAa
receptors mediating inputs to the urogenital system. GABA and GABA receptors have been
identified in a number of peripheral tissues fErdo etal.. 1991: Ong and Kerr. 1990: Erdo. 19851.
Brar etal. (20141 demonstrated that pretreatment with picrotoxin reduced the renoprotective
effects of sodium valproate (which acts on both GABAa and GABAb receptors) in a rat model of
ischemia-induced acute kidney injury, suggesting that GABAa receptors may be important in renal
function. GABA is believed to play a role in the regulation of urination and bladder capacity
(reviewed in Fowler etal. (20081 and Yoshimura and de Groat (199711. In rats, injection of a GABAa
receptor agonist inhibits the urination reflex flgawa etal.. 1993: Kontani et al.. 19871. GABAa
agonists injected into the periaqueductal gray area in rats inhibited reflex bladder activity, while
injection of an antagonist reduced bladder capacity and increased the frequency of bladder reflex
activity (Stone etal.. 20111. RDX would be expected to act like an antagonist and increase bladder
activity (which would not result in urinary stasis), although the impact of chronic exposure to RDX
acting as a GABAa receptor antagonist is not known. Evidence of GABAergic signaling regulating
bladder function, and the hypothesized disruption of that regulation by RDX via interaction with
GABAa receptors, may plausibly account for the kidney and other urogenital lesions, including
suppurative prostatitis, observed by Levine etal. Q983bl: however, no evidence to support this
hypothesized MOA is available.
Other potential mechanisms by which RDX, through GABAa binding, may lead to kidney and
urogenital effects are less apparent. Alterations in hormonal signaling or circulating levels of
estrogen or prolactin may lead to prostatitis. Prostate inflammation has been associated with
endocrine disruptors in the environment (Cowin etal.. 20101. and increased prolactin has been
shown to cause lateral lobe prostatitis f Stoker et al.. 1999b: Stoker etal.. 1999a: T angbanluekal and
Robinette. 1993: Robinette. 19881. Typically, the inflammation seen is chronic and does not reverse
over time (Robinette. 19881. Functional GABAa receptors have been identified in the anterior
pituitary (Zemkova et al.. 2008: Maverhofer etal.. 20011. which also serves as the primary source of
prolactin. Thus, the prostate inflammation observed in the rat in the 2-year study by Levine et al.
(1983b) could have been produced by disruption of pituitary prolactin or another hormonal signal
via interference with normal regulatory GABA-related hormonal control. However, no direct
evidence for this hypothesized MOA is available. Levine etal. Q983bl did not evaluate serum
endocrine measures or pituitary weights, and pituitary adenomas that could account for higher
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prolactin levels were not observed. A MOA hypothesis based on pituitary-mediated alterations in
endocrine signaling also does not explain the other urogenital lesions observed by Levine et al.
fl983bl
Another hypothesis is that the prostate effects could be mediated through an autoimmune
inflammatory response. GABAa receptor transcripts have been identified in immune cells of mouse
models fReyes-Garcia etal.. 2007: Tian etal.. 20041. and GABAa receptor agonists have decreased
cytotoxic immune responses and hypersensitivity reactions (Tian etal.. 1999: Bergeretetal.. 19981.
In a mouse autoimmune model of multiple sclerosis, Bhatetal. f 2 0101 found that treatment of
macrophages challenged with lipopolysaccharide with various GABA agonists decreased cytokine
production; addition of picrotoxin (which may have effects similar to those of RDX, as it binds to the
same site) was able to reduce this effect However, picrotoxin on its own did not significantly alter
cytokine production, suggesting that the effects are limited to reversal of agonist-induced
GABAergic activity (lohnson. 2015b). If an autoimmune mechanism was contributing to the effects
observed with RDX exposure, it is unclear why inflammation would be limited to the prostate. RDX
has also tested negative in the only battery of immunotoxicity tests to which it was subjected
(Grouse etal.. 20061.
If the urogenital effects are mediated through localized interaction with GABAa receptors,
another possibility would be that that effects would result from direct interactions with GABAa
receptors located on the prostate. GABAa receptors have been identified on the prostate
(Napoleone etal.. 19901. providing a potential mechanism by which RDX could interact directly
with the prostate. However, this would require that the prostate is actively maintained in a non-
inflamed state, mediated by GABA; RDX binding to GABAa receptor-convulsant sites on the prostate
would result in a reduction of the inhibitory effects of the GABA receptor, leading to increased
inflammation flohnson. 2015b). No evidence was found to support this potential pathway leading
to prostate inflammation.
In summary, there are no studies available that inform mechanistically how RDX might lead
to kidney and other urogenital effects. There is evidence that RDX binds to GABAa receptors in
neuronal tissues (Williams etal.. 2011: Williams and Bannon. 20091. and it is biologically plausible
that binding to the GABA receptor could occur in other tissues as well, contributing to the observed
kidney and urogenital effects. Among the mechanistic information presented above, MOAs that
require direct action on the prostate are considered less likely because the available information
suggests that the prostatitis is a secondary effect However, the ways that GABAa receptors work in
non-neuronal tissues and organs is still not well understood, and the MOA by which RDX induces
kidney and other urogenital effects is unknown.
Integration of Kidney and Urogenital System Effects
Evidence for kidney effects resulting from RDX exposure consists of human case reports and
findings of histopathological changes in rodents. In humans, evidence for kidney effects (including
decreased urine output, blood in urine, and proteinuria) is limited to individuals with acute
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accidental exposure (ingestion and inhalation) to unknown amounts of RDX. No RDX-related
changes in kidney parameters were found in a small cross-sectional study of RDX-exposed workers
fHathawav and Buck. 19771.
A dose-related increase in the incidence of suppurative prostatitis in male rats fLevine etal..
1983b) provides the strongest evidence of RDX-associated kidney and other urogenital system
effects. As discussed above, the incidence of suppurative prostatitis is considered to be an indicator
for the broader array of kidney and other urogenital effects seen in this study. Levine etal. (1983b)
identified other histopathological effects (papillary necrosis, pyelitis, luminal distension, and
cystitis) in the kidney and bladder, but at the highest dose only. A second 2-year study in Sprague-
Dawley rats found no histopathological changes in the kidney or urogenital system fHart. 19761.
but exposure levels used in this study were low compared to Levine etal. (1983b). Other measures
of kidney effects, specifically kidney weights and serum chemistry parameters, did not provide
consistent evidence of dose-related changes associated with RDX exposure. In light of the dose-
related increase in suppurative prostatitis and the lack of support for an alternative (i.e., non-RDX-
related) basis for this effect, kidney and urogenital effects are a potential human hazard of RDX
exposure.
1.2.3. Reproductive and Developmental Effects
No human studies were identified that evaluate the potential of RDX to cause reproductive
or developmental effects. Information relevant to an examination of the association between RDX
exposure and reproductive and developmental effects comes from a 2-generation reproductive
toxicity study in rats and developmental studies in rats and rabbits involving oral administration of
RDX during gestation. In addition, oral subchronic and chronic studies in experimental animals
provide information useful for examining the association between RDX exposure and effects
specifically on the male reproductive system. A summary of the reproductive and developmental
effects associated with RDX exposure is presented in Tables 1-9 and 1-10 and Figures 1-3 and 1-4.
Studies are ordered in the evidence tables and exposure-response arrays by duration of exposure
and then by species.
Reproductive Effects
Evidence of male reproductive toxicity is provided by the finding of testicular degeneration
in male mice. An increased incidence of testicular degeneration (10-11%) was observed in male
B6C3Fi mice exposed to >35 mg/kg-day RDX for 2 years in the diet compared to concurrent (0%)
and historical (1.5%) controls (Lish etal.. 1984). Reductions in absolute testicular weight were
observed, but the magnitude of this effect was small (<6% compared to controls) and not dose-
related. An increased incidence of germ cell degeneration was observed in rats exposed to
40 mg/kg-day (40%) compared with controls at 12 months (0%); by 24 months, almost all male
rats (including controls) had testicular masses (interstitial cell tumors), and no instances of germ
cell degeneration were identified in control or RDX-treated groups fLevine etal.. 1983bl. No dose-
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related histopathological changes in the testes were identified in other studies in rats (Grouse etal..
2006: Levine etal.. 1990: Levine etal.. 1981a. b; Hart. 19761 or dogs (Hart. 19741. Changes in
testicular weight were inconsistent across studies, with an equivalent number of studies identifying
decreases (Grouse etal.. 2006: Lish etal.. 1984: Cholakis etal.. 19801 or increases (Levine etal..
1990: Levine etal.. 1981a. b; Cholakis etal.. 1980: Hart. 1976.19741 in testicular weight; in most
cases, the changes in testicular weight were small (<10% change compared to control) and not
dose-related.
Reproductive function was assessed in two separate studies reported by Cholakis et al.
(19801. No specific effects on reproductive function were observed in F0 and F1 CD rats exposed to
<16 mg/kg-day RDX in the Cholakis etal. f 19801 two-generation study. The highest dose tested,
50 mg/kg-day, was associated with reductions in fertility (specifically a decreased number of
pregnancies) in the F0 generation, although these changes were not statistically significant. The
finding of lower fertility rates only at the 50 mg/kg-day dose, a dose associated with reduced body
weight and feed consumption and increased mortality (9% in male rats and 27% in female rats),
suggests that effects on reproductive function were likely due to the general toxicity of RDX rather
than a direct effect of RDX on reproduction. In the dominant lethal mutation study, which used the
F0 males from the two-generation reproductive toxicity study, no effects on fertility were observed
in male rats exposed to <16 mg/kg-day RDX. Pregnancy rates were lower in untreated females
mated to males exposed to 50 mg/kg-day RDX for 15 weeks prior to mating; the authors attributed
this effect to a treatment-related decrease in the well-being of the males in this high-dose group
(Cholakis etal.. 19801.
Developmental Effects
Animal studies have reported decreases in offspring survival following administration of
RDX. Pup survival rates in the F0 and F1 generations (including both stillborn pups and postnatal
deaths through the age of weaning) were statistically significantly decreased in RDX-exposed CD
rats compared to controls in the only available two-generation reproductive toxicity study of RDX
(Cholakis etal.. 19801. This observation was noted only at the highest dose tested (50 mg/kg-day)
that also produced toxicity in adults (mortality [18%], reduced body weights [8-14%], and reduced
food consumption [10-17%]). Decreased fetal viability was observed atthe highest dose tested, 20
mg/kg-day, in a developmental toxicity study in F344 rats fCholakis etal.. 19801. although no effect
on live fetuses was observed in a developmental toxicity study in Sprague-Dawley rats at the same
dose fAngerhofer et al.. 19861: both of these studies reported significant mortality (29-31%) in
dams at 20 mg/kg-day. Increased resorptions were similarly limited to the highest dose tested (20
mg/kg-day) fCholakis etal.. 19801. Both of these studies started treatment with RDX on gestational
day (GD) 6, which may contribute to the incidence of resorptions observed in the control and
treated groups. As noted in EPA's Guidelines for Developmental Toxicity Risk Assessment (U.S. EPA.
19911. treatment beginning around the time of implantation may result in an increase in
implantation loss that reflects variability that is not treatment related. There was no evidence of
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maternal toxicity, embryotoxicity, or decreased fetal viability in a teratology study of pregnant New
Zealand White (NZW) rabbits administered RDX by gavage from GD 7 to 29 at doses up to 20
mg/kg-day fCholakis etal.. 19801. suggesting that rabbits may be less sensitive to RDX toxicity than
rats.
Statistically significant, dose-related reductions in fetal body weight and length were
reported in Sprague-Dawley rats administered RDX by gavage from GD 6 to 15 fAngerhofer etal..
19861.9 Decreased fetal body weight (9%) and body length (5%), with statistically significant
trends, were observed at 20 mg/kg-day, a dose that produced significant (31%) mortality in the
dams. A similar reduction in fetal body weight of 7% (not statistically significant) was observed in
F344 rats exposed to RDX at 20 mg/kg-day, a dose associated with 29% maternal mortality
fCholakis etal.. 19801. Dose-related reductions in fetal body weight were not observed in NZW
rabbits at doses up to 20 mg/kg-day fCholakis etal.. 19801.
No treatment-related effects on morphological development have been reported in rats
exposed to a dose as high as 20 mg/kg-day RDX, a dose that resulted in 29-31% maternal mortality
(Angerhofer et al.. 1986: Cholakis etal.. 19801. Examination of rabbits administered RDX at doses
up to 20 mg/kg-day from GD 7 to 29 also provided no evidence of treatment-related developmental
anomalies fCholakis etal.. 19801. Although increased incidences of enlarged frontal fontanel and
unossified sternebrae were observed in fetuses of all groups of NZW rabbits administered RDX
fCholakis etal.. 19801. these developmental anomalies did not exhibit a dose-related increase in the
number of either fetuses or litters affected, and were thus interpreted as not being treatment-
related by the study authors fCholakis etal.. 19801. This interpretation is supported by the
following additional considerations. Neither individual litter data nor historical control data from
the performing laboratory were available to assist in the interpretation of these findings. A report
of historical control incidences of fetal skeletal observations in NZW rabbits for 224 prenatal
developmental toxicology studies conducted in 8 contract research laboratories during the period
of 1988-1992 fMTA. 19921 included findings from 26,166 fetuses of 3,635 litters. Background
control incidences of enlarged anterior fontanel were observed in 8 fetuses (0.031%) of 7 litters
(0.193%), while sternebrae agenesis (which may not be entirely comparable to the finding of
unossified sternebrae in Cholakis etal. (19801 was found in 10 fetuses (0.038%) of 5 litters
(0.138%). Although the use of concurrent control data is preferable for the interpretation of
developmental toxicity data, this historical information supports the low control incidences of these
findings in the Cholakis etal. f 19 8 01 study as being within typical historical parameters. It is also
noted that the non-dose-related pattern of increased enlarged fontanel and unossified sternebrae
across treated groups in Cholakis etal. f 19 8 01 was similar to the pattern of decreases in fetal body
9The statistical analyses presented by the study authors were performed on a per fetus basis; EPA's Guidelines
for Developmental Toxicity Risk Assessment (U.S. EPA. 19911 recommend that fetal data be analyzed on a per
litter (rather than per fetus) basis. In a reanalysis of the Angerhofer et al. (19861 data by EPA on a per litter
basis, fetal body weight and length showed statistically significant decreasing trends.
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1 weight in the same study, suggesting a possible link between these particular sternebral and
2 fontanel anomalies with fetal growth status. Given the lack of dose-related increases in the
3 incidences of these anomalies, and patterns that mirrored fetal body weight decreases (which were
4 also not dose-related), the findings of enlarged frontal fontanel and unossified sternebrae were not
5 considered treatment-related. Gestational administration of RDX to NZW rabbits did not result in
6 any other dose- and treatment-related skeletal abnormalities.
7 Table 1-9. Evidence pertaining to male reproductive effects in animals
Reference and study design
Results
Lish etal. (1984)
Mice, B6C3Fi, 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 urn
0,1.5, 7.0, 35, or 175/100 mg/kg-d (high
dose reduced to 100 mg/kg-d in wk 11
due to excessive mortality)
Diet
2 yrs
Doses
0 1.5 7.0 35 175/100
Testicular degeneration (incidence)
0/63 2/60 2/62 6/59 3/27a
Absolute testes weight; wk 105 (percent change compared to
control)
0% -6% 0% -2% -6%
Relative testes weight; wk 105 (percent change compared to control)
0% -4% 2% -2% -2%
Hart (1976)
Rats, Sprague-Dawley, 100/sex/dose
Purity and particle size not specified
0,1.0, 3.1, or 10 mg/kg-d
Diet
2 yrs
Doses
o
1
1
rn
o
O
Absolute testes (with epididymis) weight; wk 104
0% -2% 2% 5%
Relative testes (with epididymis) weight; wk 104
0% -1% 1% 9%
Testes were examined microscopically in control and 10 mg/kg-d
groups; no degeneration or other treatment-related effects were
observed.
Levine et al. (1983b)
Rats, F344, 75/sex/group; interim
sacrifices (10/sex/group) at 6 and 12 mo
89.2-98.7% pure, with 3-10% HMX as
contaminant; 83-89% of particles <66 nm
0, 0.3,1.5, 8.0, or 40 mg/kg-d
Diet
2 yrs
Doses
o
o
00
LO
m
o
o
Testes, germ cell degeneration; 12 mob (incidence)
SS
SDMS
0/10 0/10 0/10 0/10 4/10*
1/3 - 4/19
Testes, germ cell degeneration; 24 mo (incidence)
SS
SDMS
0/38 0/36 0/25 0/29 0/4
0/16 0/19 0/27 0/26 0/27
Testes weights were not measured at termination due to testicular
masses in nearly all males.
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Reference and study design
Results
Cholakis et al. (1980)
Mice, B6C3Fi, 10-12/sex/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 urn particle size
Experiment 1: 0,10,14, 20, 28, or
40 mg/kg-d
Diet
13 wks
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)0
Diet
13 wks
Doses
0 10 14 20 28 40
Absolute testes weight (percent change compared to control)
0s-
1
0s-
1
1
1
1
0s-
o
Relative testes weight (percent change compared to control)
0s-
1
1
0s-
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2
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5
6
7
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Reference and study design
Results
Crouse et al. (2006)
Rats, F344,10/sex/group
99.99% pure
0, 4, 8,10,12, or 15 mg/kg-d
Gavage
13 wks
Doses
0 4 8 10 12 15
Absolute testes weight (percent change compared to control)
0% -3% -5% -4% -4% -8%
Relative testes weight (percent change compared to control)
0% 4% 5% 0% -6% -10%*
Levine et al. (1981a); Levine et al. (1990);
Levine et al. (1981b)d
Rats, F344,10/sex/group; 30/sex for
control
84.7 ± 4.7% purity, ~10% HMX, median
particle diameter 20 nm, ~90% of particles
<66 nm
0,10, 30,100, 300, or 600 mg/kg-d
Diet
13 wks
Doses
0 10 30 100 300 600
Testes, germ cell degeneration (incidence)
0/10 0/10 0/10 0/10 1/9 1/10
Absolute testes weight (percent change compared to control)
0% 1% 1% -2%
Relative testes weight (percent change compared to control)
0% 4% 5% 19%*
Hart (1974)e
Dogs, Beagle, 3/sex/dose
Pre-mix with ground dog chow containing
20 mg RDX/g-chow, 60 g dog food; purity
and particle size not specified
0, 0.1,1, or 10 mg/kg-d
Diet
13 wks
Doses
o
1
1
1
o
o
Absolute testes (with epididymis) weight (percent change compared
to control)
0% - - 51%
Testes were not examined microscopically.
^Statistically significant (p < 0.05) based on analysis by study authors.
aAlthough the study authors did not observe a statistically significant increase in the incidence of testicular
degeneration, they determined that the incidences at the 35 and 175/100 mg/kg-day dose groups were "notable"
when compared to concurrent (0%) and historical (1.5%) incidences.
testicular atrophy was observed at 12 months along with a statistically reduced mean testes weight (compared
with controls). By 24 months, almost all male rats (including controls) had testicular masses (interstitial cell
tumors); testes weights were not recorded, and an increased incidence of testicular degeneration was not
observed.
cDoses were calculated by the study authors.
dLevine et al. (1981a) is a laboratory report of a 13-week study of RDX in F344 rats; two subsequently published
papers (Levine et al., 1990; Levine et al., 1981b) present subsets of the data provided in the full laboratory report.
eBecause testes weight was reported for only three treated animals in this study, organ data from this study were
considered less informative than other studies; therefore, testes weights from Hart (1974) were not presented in
the exposure-response array for male reproductive effects.
Note: A dash ("-") indicates that the study authors did not measure or report a value for that dose group.
SDMS - spontaneous death or moribund sacrifice; SS -- scheduled sacrifice
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1000
100
10
• signifcantly changed O not significantly changed
O
o
o
o
>
fl3
13
m
JE
03
O
U
a
0Q
9
4^ Absolute testes weight
s
c
ra
I
75
4-»
1
s
3
1
O
o
m
j=
00
8
o
9
tH
00
-J
s
TO
03
«
05
w
C
tu
s
m
">
to
2
o
JZ
—i
J2
u
o
JC
u
m
m
m
MI
J2
o
u
CO
a
Testicular degeneration
Note: Filled circle indicates that response was statistically significantly different from the control.
(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, they determined that the incidences at the 35 and l/b/100 trig/kg day dose groups were "notable" when
compared to concurrent (0%) and historical {1.5%) incidences.
Figure 1-3. Exposure response array of male reproductive effects following oral exposure.
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Table 1-10. Evidence pertaining to reproductive and developmental effects in
animals
Reference and study design
Results
Offspring survival
Cholakis et al. (1980)
Doses
0 5
16
50
Rats, CD, two-generation study;
F0: 22/sex/group; Fl: 26 sex/group;
F2: 10 sex/group
Stillborn pups (incidence)
Fl
8/207 6/296
4/259
16/92*
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 urn particle size
F0 and Fl parental animals: 0, 5,16, or
F2
6/288 6/290
2/250
24/46*
Offspring survival at birth (percent of fetuses)
50 mg/kg-d
Diet
Fl
96% 98%
98%
83%*
F0 exposure: 13 wks pre-mating, and
F2
98% 98%
99%
48%*
during mating, gestation, and lactation of
Fl; Fl exposure: 13 wks after weaning,
Survival at weaning (percent ofliveborn pups)
and during mating, gestation, and
Fl
87% 96%
90%
8%
lactation of F2; F2 exposure: until weaning
F2
79% 86%
79%
0%
F0 maternal deaths occurred at 50 mg/kg-d. Only six Fl females in
this group survived to serve as parental animals; none of the
surviving six died during subsequent treatment.
Note: results on a per litter basis were not provided.
Cholakis et al. (1980)
Doses
0 0.2
2
20
Rabbits, NZW, 11-12/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 nm particle size
Early resorptions (mean percent per dam)
6% 5%
4%
1%
0, 0.2, 2.0, or 20 mg/kg-d
Gavage
GDs 7-29
Late resorptions (mean percent per dam)
8% 5%
3%
3%
Viable fetuses (mean percent per dam)
85% 82%
77%
94%
Cholakis et al. (1980)
Doses
0 0.2
2.0
20
Rats, F344, 24-25 females/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants.
Early resorptions (mean percent per dam)
6.0% 2.5%
4.8%
15.3%
0, 0.2, 2.0, or 20 mg/kg-d
Gavage
GDs 6-19
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.
This document is a draft for review purposes only and does not constitute Agency policy.
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Reference and study design
Results
Angerhofer et al. (1986)
Rats, Sprague-Dawley, 39-51 mated
females/group (25-29 pregnant
dams/group)
Purity 90%; 10% HMX and 0.3% acetic acid
occurred as contaminants
0, 2, 6, or 20 mg/kg-d
Gavage
GDs 6-15
Doses
0 2 6 20
Resorptions (percent of total implantations)
4.8% 6.1% 5.9% 6.4%
Early resorptions (percent of total implantations)
4.8% 6.1% 5.9% 6.2%
Late resorptions (percent of total implantations)
0% 0% 0% 0.27%
Live fetuses (mean percent per litter)
100% 100% 100% 100%
Significant maternal mortality (16/51) occurred at 20 mg/kg-d.
Percent resorptions and live fetuses based on number of surviving
females at time of necropsy.
Offspring growth
Cholakis et al. (1980)
Rabbits, NZW, 11-12/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 urn particle size
0, 0.2, 2.0, or 20 mg/kg-d
Gavage
GDs 7-29
Doses
0 0.2 2.0 20
Fetal body weight (percent change compared to control)
0% -6.7% -2.3% -9.3%
Cholakis et al. (1980)
Rats, F344, 24-25 females/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants.
0, 0.2, 2.0, or 20 mg/kg-d
Gavage
GDs 6-19
Doses
0 0.2 2.0 20
Fetal body weight (percent change compared to control)
0% 2% 3% -7%
Significant maternal mortality (7/24 dams) occurred at 20 mg/kg-d.
Angerhofer et al. (1986)
Rats, Sprague-Dawley, 39-51 mated
females/group (25-29 pregnant
dams/group)
Purity 90%; 10% HMX and 0.3% acetic acid
occurred as contaminants
0, 2, 6, or 20 mg/kg-d
Gavage
GDs 6-15
Doses
0 2 6 20
Fetal body weight (percent change compared to control)
0% -4% -2% -9%a
Fetal body length (percent change compared to control)
(O
0s-
LO
1
0s-
1
1
0s-
1
1
0s-
o
Significant maternal mortality (16/51) occurred at 20 mg/kg-d.
This document is a draft for review purposes only and does not constitute Agency policy.
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Reference and study design
Results
Morphological development
Cholakis et al. (1980)
Rabbits, NZW, 11-12/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants; ~200 nm 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
Litters
0/88 0/99 0/94 3/110
0/11 0/11 0/11 2/12
Misshapen eye bulges (incidence)
Fetuses
Litters
0/88 0/99 0/94 3/110
0/11 0/11 0/11 1/12
Cleft palate (incidence)
Fetuses
Litters
0/39 1/46 2/44 2/52
0/11 1/11 1/11 1/12
Enlarged front fontanel (incidence)
Fetuses
Litters
0/49 5/53 2/50 8/58
0/11 2/11 2/11 2/12
Unossified sternebrae (incidence)
Fetuses
Litters
4/49 12/53 8/50 12/58
4/11 7/11 4/11 6/12
Cholakis et al. (1980)
Rats, F344, 24-25 females/group
88.6% pure, with 9% HMX and 2.2% water
as contaminants.
0, 0.2, 2.0, or 20 mg/kg-d
Gavage
GDs 6-19
No gross or soft-tissue anomalies were seen in any exposure group.
No treatment-related increase in the incidence of litters with skeletal
anomalies was observed. Significant maternal mortality (7/24 dams)
occurred at 20 mg/kg-d.
Angerhofer et al. (1986)
Rats, Sprague-Dawley, 39-51 mated
females/group (25-29 pregnant
dams/group)
Purity 90%; 10% HMX and 0.3% acetic acid
occurred as contaminants
0, 2, 6, or 20 mg/kg-d
Gavage
GDs 6-15
No treatment-related increase in the incidence of anomalies was
observed.
Doses
0 2 6 20
Total malformations (percent of fetuses with malformations)
1% 1% 0% 2%
Significant maternal mortality (16/51) occurred at 20 mg/kg-d.
1
2 ^Statistically significant (p < 0.05) based on analysis by study authors.
3 statistically significant dose-related trend (p < 0.05) by linear trend test, performed for this assessment. Average
4 fetal weights or lengths for each litter comprised the sample data for this test.
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100
>
03
72
CUD
E
o
a
10
0.1
• signficantly changed
O not signifcantly changed
o
00
01
o
_c
u
si/ Offspring survival
o
oo
CD
o
_c
u
Offpsring growth
Note: Filled circle indicates that response was statistically significantly different from the control.
(1) Statistically signficant dose-related trend (p <= 0.05) by linear trend test, performed for this assessment.
o
oo
CD
o
_c
u
Morphological
development
Figure 1-4. Exposure response array of reproductive and developmental effects following oral exposure.
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Integration of Reproductive and Developmental Effects
Testicular effects were reported in male B6C3Fi mice chronically exposed to RDX in the diet
for 24 months fLish etal.. 19841. No other studies of equivalent duration were performed in mice
to determine the consistency of this effect Germ cell degeneration was observed in F344 rats at
12 months, but not at 24 months, in a 2-year study fLish etal.. 19841: therefore, the biological
significance of the 12-month findings is uncertain. Other testicular effects were inconsistent across
rat studies. Based on the evidence of testicular degeneration in male mice reported by Lish et al.
(19841. there is suggestive evidence of male reproductive effects associated with RDX exposure.
Developmental studies in rats fAngerhofer etal.. 1986: Cholakis etal.. 19801 demonstrated
effects on offspring survival, growth, and morphological development only at doses associated with
severe maternal toxicity and mortality. No dose-related developmental effects were observed in
rabbits fCholakis etal.. 19801. As noted in EPA's Guidelines for Developmental Toxicity Risk
Assessment (U.S. EPA. 19911. where adverse developmental effects are produced only at doses that
cause minimal maternal toxicity, developmental effects should not be discounted as being
secondary to maternal toxicity; however, at doses causing excessive toxicity, as is the case with
RDX, information on developmental effects may be difficult to interpret and of limited value.
Therefore, at this time, no conclusions are drawn regarding developmental effects as a human
hazard of RDX exposure.
1.2.4. Liver Effects
One occupational epidemiology study examined the association between RDX exposure and
changes in serum liver enzymes. Case reports involving accidental exposure to RDX provide
information on the potential for acute exposure to RDX to affect the liver in humans. In addition,
organ weight, histopathology, and serum chemistry findings from experimental animal studies
involving subchronic and chronic exposure to ingested RDX provide data relevant to an
examination of the association between RDX exposure and liver effects. A summary of the liver
effects associated with RDX exposure is presented in Tables 1-11 and 1-12 and Figure 1-5.
Experimental animal studies are ordered in the evidence table and exposure-response array by
duration of exposure and then by species.
Reports in humans provide inconsistent evidence of liver toxicity associated with acute
exposure to RDX. Elevated serum levels of aspartate aminotransferase (AST) and/or alanine
aminotransferase (ALT) were reported in several case reports of individuals who ingested
unknown amounts of RDX (Kiiciikardali etal.. 2003: Woody etal.. 1986: Knepshield and Stone.
1972: Hollander and Colbach. 1969: Stone etal.. 1969: Merrill. 19681 (see Appendix C, Section C.2).
Liver biopsies did not reveal any abnormal observations (Stone etal.. 19691. In other case reports,
no significant changes in serum levels of liver enzymes were observed fTestud etal.. 1996a: Ketel
and Hughes. 19721. In a cross-sectional epidemiologic study of workers from five U.S. Army
munitions plants (69 exposed to RDX alone and 24 to RDX and HMX; RDX exposure range:
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undetectable (<0.01 mg/m3) to 1.6 mg/m3) (Hathaway and Buck. 19771. serum chemistry analysis
(including the serum liver enzymes AST, ALT, and alkaline phosphatase [ALP]) revealed no
statistically significant differences between exposed and unexposed workers (Table 1-11).
In experimental animals, some, but not all, subchronic studies reported increased liver
weight associated with RDX exposure (Table 1-12 and Figure 1-5). Dose-related increases in
relative liver weight10 (11-25% in high-dose groups) were observed in male and female B6C3Fi
mice given RDX in the diet for 90 days fCholakis etal.. 1980) and in female F344 rats in two
separate 90-day dietary studies of RDX (Levine etal.. 1990: Levine etal.. 1981a. b; Cholakis etal..
1980): however, relative liver weights were not increased in female F344 rats in another 90-day
gavage study fCrouse etal.. 20061. Male F344 rats exhibited an increase in relative liver weight
only in one of these subchronic studies (Levine etal.. 1990: Levine etal.. 1981a. b). In subchronic
studies in other species, absolute liver weights were increased in male and female monkeys
(6-16% relative to control at 1 and 10 mg/kg-day) f Martin and Hart. 19741 and in male, but not
female, beagle dogs (53% relative to control in male dogs at 10 mg/kg-day) (Hart. 1974).
Chronic RDX exposures in B6C3Fi mice and F344 or Sprague-Dawley rats showed a less
consistent pattern of liver weight increases. Interpretation of liver weight increases in the 2-year
mouse study is complicated by the incidence of adenomas and carcinomas in each dose group; the
apparent increase in liver weights in male and female mice exposed to RDX in diet (Lish etal.. 1984)
was reduced when mice with liver adenomas or carcinomas were removed from the analysis. In a
2-year rat study fLevine etal.. 1983bl. relative liver weights were increased in high-dose
(40 mg/kg-day) males and females (by 11 and 18% compared to controls, respectively), likely
reflecting the depressed weight gain in the high-dose rats (2-30% in males and 10-15% in
females). In evaluating organ weight data across studies of all durations, less weight is placed on
evidence of organ weight changes from chronic (2-year) studies because normal physiological
changes associated with aging and intercurrent disease contributes to inter-animal variability that
could confound organ weight interpretation (Sellers etal.. 20071. as is true of the mouse liver
weight data for RDX.
Nonneoplastic histopathological changes in the liver were not associated with RDX
exposure in the majority of experimental animal studies fCrouse etal.. 2006: Levine etal.. 1990:
Lish etal.. 1984: Levine etal.. 1983b: Levine etal.. 1981a. b; Hart. 1976.1974: Martin and Hart.
1974). including 2-year oral studies in mice at doses up to 100 mg/kg-day (Lish etal.. 1984) and in
rats at doses up to 40 mg/kg-day fLevine etal.. 1983bl. The few findings of liver lesions were
10Based on an evaluation of the relationship between organ weight and body/brain weight to determine
which endpoint (organ weight, organ-to-body weight ratio, or organ-to-brain weight ratio) is likely to more
accurately detect target organ toxicity, Bailey etal. (20041 concluded that evaluation of the effects of a test
chemical on liver weight are optimally analyzed using organ-to-body weight ratios. Therefore, the analysis of
liver weight here focuses on relative weight data where study authors reported both relative and absolute
weights, although both relative and absolute data are summarized in the evidence table (Table 1-12).
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Toxicological ReviewofHexahydro-l,3,5-trinitro-l,3,5-triazine
reported in studies with more limited histopathological analyses, and were not confirmed in the
studies with more complete histopathologic examination and longer exposure durations fLish etal..
1984: Levine etal.. 1983bl. For example, the incidence of liver portal inflammation was increased
in female rats, but not male rats, exposed to 40 mg/kg-day in the diet for 90 days fCholakis etal..
19801. There was an increase in the incidence of mild liver microgranulomas in female mice only
fCholakis etal.. 19801 and karyomegaly of hepatocytes in male mice only exposed to
320 mg/kg-day RDX in the diet for 90 days fCholakis etal.. 19801. Because both the rat and mouse
studies by Cholakis etal. (19801 used relatively small group sizes (n = 10/sex/group) and provided
histopathologic findings for the control and high-dose groups only, less weight is placed on these
findings than on those from the 2-year bioassays. It should be noted that exposure to HMX, the
primary contaminant in several of the RDX studies, was associated with histopathological changes
in the livers of male rats fed doses >450 mg/kg-day for 13 weeks. However, similar findings were
not observed in the RDX studies, where the doses of RDX employed in the studies would have
resulted in HMX exposures of <60 mg/kg-day. The contribution of HMX exposure to the overall
liver findings in the studies of RDX toxicity is therefore expected to be negligible.
Clinical chemistry parameters, including serum ALT, AST, and ALP, showed no treatment-
related changes indicative of liver toxicity. Statistically significant changes in these parameters in
some subchronic and chronic toxicity studies in rats and mice were relatively small (generally
<50% of the control mean), were not dose-related in most instances, and showed no consistent
pattern of change between sexes or across studies.
Some subchronic and chronic oral toxicity studies in rats and mice reported dose-related
changes in serum cholesterol and triglyceride levels; however, these changes were not consistently
observed in males and females within the same study, and patterns of changes were not consistent
across studies. Specifically, serum triglyceride levels were elevated (up to 41%) in female B6C3Fi
mice exposed to RDX in the diet for 2 years, although increases were not dose-related (Lish etal..
19841: male mice in the same study did not show a similar increase in triglycerides. In contrast,
serum triglycerides showed dose-related decreases in male and female F344 rats (50-62% at the
high doses) in a subchronic oral (dietary) study (Levine etal.. 1990: Levine etal.. 1981a. b). In a
chronic toxicity study by the same investigators (Levine etal.. 1983b). serum triglyceride levels
were generally decreased in male and female rats (52 and 51%, respectively, at the highest dose of
40 mg/kg-day); however, triglyceride levels across the four dose groups in this study did not show
a dose-related response.
Serum cholesterol levels showed a dose-related increase (38% at the high dose of
100 mg/kg-day) in female B6C3Fi mice exposed to RDX in the diet for 2 years fLish etal.. 19841:
however, changes in cholesterol in male mice in the same study were not dose related. Changes in
serum cholesterol in male and female F344 rats exposed to RDX in the diet for 2 years at doses up
to 40 mg/kg-day (Levine etal.. 1983b). in rats exposed to RDX by gavage for 90 days at doses up to
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1 15 mg/kg-day (Grouse etal.. 20061. and in monkeys exposed to RDX in the diet for 90 days f Martin
2 and Hart. 19741 were relatively small (within 38% of control mean) and were not dose related.
3 Table 1-11. Evidence pertaining to liver effects in humans
Reference and study design
Results
Hathawav and Buck (1977) (United States)
Mean laboratory values of liver enzymes in men (mean; standard
Cross-sectional study, 2,022 workers,
deviation not reported)
1,491 participated (74% response rate).
Analysis group: limited to whites;
RDX exposed*
69 exposed to RDX alone and 24 exposed
Referent
Undetected (0.01 mg/m3
to RDX and HMX; 338 not exposed to RDX,
Test
(n = 237)
(n = 22)
LO
II
C
HMX, or TNT.
Exposure measures: Exposure
LDH
173
191
174
determination based on job title and
ALP
82
78
80
industrial hygiene evaluation. Exposed
subjects assigned to two groups: 0.01 mg/m3 (mean for employees with
AST (SGPT)
21
26
18
exposures >LOD: 0.28 mg/m3).
Effect measures: Liver function tests.
Bilirubin
0.5
0.4
0.4
Analysis: Types of statistical tests were not
includes both workers exposed to RDX alone and RDX and HMX.
reported (assumed to be t-tests for
No differences were statistically significant as reported by study
comparison of means and x2 tests for
authors. Similar results in women.
comparison of proportions).
Liver function tests in men (prevalence of abnormally elevated
values)
Test
RDX exposed*
(abnormal
range)
Referent
Undetected (0.01 mg/m3
LDH (>250)
2/237
1/22
0/45
ALP (>1.5)
34/237
1/22
6/45
AST (SGOT)
20/237
4/22
2/45
(>35)
ALT (SGPT)
15/237
2/22
0/45
(>35)
Bilirubin
5/237
1/22
1/45
(>1.0)
includes both workers exposed to RDX alone and RDX and HMX.
No differences were statistically significant as reported by study
authors. Similar results in women.
4 LDH = lactate dehydrogenase; SGOT = glutamic oxaloacetic transaminase; SGPT = glutamic pyruvic transaminase
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Toxicological ReviewofHexahydro-l,3,5-trinitro-l,3,5-triazine
Table 1-12. Evidence pertaining to liver effects in animals
Reference and study design
Results
Liver weight
Lish etal. (1984)
Mice, B6C3Fi, 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 urn
0,1.5, 7.0, 35, or 175/100 mg/kg-d
(high dose reduced to 100 mg/kg-d in
wk 11 due to excessive mortality)
Diet
2 yrs
Doses
0 1.5 7.0 35 175/100
Absolute liver weight at 104 wks (percent change compared to control)
M
F
0% 28%* 11% 12% 35%*
0% 7% 7% 15% 18%*
Relative liver weight at 104 wks (percent change compared to control)
M
F
0% 32%* 12% 14% 46%*
0% 6% 8% 18% 45%*
Note: Percent change in liver weights of male and female mice was reduced
in all dose groups when mice with liver tumors were removed from the
analysis, suggesting no real effect on liver weight.
Hart (1976)
Rats, Sprague-Dawley, 100/sex/group
Purity and particle size not specified
0, 1.0, 3.1, or 10 mg/kg-d
Diet
2 yrs
Doses
o
1
1
rn
o
O
Absolute liver weight (percent change compared to control)
M
F
0% -6% -6% -6%
0% 7% -11% 1%
Relative liver weight (percent change compared to control)
M
F
0% -5% -2% -3%
0% 17% -2% 13%
Levine et al. (1983b)
Rats, F344, 75/sex/group; interim
sacrifices (10/sex/group) at 6 and
12 mo
89.2-98.7% pure, with 3-10% HMX as
contaminant; 83-89% of
particles <66 nm
0, 0.3,1.5, 8.0, or 40 mg/kg-d
Diet
2 yrs
Doses
o
o
00
LO
m
o
o
Absolute liver weight at 105 wks (percent change compared to control)
M
F
0% 3% -7% 1% -8%
0% 1% -4% 3% 0%
Relative liver weight at 105 wks (percent change compared to control)
M
F
0% 1% 0% 2% 11%
0% 1% -2% 6% 18%*
Cholakis et al. (1980)
Mice, B6C3Fi, 10-12/sex/group
88.6% pure, with 9% HMX and 2.2%
water as contaminants; ~200 nm
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
F
0% - - - -6% -5%
0% - - - -4% -1%
Relative liver weight (percent change compared to control)
M
F
0% - - - -4% -4%
0% - -6% 1%
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Toxicological ReviewofHexahydro-l,3,5-trinitro-l,3,5-triazine
Reference and study design
Results
Experiment 2: 0, 40, 60, or 80 mg/kg-d
for 2 wks followed by 0, 320,160, or
80 mg/kg-d (TWA doses of 0, 79.6,
147.8, or 256.7 mg/kg-d for males and
0, 82.4,136.3, or 276.4 mg/kg-d for
females)3
Diet
13 wks
Doses
0 80 160 320
Absolute liver weight (percent change compared to control)
M
F
0% 2% 12% 26%*
0% 4% 9% 29%*
Relative liver weight (percent change compared to control)
M
F
0% 0% 9% 25%*
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 nm
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
F
0% - -2% -5%
0% - - - 6% 4%
Relative liver weight (percent change compared to control)
M
F
0% - - - 2% 3%
0% - 10% 11%
Cholakis et al. (1980)
Rats, CD, two-generation study; F0:
22/sex/group; Fl: 26/sex/group;
F2:10/sex/group
88.6% pure, with 9% HMX and 2.2%
water as contaminants; ~200 nm
particle size
F0 and Fl parental animals: 0, 5,16, or
50 mg/kg-d
Diet
F0 exposure: 13 wks pre-mating, and
during mating, gestation, and lactation
of Fl; Fl exposure: 13 wks after
weaning, and during mating, gestation,
and lactation of F2; F2 exposure: until
weaning
Doses
0 5 16 50
Absolute liver weight (percent change compared to control)
M
F
0% 7% -16%
0% 0% -14%
Crouse et al. (2006)
Rats, F344,10/sex/group
99.99% pure
0, 4, 8,10,12, or 15 mg/kg-d
Gavage
13 wks
Doses
0 4 8 10 12 15
Absolute liver weight (percent change compared to control)
M
F
0% -6% -9% 0% 7% 5%
0% 1% 7% 18%* 15% 28%*
Relative liver weight (percent change compared to control)
M
F
0% 0% -1% 2% 5% 2%
0% 1% -2% 2% -3% 2%
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Toxicological ReviewofHexahydro-l,3,5-trinitro-l,3,5-triazine
Reference and study design
Results
(Levine et al. (1990); Levine et al.
(1981a), 1981b))b
Rats, F344, 3-4 wks old; 10/sex/group;
30/sex/group for controls
84.7 ± 4.7% purity, ~10% HMX, median
particle diameter 20 urn, ~90% of
particles <66 urn
0, 10, 30,100, 300, or 600 mg/kg-d
Diet
13 wks
Doses
0 10 30 100 300 600
Absolute liver weight (percent change compared to control)
M
F
0% 5% -1% -2%
0% 2% 4% 16%*
Relative liver weight (percent change compared to control)
M
F
0% 9% 6% 20%
0% 3% 5% 19%*
Hart (1974)°
Dogs, Beagle, 3/sex/group
Pre-mix with ground dog chow
containing 20 mg RDX/g-chow, 60 g
dog food; purity and particle size not
specified
0, 0.1,1, or 10 mg/kg-d
Diet
13 wks
Doses
o
1
1
1
o
o
Absolute liver weight (percent change compared to control)
M
F
0% - - 53%
0% - - 3%
Martin and Hart (1974)°
Monkeys, Cynomolgus or Rhesusd,
3/sex/group
Purity of test material not specified
0, 0.1,1, or 10 mg/kg-d
Gavage
13 wks
Doses
o
1
1
1
o
o
Absolute liver weight (percent change compared to control)
M + F
0% 2% 6% 16%
Histopathological lesions
Lish et al. (1984)
Mice, B6C3Fi, 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 nm
0,1.5, 7.0, 35, or 175/100 mg/kg-d
(high dose reduced to 100 mg/kg-d in
wk 11 due to excessive mortality)
Diet
2 yrs
Histopathological lesions in liver other than adenomas and carcinomas
were not significantly different compared to controls, as reported by study
authors.
Hart (1976)
Rats, Sprague-Dawley, 100/sex/group
Purity and particle size not specified
0, 1.0, 3.1, or 10 mg/kg-d
Diet
2 yrs
Histopathological examination performed only for controls and 10 mg/kg-d
rats; no significant differences compared to controls were reported by
study authors.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological ReviewofHexahydro-l,3,5-trinitro-l,3,5-triazine
Reference and study design
Results
Levine et al. (1983b)
Rats, F344, 3-4 wks old; 75/sex/group;
interim sacrifices (10/sex/group) at
6 and 12 mo
89.2-98.7% pure, with 3-10% HMX as
contaminant; 83-89% of particles
<66 urn
0, 0.3,1.5, 8.0, or 40 mg/kg-d
Diet
2 yrs
Doses
o
o
00
LO
m
o
o
Microgranulomas (incidence)
M
F
0/38 0/36 0/25 0/29 0/4
10/43 19/45 12/42 17/41 4/28
Cholakis et al. (1980)
Mice, B6C3Fi, 10-12/sex/group
88.6% pure, with 9% HMX and 2.2%
water as contaminants; ~200 nm
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)3
Diet
13 wks
Doses
0 80 160 320
Liver microgranulomas; mild (incidence)
M
F
2/10 - - 1/9
2/11 - - 7/11*
Increased karyomegaly of hepatocytes (incidence)
M
F
0/10 - - 5/9*
Cholakis et al. (1980)
Rats, F344,10/sex/group
88.6% pure, with 9% HMX and 2.2%
water as contaminants; ~200 nm
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
F
0/10 - - 1/10
Liver portal inflammation (incidence)
M
F
2/10 - - 3/10
1/10 - - 7/10
Crouse et al. (2006)
Rats, F344,10/sex/group
99.99% pure
0, 4, 8,10,12, or 15 mg/kg-d
Gavage
13 wks
Histopathology examination of the 15 mg/kg-d group showed one male
with mild liver congestion and one female with a moderate-sized focus of
basophilic cytoplasmic alteration; neither finding was attributed by study
authors to RDX treatment.
(Levine et al. (1990); Levine et al.
(1981a), 1981b))b
Rats, F344,10/sex/group; 30/sexfor
control
84.7 ± 4.7% purity, ~10% HMX, median
particle diameter 20 nm, ~90% of
particles <66 nm
0, 10, 30,100, 300, or 600 mg/kg-d
Diet
13 wks
Histopathological examination of liver did not reveal any significant
differences compared to controls, as reported by study authors. No
histopathology findings available for the 300 or 600 mg/kg-d dose groups
because all rats in these groups died before the 13-wk necropsy.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological ReviewofHexahydro-l,3,5-trinitro-l,3,5-triazine
Reference and study design
Results
Hart (1974)
Dogs, Beagle, 3/sex/group
Pre-mix with ground dog chow
containing 20 mg RDX/g-chow, 60 g
dog food; purity and particle size not
specified
0, 0.1,1, or 10 mg/kg-d
Diet
13 wks
Histopathological examination performed only for controls and 10 mg/kg-d
dogs; no significant differences compared to controls were reported.
Martin and Hart (1974)
Monkeys, Cynomolgus or Rhesusd,
3/sex/group
Purity of test material not specified
0, 0.1,1, or 10 mg/kg-d
Gavage
13 wks
An increase in the amount of iron-positive material in liver cord cytoplasm
was reported in monkeys treated with 10 mg/kg-d RDX, which the study
authors considered to be of uncertain toxicological significance. Because
iron-positive stain was present in controls and no further characterization
of the staining was provided in the study report, the toxicological
significance of this finding could not be determined.
Serum chemistry
Lish etal. (1984)
Mice, B6C3Fi, 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 nm
0,1.5, 7.0, 35, or 175/100 mg/kg-d
(high dose reduced to 100 mg/kg-d in
wk 11 due to excessive mortality)
Diet
2 yrs
Doses
0 1.5 7.0 35 175/100
Serum cholesterol at 105 wks (percent change compared to control)
M
F
0% 11% -11% 5% 39%
0% 5% 15% 25% 38%
Serum triglycerides at 105 wks (percent change compared to control)
M
F
0% 21% -20% 10% -25%
0% 34% 28% 41% 28%
Levine et al. (1983b)
Rats, F344, 75/sex/group; interim
sacrifices (10/sex/group) at 6 and
12 mo
89.2-98.7% pure, with 3-10% HMX as
contaminant; 83-89% of particles
<66 nm
0, 0.3,1.5, 8.0, or 40 mg/kg-d
Diet
2 yrs
Doses
o
o
00
LO
m
o
o
Serum cholesterol at 104 wks (percent change compared to control)
M
F
0% 15% 38% 19% -6%
0% 6% 3% -7% -9%
Serum triglycerides at 104 wks (percent change compared to control)
M
F
0% 14% -15% -12% -52%
0% 18% 5% -42% -51%*
Crouse et al. (2006)
Rats, F344,10/sex/group
99.99% pure
0, 4, 8,10,12, or 15 mg/kg-d
Gavage
13 wks
Doses
0 4 8 10 12 15
Serum cholesterol (percent change compared to control)
M
F
0% -3% -10%* -16%* -18%* -11%*
0% -1% -8% -4% -4% -1%
Serum triglycerides (percent change compared to control)
M
0% 1% 1% -7% -2% -19%
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Toxicological ReviewofHexahydro-l,3,5-trinitro-l,3,5-triazine
Reference and study design
Results
F
0% -16% -21% 7% -37% 18%
(Levine et al. (1990); Levine et al.
(1981a), 1981b))b
Rats, F344,10/sex/group; 30/sexfor
control
84.7 ± 4.7% purity, ~10% HMX, median
particle diameter 20 nm, ~90% of
particles <66 nm
0, 10, 30,100, 300, or 600 mg/kg-d
Diet
13 wks
Doses
0 10 30 100 300 600
Serum triglyceride levels (percent change compared to control)
M
F
0% -14% -34% -62%*
0% -12% -29% -50%*
Martin and Hart (1974)
Monkeys, Cynomolgus or Rhesusd,
3/sex/group
Purity of test material not specified
0, 0.1,1, or 10 mg/kg-d
Gavage
13 wks
Serum biochemistry analysis revealed scattered deviations, but study
authors indicated they appear to have no toxicological significance.
Doses
o
1
1
1
o
o
Serum cholesterol (percent change compared to control)
M
F
0% -17% -2% -7%
0% 7% 7% 7%
^Statistically significant (p < 0.05) based on analysis by study authors.
aDoses were calculated by the study authors.
bLevine et al. (1981a) is a laboratory report of a 13-week study of RDX in F344 rats; two subsequently published
papers (Levine et al., 1990; Levine et al., 1981b) present subsets of the data provided in the full laboratory report.
cLiver weight data from the Hart (1974) and Martin and Hart (1974) studies were considered less informative than
other studies. Hart (1974) reported organ weight data for high-dose dogs (3/sex/group) only, and the liver weights
from Martin and Hart (1974) were highly variable across monkeys (e.g., liver weights for the control animals
ranged from 46 to 141 g). Therefore, liver weight data from these two studies were not presented in the
exposure-response array for liver effects (Figure 1-5).
dThe species of monkey used in this study was inconsistently reported in the study as either Cynomolgus (in the
methods section) or Rhesus (in the summary).
Note: A dash ("-") indicates that the study authors did not measure or report a value for that dose group.
This document is a draft for review purposes only and does not constitute Agency policy.
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ToxicologicalReviewofHexahydro-l,3,5-trinitro-l,3,5-triazine
• significantly changed O not significantly changed x not determined
¦&— -e
o «
<> '
I
I
I
1
I
G) t
i
111!
KD
h*
u\
8
Chronic
$ %
o
£
o
00
3
o
o
0D
a
I
o
00
a
+- ¦*- 2
¦
J
c
o
£
r-.
cn
rH
t
ro
X
Chronic Subchronic
4- Cholesterol
eu
CL5
Chronic Subchr
4- Trigylcerides
Serum biochemistry changes
Note; Filled circle indicates that response was statistically significantly different from the control.
X - Not considered due to confounding caused by presence of tumors.
Figure 1-5. Exposure response array of liver effects following oral exposure.
This document is a draft for review purposes only and does not constitute Agency policy.
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Toxicological ReviewofHexahydro-l,3,5-trinitro-l,3,5-triazine
Integration of Liver Effects
There is limited evidence from human studies and from studies in experimental animals
that RDX may affect the liver. The observation of short-term elevations of serum liver enzymes in
several human case reports of individuals who ingested unknown amounts of RDX suggests that
RDX might target the liver; however, serum liver enzymes were not elevated in a small cross-
sectional study of munition plant workers exposed to RDX. In experimental animals, dose-related
increases in liver weight were observed in some studies following subchronic oral exposure, but
liver weight changes were not consistent across sexes within a study or across different studies.
Changes in serum chemistry were not consistent across studies and the magnitude of change
relative to concurrent controls was not indicative of liver damage. Nonneoplastic histopathologic
lesions of the liver were also not consistently associated with RDX exposure. At this time, no
conclusions are drawn regarding liver effects as a human hazard of RDX exposure.
1.2.5. Carcinogenicity
The relationship between exposure to RDX and cancer in human populations has not been
investigated. The carcinogenicity of RDX has been examined in one oral chronic/carcinogenicity
bioassay in mice fLish etal.. 19841 and two bioassays in rats (Levine etal.. 1983b: Hart. 19761. The
2-year studies by Lish etal. f 19841 and Levine etal. Q983bl included comprehensive
histopathological examination of major organs, multiple dose groups and a control, and
>50 animals/dose group (plus additional interim sacrifice groups). In both studies, the maximum
tolerated dose was reached or exceeded in high-dose animals (based on decreased terminal body
weight in high-dose male and female mice of 5 and 19%, respectively, and decreased survival in
male and female rats by approximately 50 and 25%, respectively, compared to the control). The
earlier Hart (19761 study is largely limited by the lack of characterization of the test material and
the pathology examination in control and high-dose groups only. A temperature spike in the animal
rooms on study day 76 resulted in significant mortality across all dose groups and control animals;
however, there were still >80 rats/sex/group after the overheating incident and
>50 rats/sex/group at termination, and it seems unlikely that the mortality associated with the
temperature spike would have affected a tumor response in the rats. A summary of the evidence
for liver and lung tumors in experimental animals from these three bioassays is provided in
Tables 1-13 and 1-14.
Liver Tumors
An increased incidence of liver tumors was observed in one chronic mouse study (Lish etal..
19841 and one of two chronic rat studies (Levine etal.. 1983b). Incidences of hepatocellular tumors
are presented in Table 1-13 and discussed in further detail below.
The incidence of hepatocellular carcinomas and the combined incidence of hepatocellular
adenomas or carcinomas showed a statistically significant positive trend with RDX dose in female,
but not male, B6C3Fi mice as compared to concurrent controls in a 2-year dietary study fLish etal..
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19841. In female B6C3Fi mice, Lish etal. (1984) observed that the liver tumor incidence in the
concurrent female control mice was relatively low (1 /65), and significantly lower than the
incidence from historical controls (historical incidence data not provided by study authors). The
study authors also compared liver tumor incidence in RDX-exposed female mice to mean historical
control incidence for female mice of the same strain from National Toxicology Program (NTP)
studies conducted during the same time period (147/1,781 or 8%; range: 0-20%) fHaseman etal..
1985).11 The combined incidence of hepatocellular adenomas or carcinomas in female mice atRDX
doses >35 mg/kg-day (19% at both doses) was statistically significantly elevated when statistical
analysis was performed using NTP historical control data; limitations associated with comparisons
to historical control data originating from a different laboratory are acknowledged given cross-
study differences in diet, laboratory, pathological evaluation, and animal provider.
A Pathology Working Group (PWG) with substantial participation by NTP pathologists
reviewed the slides of female mouse liver lesions from the Lish etal. (1984) study (Parker etal..
2006: Parker. 2001). Some malignant tumors were downgraded to benign status and several
lesions initially characterized as adenomas were changed to non-neoplastic lesions based on more
recent diagnostic criteria used by the PWG (Harada etal.. 1999). There remained a statistically
significant positive trend in the combined incidence of hepatocellular adenomas or carcinomas,
consistent with the original findings of Lish etal. (1984). Because the PWG analysis reflects more
recent histopathological criteria for the grading of tumors, the incidence of hepatocellular
adenomas or carcinomas as reported by Parker etal. (2006) were considered the more reliable
measure of liver tumor response in female mice from the Lish etal. (1984) bioassay.
In male mice from the Lish etal. (1984) study, the incidences of hepatocellular carcinomas
in treated groups were higher than in the control, and the combined incidences of hepatocellular
adenomas or carcinomas of male mice were higher in three of four treated groups than in the
control; however, there were no statistically significant trends in either case. The incidences of
liver carcinoma in control (21%) and treated groups of male mice (22-33%) were generally within
the range for the same mouse strain reported by NTP (8-32%) fHaseman et al.. 1985). Similarly,
the combined incidences of liver adenoma or carcinoma in control (32%) and treated groups
"Comparison of control incidences of hepatocellular adenomas or carcinomas between Lish et al. (1984) and
Haseman etal. f!9851 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. f19851 did not include
the lab contracted to perform the Lish et al. f 19841 study, and it is not clear if the diet used in the Lish et al.
(19841 study was included in the diets reported in the NTP studies. Further, the NTP studies included three
different suppliers of mice; one supplier was also used in the Lish et al. (19841 study. EPA Guidelines for
Carcinogenic Risk Assessment (U.S. EPA. 2005al also note that, unless the tumor is rare, the standard for
determining statistical significance of tumor incidence is a comparison of dosed animals with the concurrent
controls.
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(27-48%) were within the range for the same mouse strain reported by NTP (14-58%) (Haseman
etal.. 1985).12 The PWG did not re-analyze liver tumor slides from male mice.
A statistically significant positive trend with dose was observed in the incidence of
hepatocellular carcinomas in male, but not female, F344 rats exposed to RDX in the diet for 2 years
fLevine etal.. 1983bl In the Levine etal. Q983bl study, there were only a few tumors observed in
the exposed groups of male rats (0/55, 0/52, 2/55, 2/31) relative to the control (1/55), and
inferences made from such a sparse response are uncertain. Because liver tumors are rare tumors
in the rat13, some perspective is obtained by considering historical control data. In a paper
published concurrently with the Levine etal. (1983b) study, NTP reported an incidence of liver
carcinomas in untreated control male F344 rats of 0.7% (12/1,719; range: 0-2%) fHaseman etal..
1985). In Levine etal. (1983b). the incidence of liver carcinomas in control male rats (1 /55 or
1.8%) was at the upper end of this NTP range, and the incidence in RDX-treated male F344 rats in
the highest two dose groups (3.6 and 6.4%) exceeded the NTP historical control range. Using
incidence data from NTP historical controls, the trend for carcinoma in the RDX-treated F344 rats
was statistically significant (p-value = 0.003; one-sided exact Cochrane-Armitage trend test). It
should be noted that although the NTP historical controls fHaseman et al.. 1985) are comparable
with Levine etal. Q983bl in terms of the time period, they may not be directly comparable in terms
of diet, laboratory, pathological evaluation, and animal provider. However, other historical control
datasets from male F344 rats, both recent and of the time period of the Levine study, indicate
similar low incidences of liver carcinomas (0.36%, (NTP. 2009): 0.31%, fMaita et al.. 1987)). In the
Levine etal. (1983b) study, the mortality in the highest dose group is substantially higher than in
the other dose groups during the second year leading to uncertainty in the true cancer incidence in
the high dose group. It was not possible to estimate mortality-adjusted incidences because no time-
to-death information was available.
Nonmalignant liver tumors (neoplastic nodules) in F344 male rats in NTP historical
controls were reported more frequently than carcinomas, with an average incidence of 3.5%
(61/1,719; range: 0-12%) fHaseman etal.. 1985): Levine etal. (1983b) reported an incidence of
neoplastic nodules of 7.3% in their control male rats, consistent with the NTP historical control
data, and a decline in incidence with increasing RDX exposure. The combined incidence of liver
neoplastic nodules or carcinomas did not show a significant trend with dose.
In a second 2-year dietary study in the rat study using a different strain (Sprague-Dawley),
the combined incidence of hepatocellular adenomas or carcinomas was not increased with dose in
rats of either sex at doses up to 10 mg/kg-day fHart. 19761. However, interpretation of results
12Ibid.
13NTP historical control data for hepatocellular carcinomas F344 rats as reported in Haseman etal. (1985):
12/1,719 (0.7%) in males; 3/1,766 (0.17%) in females. Historical control data for Charles River Sprague-
Dawley rats as reported in Chandra etal. fl9921: 6/1,340 (0.45%) in males; 1/1,329 (0.08%) in females.
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1 from this study is limited by the comparatively lower doses employed in the study, and the
2 recording of effects only at the control and high dose groups.
3 Table 1-13. Liver tumors observed in chronic animal bioassays
Reference and study design
Results3
Lish etal. (1984)
Doses
o
LO
O
35
175/100b
Mice, B6C3Fi, 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 urn
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
Hepatocellular adenomas (incidence)c
M
F
8/63 6/60 1/62*
(12.7) (10.0) (1.6)
1/65 1/62 6/64
(1.5) (1.6) (9.4)
7/59
(11.9)
6/64
(9.4)
7/27
(25.9)
3/31
(9.7)
Hepatocellular carcinomas (incidence)c
2 yrs
M
13/63 20/60 16/62
(20.6) (33.3) (25.8)
18/59
(30.5)
6/27
(22.2)
F
0/65 4/62 3/64
(0.0) (6.5) (4.7)
6/64
(9.4)
3/3 ld
(9.7)
Hepatocellular adenoma or carcinoma combined (incidence)c
M
20/63 26/60 17/62
(31.7) (43.3) (27.4)
25/59
(42.4)
13/27
(48.1)
F
1/65 5/62 9/64*
(1.5) (8.1) (14.1)
12/64*
(18.8)
6/31*d
(19.4)
PWG reanalvsis of liver lesion slides from female mice (Parker et al.,
2006; Parker, 2001).e
Doses
o
LO
O
35
175/100b
Hepatocellular adenomas (incidence)c
F
1/67 3/62 2/63
(1.5) (4.8) (3.2)
8/64
(12.5)
2/31
(6.5)
Hepatocellular carcinomas (incidence)c
F
0/67 1/62 3/63
(0.0) (1.6) (4.8)
2/64
(3.1)
2/31
(6.5)
Hepatocellular adenoma or carcinoma combined (incidence)c
F
1/67 4/62 5/63
(1.5) (6.5) (7.9)
10/64
(15.6)
4/3 ld
(12.9)
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Reference and study design
Results3
Hart (1976)
Doses
0 1.0
3.1
10
Rats, Sprague-Dawley, 100/sex/group
Purity and particle size not specified
0,1.0, 3.1, or 10 mg/kg-d
Neoplastic nodules (incidence)c
M
0/82
-
3/77
Diet
2 yrs
F
1/72
-
1/81
Hepatocellular carcinomas (incidence)c
M
1/82
-
1/77
F
1/72
-
1/8 lf
Neoplastic nodules or hepatocellular carcinomas combined
(incidence)c
M
1/82
-
4/77
F
2/72
-
2/81
Levine et al. (1983b)
Doses
LO
m
o
o
8.0
40
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
Neoplastic nodules (incidence)c
M
4/55 3/55 0/52
2/55
1/31
contaminant; 83-89% of particles <66 nm
(7.3) (5.5) (0.0)
(3.6)
(3.2)
0, 0.3,1.5, 8.0, or 40 mg/kg-d
F
3/53 1/55 1/54
0/55
4/48
Diet
2 yrs
(5.6) (1.8) (1.9)
(0.0)
(8.3)
Hepatocellular carcinomas (incidence)c
M
1/55 0/55 0/52
2/55
2/3 ld
(1.8) (0.0) (0.0)
(3.6)
(6.5)
F
0/53 1/55 0/54
0/55
0/48
(0.0) (1.8) (0.0)
(0.0)
(0.0)
Neoplastic nodules or hepatocellular carcinomas combined
(incidence)c
M
5/55 3/55 0/52
4/55
3/31
(9.1) (5.5) (0.0)
(7.3)
(9.7)
F
3/53 2/55 1/54
0/55
4/48
(5.6) (3.6) (1.9)
(0.0)
(8.3)
^Statistically significant difference compared to the control group (p < 0.05), identified by the authors.
aSelected percent incidences are provided in parentheses below the incidences to help illustrate patterns in the
responses.
bThe lower dose of 100 mg/kg-day was started in week 11, resulting in a duration-weighted average dose of
107 mg/kg-day.
cThe incidences reflect the animals surviving to month 12.
Statistically significant trend (p < 0.05) was identified using a one-sided Cochran-Armitage trend tests performed
by EPA.
eThe numbers of animals at risk (i.e., the denominators) in the control group (n = 67) and 7 mg/kg-day dose group
(n = 63) as reported in the PWG reanalysis (Parker et al., 2006; Parker, 2001) differed from the numbers reported
in the original study by Lish et al. (1984) (n = 65 and 64, respectively). Further investigation of these differences
by the U.S. Army (sponsor of the mouse bioassay and subsequent PWG reevaluation) was unable to resolve the
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discrepancy (email to Louis D'Amico, U.S. EPA, from Mark Johnson, U.S. Army Public Health Command, February
13, 2015).
fHart (1976) distinguishes the single high-dose carcinoma in the liver from a hepatocellular carcinoma; the
incidence of hepatocellular carcinomas in this dose group is shown as 0/81 (p. 119 of the publication).
Note: A dash ("-") indicates that the study authors did not measure or report a value for that dose group.
Lung Tumors
Lung tumors were observed in female and male B6C3Fi mice exposed to RDX in the diet for
2 years (Lish etal.. 19841 (see Table 1-14). Incidence of alveolar/bronchiolar carcinomas and the
combined incidence of alveolar/bronchiolar adenomas or carcinomas showed a statistically
significant positive trend (one-sided p-values of 0.016 and 0.009, respectively, for the Cochran-
Armitage trend test) in female mice. Incidence of alveolar/bronchiolar carcinomas in male mice
showed a statistically significant positive trend (p-value = 0.015; one-sided Cochran-Armitage trend
test). However, the combined incidence of adenomas and carcinomas was not elevated in male
mice. In such a case, NTP policy recommends analyzing the tumors both separately and in
combination (McConnell etal.. 1986). This recommendation arose out of concern that combining
benign and malignant neoplasms can result in a false negative if the chemical shows a statistically
significant increase in malignant tumors without an increase in the combined incidence. In an
addendum to the study report that included results of additional examination and sectioning of
lung specimens from the mid-dose groups in the mouse study, Lish etal. (1984) noted an increase
in the combined incidences of primary pulmonary neoplasms in males of all dose groups and in
females in the 7.0, 35, and 175/100 mg/kg-day dose groups, but regarded these neoplasms as
random and not biologically significant (rationale for this conclusion not provided).
Bioassays in rats provide no evidence of an association between RDX exposure and
induction of lung tumors. The incidence of alveolar/bronchiolar adenomas or carcinomas was not
increased in either sex of Sprague-Dawley rats exposed chronically to RDX at doses up to
10 mg/kg-day (Hart. 1976) or in F344 rats of either sex exposed chronically to RDX at doses up to
40 mg/kg-day (Levine etal.. 1983b). Alveolar/bronchiolar carcinomas are rare tumors in both
species of rats, male or female (Chandra etal.. 1992: Haseman etal.. 1985).
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1 Table 1-14. Lung tumors observed in chronic animal bioassays
Reference and study design
Results3
Lish etal. (1984)
Doses
o
LO
O
35
175/100b
Mice, B6C3Fi, 85/sex/group; interim
sacrifices (10/sex/group) at 6 and 12 mo
89.2-98.7% pure, with 3-10% HMX as
Alveolar/bronchiolar adenomas (incidence)c
M
6/63 5/60 5/62
7/59
1/27
contaminant; 83-89% of particles
(9.5) (8.3) (8.1)
(11.9)
(3.7)
<66 nm
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)
F
4/65 2/62 5/64
(6.2) (3.2) (7.8)
9/64
(14.1)
3/31
(9.7)
Alveolar/bronchiolar carcinomas (incidence)c
Diet
M
3/63 6/60 3/62
7/59
5/27d
2 yrs
(4.8) (10.0) (4.8)
(11.9)
(18.5)
F
3/65 1/62 3/64
3/64
4/3 ld
(4.6) (1.6) (4.7)
(4.7)
(12.9)
Alveolar/bronchiolar adenoma or carcinoma combined (incidencef
M
9/63 11/60 8/62
14/59
6/27
(14.3) (18.3) (12.9)
(23.7)
(22.2)
F
7/65 3/62 8/64
12/64
7/3 ld
(10.8) (4.8) (12.5)
(18.8)
(22.6)
Hart (1976)
Doses
1
rn
o
O
10
Rats, Sprague-Dawley, 100/sex/group
Purity and particle size not specified
0,1.0, 3.1, or 10 mg/kg-d
Alveolar/bronchiolar adenoma (incidence)
M
2/83
-
1/77
Diet
2 yrs
F
0/73
-
0/82
No alveolar/bronciolar carcinomas reported by study authors.
Levine et al. (1983b)
Doses
LO
m
o
o
8.0
40
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
Alveolar/bronchiolar adenomas (incidence)c
M
1/55 0/15 1/17
0/16
1/31
contaminant; 83-89% of particles
<66 nm
0, 0.3,1.5, 8.0, or 40 mg/kg-d
F
3/53 0/7 0/8
1/10
0/48
Alveolar/bronchiolar carcinomas (incidence)c
Diet
2 yrs
M
- - -
-
-
F
0/53 0/7 1/8
0/10
0/48
Alveolar/bronchiolar adenoma or carcinoma combined (incidencef
M
F
3/53 0/7 1/8
1/10
0/48
2
3 aSelected percent incidences are provided in parentheses below the incidences to help illustrate patterns in the
4 responses.
5 bThe lower dose of 100 mg/kg-day was started in week 11, resulting in a duration-weighted average dose of
6 107 mg/kg-day.
7 cThe incidences reflect the animals surviving to month 12.
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Statistically significant trend (p < 0.05) was identified using a one-sided Cochran-Armitage trend test performed by
EPA.
Note: A dash indicates that the study authors did not measure or report a value for that dose group.
Mechanistic Evidence
There are few mechanistic data to inform a MOA determination for either liver or lung
tumors induced by exposure to RDX.
The available in vitro and in vivo genotoxicity assay results are largely negative for parent
RDX or its oxidative metabolites (see Appendix C, Section C.3.2), supporting the hypothesis that
parent RDX or its oxidative metabolites do not interact directly with deoxyribonucleic acid (DNA).
In contrast, there are some positive genotoxicity results for the N-nitroso metabolites of RDX,
specifically hexahydro-l-nitroso-3,5-dinitro-l,3,5-triazine (MNX) and hexahydro-l,3,5-trinitroso-
1,3,5-triazine (TNX). MNX and TNX have been identified from minipigs; minipigs were chosen as
the animal model for investigation of RDX metabolism because the GI tract of pigs more closely
resembles that of humans fMusick etal.. 2010: Major etal.. 20071. MNX has tested positive in some
in vitro assays, including unscheduled DNA synthesis in primary rat hepatocytes and the mouse
lymphoma forward mutation assay fSnodgrass. 19841. although MNX tested negative in the only in
vivo test performed, a mouse dominant lethal mutation test fSnodgrass. 19841. MNX was not
mutagenic in Salmonella typhimurium (strains TA98, TA100, TA1535, TA1537, and TA1538), with
or without the addition of the S9 metabolic activating mixture (Pan etal.. 2007b: Snodgrass. 19841.
When S. typhimurium strains TA97a and TA102, strains sensitive to frame shift and oxidative DNA
damage, were used in conjunction with elevated concentrations of the metabolizing system (S9),
MNX and TNX were mutagenic. N-nitroso metabolites, including MNX and TNX, are generated
anaerobically and are likely a result of bacterial transformation of parent RDX in the GI tract to
various N-nitroso derivatives fPan etal.. 2007bl. Exposure to potentially mutagenic N-nitroso
metabolites of RDX generated in the GI tract of mice may occur in the liver (and subsequently in the
systemic circulation) via enterohepatic circulation. However, in pigs, the N-nitroso metabolites of
RDX have been identified only in trace amounts in urine compared to the major metabolites,
4-nitro-2,4-diazabutanal and4-nitro-2,4-diaza-butanamide (Major etal.. 20071. Thus, the
contribution of the N-nitroso metabolites to the overall carcinogenic potential of RDX is unclear.
Aberrant expression of miRNAs was observed in the brains and livers of female B6C3Fi
mice fed 5 mg RDX/kg in the diet for 28 days fZhang and Pan. 2009bl (dose of 0.75-1.5 mg/kg-day
estimated by Bannon etal. (2009bll. with several oncogenic miRNAs being upregulated, while
several tumor-suppressing miRNAs were downregulated. However, the pattern of induction was
not always consistent in the livers of RDX-treated mice (e.g., miR-92a was downregulated in liver
tissue samples when it is typically upregulated in hepatocellular carcinomas) (Sweeney etal..
2012b). miRNAs have been associated with several cancers fWiemer. 2007: Zhang etal.. 20071. but
the utility of miRNAs as predictive of carcinogenesis has not been demonstrated (Bannon etal..
2009b). Further, it is unknown whether or not aberrant expression of a specific miRNA (or suite of
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1 miRNAs) plays a role in the MOA of RDX carcinogenicity. Microarray analysis of gene expression in
2 male Sprague-Dawley rats after exposure to a single oral (capsule) dose of RDX revealed a general
3 upregulation in gene expression (predominantly genes involved in metabolism) in liver tissues
4 (Bannon et al.. 2009a): however, the relevance of this finding to the carcinogenicity of RDX is
5 unclear.
6 Sweeney etal. (2012b) hypothesized a set of MOAs for the liver tumors:
• Genotoxicity mediated by either: (1) RDX; (2) tissue-generated oxidative metabolites; or
(3) N-nitroso metabolites generated anaerobically in the GI tract. The key events in this
hypothesized MOA are: production of DNA damage, gene mutation, formation of neoplastic
lesions, and promotion/progression of tumors. The largely negative results for genotoxicity
led Sweeney etal. (2012b) to conclude that this MOA is not plausible for RDX or its
oxidative metabolites. Although there are some positive results for the N-nitroso
metabolites, the limited evidence to support systemic uptake and distribution of
metabolites to the liver led Sweeney etal. (2012b) to conclude that this MOA is not
sufficiently plausible.
7 • Cell proliferation. The key events in this hypothesized MOA are Gl-tract generation of
8 N-nitroso metabolites, absorption, distribution to the liver, cytoxocity (optional), and
9 enhanced cell proliferation, leading to preneoplastic foci that progress to hepatocellular
10 adenomas and carcinomas. Sweeney etal. (2012b) cited evidence of increased liver weights
11 in mice as consistent with cell proliferation, but noted that increased liver weights were
12 also observed in rats without proceeding to liver tumors. They considered this MOA
13 "plausible, but not particularly well supported."
14 In addition, there are other results that do not support a metabolite-based proliferative
15 response as the MOA for carcinogenesis.
16 • The absence of significant liver histopathology in mice after subchronic or chronic exposure
17 to RDX at doses that induced liver tumors (Lish etal.. 1984: Cholakis etal.. 19801 suggests
18 that cellular toxicity is not a precursor to these tumors.
19 • As discussed in Section 1.2.4, changes in liver weight showed no consistent pattern across
20 studies or sexes, and did not correlate with tumor response.
21 • No studies were available that directly measured RDX-induced cell proliferation rates.
22 • No information was available to rule out non-precancerous causes of liver weight increase.
23 In summary, the available evidence indicates that RDX is likely not mutagenic (see
24 Appendix C, Section C.3.2), although anaerobically-derived N-nitroso metabolites have
25 demonstrated some genotoxic potential. While these metabolites have been measured in the
26 mouse (Pan etal.. 2007b) and minipig fMusick etal.. 2010: Major etal.. 2007). they have not been
27 identified in humans, and may not be the predominant metabolites of RDX. A MOA involving a
28 proliferative response generated by tissue-derived oxidative metabolites of RDX has been
29 proposed, but is not supported by the available data. In light of limited information on precursor
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events leading to the observed liver and lung tumor response in RDX-exposed rodents and lack of
toxicokinetic information on RDX metabolites, neither a cell proliferative MOA nor a mutagenic N-
nitroso metabolite MOA is supported. Thus, the MOA leading to the increased incidence of liver and
lungs tumors is not known.
1.2.6. Other Noncancer Effects
There are isolated reports of RDX inducing systemic effects in several organs/systems,
including the eyes and the musculoskeletal, cardiovascular, immune, and GI systems. However,
there is less evidence for these effects compared to organ systems described earlier in Section 1.2.
Generally, evidence for toxicological effects in these organ systems was limited to human case
reports, lacked reproduction or were not observed in other studies of similar duration in the same
species, or lacked consistent, dose-related patterns of increasing or decreasing effect. A longer
discussion of the evidence for each of the other noncancer effects noted above is provided in
Appendix C.3.2. At this time, no conclusions are drawn regarding the other noncancer effects as
human hazards of RDX exposure.
1.3. INTEGRATION AND EVALUATION
1.3.1. Effects Other Than Cancer
The majority of evidence for the health effects of RDX comes from oral toxicity studies in
animals. The three epidemiology studies that document possible inhalation exposure are limited by
various study design features, including inability to distinguish exposure to TNT (associated with
liver and hematological system toxicity), inability to adequately characterize exposure levels, small
sample sizes, and inadequate reporting. The single animal inhalation study identified in the
literature search had deficiencies (e.g., lack of a control and incomplete exposure information) that
precluded its inclusion in this assessment (see literature search section).
The strongest evidence for hazards following exposure to RDX is for nervous system effects.
Toxicity studies in multiple animal species involving chronic, subchronic, and gestational exposures
provide consistent evidence of nervous system effects following oral exposure. Effects included
dose-related increases in seizures and convulsions, as well as observations of tremors,
hyperirritability, hyper-reactivity, and other behavioral changes fCrouse etal.. 2006: Angerhofer et
al.. 1986: Levine etal.. 1983b: Levine etal.. 1981a. b; Cholakis etal.. 1980: von Oettingen etal..
19491.
Human studies provide supporting evidence for RDX as a neurotoxicant and that the
nervous system effects observed in experimental animals are plausible in, and relevant to, humans.
A cross-sectional study described memory impairment and visual-spatial decrements in RDX-
exposed workers (Ma and Li. 1993). although confidence in these findings is relatively low because
of issues with design and reporting. Several case reports provide additional evidence of
associations between exposure to RDX (via ingestion, inhalation, and possibly dermal exposure)
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and seizures and convulsions fKasuske etal.. 2009: Kiiciikardali etal.. 2003: Testud etal.. 1996a:
Testud etal.. 1996b: Woody et al.. 1986 and others, see Appendix C.2). Other nervous system
effects identified in human case reports include dizziness, headache, confusion, and
hyperirritability.
Additional support for an association between RDX exposure and nervous system effects
comes from consistent evidence of neurotoxicity across taxa, including several species of wildlife
fOuinn etal.. 2013: Garcia-Revero etal.. 2011: McFarland et al.. 2009: Gogal etal.. 20031. 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 fWilliams etal.. 2011: Bannon etal.. 2009al. Additionally, the affinity of RDX for the
picrotoxin convulsant site of the GABAa channel suggests that the resulting disinhibition could lead
to the onset of seizures fWilliams etal.. 20111.
Induction of convulsions and seizures appears to be more strongly correlated with dose
than with duration of exposure. However, there is some mechanistic information to suggest that
repeated exposure to a chemical binding to the receptor convulsant site of GABAa may promote a
state of increased neuronal activity that could increase the likelihood of subsequent neurological
effects fGerkin etal.. 20101. It is unclear if nervous system effects progressed in severity (e.g., from
behavioral change to seizures and convulsions) with increasing dose, as many of the studies that
reported more subtle neurobehavioral changes did not provide detailed dose-response
information, and the majority of studies were not designed to capture this information.
The nervous system effects following oral exposure to RDX were observed in humans
acutely exposed to RDX and in multiple experimental animal studies in rats, mice, monkeys, and
dogs following exposures ranging from 10 days to 2 years in duration. Across the database,
behavioral manifestations of seizure activity were the most consistently observed nervous system
effect associated with RDX exposure. This most commonly included evidence of increased
convulsions, as well as other related effects such as tremors, shaking, hyperactivity, or nervousness,
which were generally observed at doses that were the same as or higher than doses that induced
convulsions. Nervous system effects are a human hazard of RDX exposure and are carried forward
for consideration for dose-response analysis. Convulsions, considered a severe adverse effect, were
selected as a consistent and sensitive endpoint representative of nervous system effects.
Evidence for kidney and other urogenital toxicity is more limited than evidence for
neurotoxicity. Histopathological changes in the urogenital system (suppurative prostatitis,
medullary papillary necrosis, suppurative pyelitis, uremic mineralization, and luminal distention
and cystitis of the urinary bladder) were reported in male rats exposed to RDX in the diet for
2 years (Levine etal.. 1983b). Similar histopathological changes of the urogenital system were not
observed in mice, and no other rat studies of similar duration that examined the prostate were
available. As discussed earlier, among the lesions identified in the rat, the incidence of suppurative
prostatitis is considered a marker for RDX-related urogenital effects as it is plausibly associated
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with a progression of effects resulting from other kidney or bladder lesions. The plausibility of a
MOA that shares a common molecular initiating event (binding to the GABAa receptor convulsant-
site) with the neurotoxic effects of RDX provides some support for an association between RDX
exposure and kidney and other urogenital effects. Kidney and other urogenital system effects are a
potential human hazard of RDX exposure and were carried forward for consideration for dose-
response analysis. Prostatitis, considered a marker for the kidney and urogenital effects, was
selected as a sensitive endpoint representative of the urogenital system effects.
There is some evidence for male reproductive toxicity that comes from the finding of
testicular degeneration in male B6C3Fi mice chronically exposed to RDX in the diet fLish etal..
19841 in the only mouse study conducted of that duration (24 months). The effect was noted by the
study authors at both the penultimate and highest dose tested in the study. However, studies in
different rat strains did not consistently observe testicular effects. Although the available data are
limited, given the dose-related findings of mouse testicular degeneration, there is suggestive
evidence of male reproductive effects associated with RDX exposure; these effects were carried
forward for consideration for dose-response analysis. Testicular degeneration, the only endpoint
observed, was selected as the endpoint representative of male reproductive effects.
Evidence for developmental toxicity and liver toxicity was more limited than that for the
endpoints discussed above. In animal studies, developmental effects, including offspring survival,
growth, and morphological development, were observed only at doses associated with maternal
mortality (Angerhofer etal.. 1986: Cholakis etal.. 19801. Evidence for potential hepatic effects
comes from observations of increases (generally dose-related) in liver weight in some subchronic
oral animal studies (Lish etal.. 1984: Levine etal.. 1983b: Levine etal.. 1981a. b; Cholakis etal..
1980: Hart. 19761. However, these elevations in liver weight were not consistently observed across
studies nor were they accompanied by RDX-related histopathological changes in the liver or
increases in serum liver enzymes. In addition, the interpretation of liver weight changes in the
mouse bioassay by Lish etal. f 19841 is complicated by the relatively high incidence of liver tumors
in this study. At this time, no conclusions are drawn regarding developmental and liver toxicity as
human hazards of RDX exposure; these effects were not considered further for dose-response
analysis and derivation of reference values.
As discussed in Section 1.2, mortality is not addressed in this assessment as a hazard by
itself, but rather in the context of nervous and urogenital system hazards. Histopathological
changes in the urogenital system observed in male rats exposed to 40 mg/kg-day in the diet for 2
years were considered the principal cause of treatment-related morbidity and mortality f Levine et
al.. 1983b). However, the incidence of suppurative prostatitis, considered a sensitive marker of the
urogenital effects, was increased at doses of >1.5 mg/kg-day. Therefore, the mortality
characterized as secondary to renal effects in Levine etal. (1983b) is a less sensitive endpoint (by
more than 10-fold) than the effect that is the selected as the basis dose-response analysis (i.e.,
suppurative prostatitis).
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In a number of the animal studies reporting nervous system effects, unscheduled deaths
occurred at RDX doses as low as those that induced nervous system effects (Grouse etal.. 2006:
Angerhofer etal.. 1986: Levine etal.. 1983b: Levine etal.. 1981a: Cholakis etal.. 1980: von
Oettingen etal.. 19491. In a 90-day study that recorded nervous system effects and survival more
thoroughly than earlier studies, Grouse etal. (2006) reported that nearly all pre-term deaths were
preceded by neurotoxic signs such as tremors and convulsions. Convulsions did not, however,
necessarily lead to early mortality; of the animals observed to have convulsed in the Grouse etal.
(2006) study, approximately 75% survived to the end of the 90-day study. Most of the earlier
studies provide a limited understanding of the association between mortality and nervous system
effects because the frequency of clinical observations was likely insufficient to observe convulsions
prior to death. In humans, mortality has not been reported in case reports involving workers with
symptoms of neurotoxicity exposed to RDX during manufacture or in individuals exposed acutely as
a result of accidental or intentional ingestion; however, survival has not been specifically evaluated
in studies of worker populations exposed chronically to RDX.
Regarding mortality, the preference, in general, is not to use a frank health effect as severe
as mortality as the basis for a reference value. As noted in U.S. EPA (2002). a chemical may cause a
variety of effects ranging from severe—such as death—to more subtle biochemical, physiological,
or pathological changes; primary attention in assessing health risk should be given to those effects
in the lower exposure range and/or the effects most biologically appropriate for a human health
risk assessment Where mortality occurs as a consequence of a chemical's effects on a specific
organ/system (e.g., in the case of RDX, evidence suggests some relationship between mortality and
effects on the nervous or kidney/urogenital systems), the preference would be to develop a
quantitative assessment based on the initial hazard and not on death. Because unscheduled deaths
were observed with some consistency across studies and, in some studies, at doses as low as those
associated with convulsions, two additional analyses of mortality data are presented in Chapter 2.
In the first analysis, BMDs derived using mortality data sets are compared to the BMD used to
derive the RfC (Section 2.1.6). In addition, the relationship between convulsions and mortality is
not clear and raises concerns for the potential underreporting of convulsions (see Section 1.2.1).
An analysis, described in Section 2.1.7, addresses the possibility that the analyses of convulsions
brought forward for dose-response analysis resulted in an underestimate of the toxicity for RDX.
1.3.2. Carcinogenicity
As presented in Section 1.2.5, dietary administration of RDX induced dose-related increases
in the incidence of hepatocellular adenomas or carcinomas in male and female B6C3Fi mice mice
(Parker etal.. 2006: Lish etal.. 1984). In the same study, RDX also induced dose-related increases
in the incidence of alveolar/bronchiolar adenomas or carcinomas in both sexes. Some of these
trends in liver and lung were statistically significant. In Fischer 344 rats, dietary administration of
RDX yielded a statistically significant trend in the incidence of hepatocellular carcinomas in males,
but not in females fLevine etal.. 1983b). A 2-year dietary study in Sprague-Dawley rats was
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negative in both sexes Hart (1976). although doses in this study were somewhat lower (no
carcinomas at doses up to 10 mg/kg-d in Hart (1976). versus hepatocellular carcinomas at 8 and
40 mg/kg-d in the Levine etal. Q983bl study. The human studies are not informative.
This evidence leads to consideration of two hazard descriptors under the EPA's cancer
guidelines fU.S. EPA. 2005al. The descriptor likely to be carcinogenic to humans is appropriate
when the evidence is "adequate to demonstrate carcinogenic potential to humans" but does not
support the descriptor carcinogenic to humans. One example from the cancer guidelines is "an
agent that has tested positive in animal experiments in more than one species, sex, strain, site, or
exposure route, with or without evidence of carcinogenicity in humans." RDX matches the
conditions of this example, having induced dose-related increases in tumors in two species (mouse
and rat), in both sexes, and at two sites (liver and lung).
Alternatively, the descriptor suggestive evidence of carcinogenic potential is appropriate
when the evidence raises "a concern for potential carcinogenic effects in humans" but is not
sufficient for a stronger conclusion. The hepatocellular carcinoma result in male F344 rats is based
on a small number of tumors (two each at 8 and 40 mg/kg-day, versus one in controls), and RDX
did not increase the incidence of carcinomas at any other site in F344 or Sprague-Dawley rats of
either sex.
As noted in the EPA's cancer guidelines (U.S. EPA. 2005a). choosing a hazard descriptor
cannot be reduced to a formula, as descriptors may be applicable to a variety of potential data sets
and represent points along a continuum of evidence. In the case of RDX, there are plausible
scientific arguments for more than one hazard descriptor. Overall, the considerations discussed
above, interpreted in light of the cancer guidelines, lead to the conclusion that there is suggestive
evidence of carcinogenic potential for RDX. Although the evidence includes dose-related tumor
increases in two species, two sexes, and two sites, the evidence of carcinogenicity outside the
B6C3Fi mouse is not robust, and this factor was decisive in choosing a hazard descriptor. Within
the spectrum of results covered by the descriptor suggestive evidence, the evidence for RDX is
strong. There are well-conducted studies that tested large numbers of animals at multiple dose
levels, making the cancer response suitable for dose-response analysis (Section 2).
The descriptor suggestive evidence of carcinogenic potential applies to all routes of human
exposure. Dietary administration of RDX to mice and rats induced tumors of the liver or lung, sites
beyond the point of initial contact, and human case reports have demonstrated absorption and
distribution of inhaled RDX into the systemic circulation. Under the cancer guidelines, this
information provides sufficient basis to apply the cancer descriptor developed from oral studies to
other exposure routes.
1.3.3. Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes
Susceptibility refers to factors such as lifestage, genetics, sex, and health status that may
predispose a group of individuals to greater response to an exposure. This greater response could
be achieved either through differences in exposure to the chemical underlying RDX toxicokinetics
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or differences in RDX toxicodynamics between susceptible and other populations. Little
information is available on populations that may be especially vulnerable to the toxic effects of RDX.
Lifestage, and in particular childhood, susceptibility has not been observed in human or
animal studies of RDX toxicity. Transfer of RDX from dam to the fetus during gestation has been
reported, and the presence of RDX in the milk of dams administered 6 mg/kg-day by gavage has
been documented fHess-Ruth etal.. 20071: however, reproductive and developmental toxicity
studies generally did not identify effects in offspring at doses below those that also caused severe
maternal toxicity (Angerhofer et al.. 1986: Cholakis etal.. 19801. Thus, the existing toxicity
literature does not provide evidence of early lifestage susceptibility to RDX.
Limited data suggest that male laboratory animals may be more susceptible to noncancer
toxicity associated with RDX exposure. In general, male animals were more sensitive to RDX
neurotoxicity than females (i.e., more convulsions; more hyperactive; greater brain weight
changes). Urogenital effects have been observed in males at lower doses than in females fLevine et
al.. 1983b: Levine etal.. 1981a. b; Cholakis etal.. 19801. suggesting a possible sex-based difference
in susceptibility to RDX toxicity.
Data on the incidence of convulsions and mortality from gavage studies of RDX in the rat
provide some indication that pregnant animals may be a susceptible population. In the
developmental toxicity study by Cholakis etal. (19801. deaths were observed in pregnant F344 rats
only at a dose of 20 mg/kg-day, but convulsions were reported in a single rat at 2 mg/kg-day. In a
range-finding developmental toxicity study (Angerhofer etal.. 19861. mortality and convulsions
were reported in pregnant_Sprague-Dawley rats at a dose of >40 mg/kg-day, but not at <20 mg/kg-
day, although the relatively small group sizes in this study should be noted. In the main study by
these investigators, convulsions were reported in pregnant rats only at 20 mg/kg-day, but one
death (in dose groups of 40 rats) was reported at both 2 and 6 mg/kg-day f Angerhofer et al.. 19861.
In comparison, increased mortality and convulsions were reported at >8 mg/kg-day in a 90-day
gavage study in F344 rats fCrouse etal.. 20061. The instances of one convulsion and two deaths in
pregnant rats in the Cholakis etal. (19801 and Angerhofer et al. (19861 studies at doses of 2 or 6
mg/kg-day raise the possibility thatpregnants animals may be more susceptible to the effects of
RDX; however, direct comparison between the available gavage studies in pregnants and
nonpregnant rats is uncertain because of differences in study design, including numbers of animals
tested per group, test material characteristics, and rat strain. Overall, the available information is
not considered sufficient to conclude that pregnant animals are a susceptible population.
There is limited evidence that CYP450 or similar enzymes are involved in the metabolism of
RDX (Bhushan et al.. 20031. indicating a potential for genetic polymorphisms in these metabolic
enzymes to affect susceptibility to RDX. This susceptibility may also be influenced by differential
expression of these enzymes during development Individuals with epilepsy or other seizure
syndromes, and in particular those that have their basis in genetic mutation to GABAa receptors,
may represent another group that may be susceptible to RDX exposure. However, there is currently
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no information to support predictions of how genetic polymorphisms or the presence of seizure
syndromes may affect susceptibility to RDX exposure.
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2. DOSE-RESPONSE ANALYSIS
2.1. ORAL REFERENCE DOSE FOR EFFECTS OTHER THAN CANCER
The oral reference dose (RfD, expressed in units of mg/kg-day) is defined as an estimate
(with uncertainty spanning perhaps an order of magnitude) of a daily exposure to the human
population (including sensitive subgroups) that is likely to be without an appreciable risk of
deleterious effects during a lifetime. It can be derived from a no-observed-adverse-effect level
(NOAEL), lowest-observed-adverse-effect level (LOAEL), or the 95% lower bound on the
benchmark dose (BMDL), with uncertainty factors (UFs) generally applied to reflect limitations of
the data used.
2.1.1. Identification of Studies for Dose-Response Analysis of Selected Effects
As discussed in Section 1.3.1, based on findings from oral studies in experimental animals,
nervous system effects are a human hazard of hexahydro-l,3,5-trinitro-l,3,5-triazine (RDX)
exposure, and kidney and other urogenital system effects are a potential human hazard of RDX
exposure. There is suggestive evidence of male reproductive effects associated with RDX exposure.
Although animal mortality has been reported in a number of the toxicology studies conducted for
RDX, it was not considered a hazard by itself or as the basis for the derivation of a reference
value. Rather, the mortality evidence was evaluated in the context of that system-specific hazard
(see Sections 2.1.6 and 2.1.7 for further discussion).
The effects selected to best represent each of the hazards (see discussion in Section 1.3.1)
are noted below. In order to identify the stronger studies for dose-response analysis, several
attributes of the studies reporting the endpoints selected for each hazard were reviewed (i.e., study
size and design, relevance of the exposure paradigm, and measurement of the endpoints of
interest). In considering the study size and design, preference was given to studies using designs
reasonably expected to have power to detect responses of suitable magnitude. Exposure paradigms
including a route of human environmental exposure (i.e., oral and inhalation) are preferred. When
developing a chronic reference value, chronic or subchronic studies are preferred over studies of
acute exposure durations. Studies with a broad exposure range and multiple exposure levels are
preferred to the extent that they can provide information about the shape of the exposure-response
relationship. Additionally, with respect to measurement of the endpoint, studies that can reliably
distinguish the presence or absence (or degree of severity) of the effect are preferred.
Human studies are generally preferred over animal studies as the basis for a reference value
when quantitative measures of exposure are reported, and the reported effects are determined to
be associated with exposure. The available epidemiological studies of worker populations exposed
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to RDX examined the relationship between certain health endpoints and inhalation exposure;
however, no epidemiological studies of ingested RDX are available. Multiple case reports support
the identification of hazards associated with RDX exposure but are inadequate for dose-response
analysis because they do not yield incidence estimates, exposure durations are short, and
quantitative exposure information is lacking. Therefore, human studies could not be used for oral
dose-response analysis or to serve as the basis for the RfD. In the absence of human data, the
animal studies were considered for dose-response analysis.
Experimental animal studies considered for each health effect were evaluated using general
study quality considerations discussed in Section 6 of the Preamble and in the literature search
section, and the attributes described above. The rationales for selecting the strongest studies that
reported the health effects are summarized below.
Nervous System Effects
Convulsions were selected for dose-response analysis as a consistent and sensitive endpoint
of nervous system effects (see Section 1.3.1 for discussion). This endpoint was reported in seven
studies (Grouse etal.. 2006: Lish etal.. 1984: Levine etal.. 1983b: Levine etal.. 1981a: Cholakis et
al.. 1980: Martin and Hart. 1974: von Oettingen et al.. 19491. Table 2-1 provides an overview of the
information considered in the studies reporting nervous system effects (i.e., convulsions) evaluated
for dose-response analysis.
Table 2-1. Information considered for evaluation of studies that examined
convulsions
Study
reference
Study design and size
Relevance of exposure paradigm
Measurement
of endpoint
Design
Number of
animals
Route
Duration
Number of
dose
groups1
Levels
(mg/kg-d)
Purity
(%)
Incidence
data reported
Crouse et al.
(2006)
Toxicity study
10 rats/sex/
group
Gavage
13-wk
5
4-15
99.99
Yes
Cholakis et al.
(1980)
Developmental
study
24-25 female
rats/group
Gavage
14-d
3
0.2-20
89
Yes
Martin and
Hart (1974)
Toxicity study
3 monkeys/
sex/group
Gavage
13-wk
3
0.1-10
Not
specifie
d
Yes
Levine et al.
(1983b)
Toxicity and
carcinogenicity
bioassay
75 rats/sex/
group
Diet
2-yr
4
0.3-40
89-99
No
Lish et al.
(1984)
Toxicity and
carcinogenicity
bioassay
85 mice/sex/
group
Diet
2-yr
4
1.5-175
89-99
No
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Study
reference
Study design and size
Relevance of exposure paradigm
Measurement
of endpoint
Design
Number of
animals
Route
Duration
Number of
dose
groups1
Levels
(mg/kg-d)
Purity
(%)
Incidence
data reported
Levine et al.
Toxicity study
10 rats/sex/
group
diet
13-wk
5
10-600
85
No
(1981a)
von Oettingen
Toxicity study
20 rats/group
diet
13-wk
3
15-50
90-97
No
etal. (1949)
Excluding the control group.
Incidence of convulsions was reported in three studies of RDX—all involving gavage
administration: Grouse etal. (20061 Cholakis etal. (19801 (developmental toxicity study), and
Martin and Hart (19741 Qualitative findings of nervous system effects were reported in other
chronic and subchronic studies—all involving dietary administration: Lish etal. (1984). Levine et
al. (1983b). Levine etal. (1981a). and von Oettingen etal. (1949). Incidence data on neurotoxic
effects of RDX were not collected in any of the dietary studies. For example, Levine etal. (1983b)
reported only that convulsions and other nervous system effects were noted in rats exposed to RDX
for 2 years at the highest dose (40 mg/kg-day) tested. The studies that included incidence data (i.e.,
the gavage studies) were preferred over those studies only reporting qualitative results (i.e., the
dietary studies).
The three gavage studies reporting incidence data were further considered. Grouse et al.
(2006) reported a dose-related increase in convulsions and tremors in both male and female F344
rats following a 90-day oral (gavage) exposure to RDX. This study used a test material of high
purity and six dose groups (including the control) that provided good resolution of the dose-
response curve. Cholakis etal. (1980) reported a dose-related increase in convulsions in a
developmental toxicity study in F344 rats, following a 14-day exposure to RDX on gestational days
(GDs) 6-19. Although this study was designed as a standard developmental toxicity study (i.e., not
specifically to examine nervous system effects), it reported information on the identity of the test
material and used three dose groups that adequately characterized the dose-response curve.
Further, this study provided evidence of nervous system effects at a relatively low dose. The study
in monkeys by Martin and Hart (1974) provides supporting evidence of nervous system effects
(trembling, shaking, ataxia, hyperactive reflexes, and convulsions); however, this study was not
selected for dose-response analysis because of small group sizes (n = 3/sex) and uncertainty in
measures of exposures (e.g., purity of the test material was not specified, and reported emesis in
some animals likely influenced the delivered dose).
Although the gavage studies reporting incidence data were preferred over four dietary
studies (Lish etal.. 1984: Levine etal.. 1983b: Levine etal.. 1981a: von Oettingen et al.. 1949) that
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did not provide incidence data, it is important to note that the reported neurotoxic effects in the
dietary studies were observed at dose levels higher than the doses at which effects were observed
in the gavage studies fCrouse etal.. 2006: Cholakis etal.. 1980: Martin and Hart. 19741. Given this
potential difference based on dosing method, the dietary studies were also considered for
quantitative analysis, despite the lack of incidence data, to evaluate the influence of oral dosing
method on candidate reference values. In the 2-year study by Levine etal. f!983bl. a LOAEL for
nervous system effects (convulsions, tremors, and hyper-irritability) of 40 mg/kg-day and a NOAEL
of 8 mg/kg-day were identified. Other studies identified higher effect levels (i.e., 100 mg/kg-day in
the 2-year mouse study by Lish etal. (1984) and 50 mg/kg-day in the 3-month rat study by von
Oettingen et al. Q9491I and, with the exception of Lish etal. Q9841. used shorter exposure
durations. The unusual dosing regimen in the Cholakis etal. (1980) 13-week mouse study
precluded identification of a NOAEL and LOAEL, and the single-dose design of the 6-week dog study
by von Oettingen et al. (1949) did not allow identification of a NOAEL. As discussed in Section 1.2.1
and T able 1-3, the technical report of the 13-week study by Levine etal. (1981a) inconsistently
identified the dose level at which convulsions occurred; therefore, a reliable NOAEL and LOAEL
from this study could not be identified.
Therefore, two gavage studies, Grouse etal. f20061 and Cholakis etal. f!980I and one
dietary study, Levine etal. (1983b). were selected for dose-response analysis.
Kidney and Other Urogenital System Effects
Suppurative prostatitis was selected for dose-response analysis. It is considered to be a
sensitive marker for the broader range of urogenital effects observed in F344 male rats in a 2-year
study by Levine etal. (1983b). The Levine etal. (1983b) study: (1) included a histopathological
examination of the kidney and other urogenital system tissues at 6-, 12-, and 24-month time points;
(2) included four dose groups and a control group, and adequate numbers of animals per dose
group (75/sex/group, with interim sacrifice groups of 10/sex/group at 6 and 12 months); and
(3) reported individual animal data. This study, the only one to identify suppurative prostatisis,
was selected for dose-response analysis.
Male Reproductive Toxicity
Testicular degeneration was selected for dose-response analysis. Lish etal. f 19841
observed a dose-related increase in the incidence of testicular degeneration in mice following
chronic administration of RDX in the diet This 2-year study: (1) included histopathological
examination of male reproductive organs; (2) included four dose groups and a control group, and
adequate numbers of animals per dose group (85/sex/group, with interim sacrifice groups of
10/sex/group at 6 and 12 months); and (3) reported individual animal data. This study, the only
one to identify testicular degeneration, was selected for dose-response analysis.
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2.1.2. Methods of Analysis
No biologically based dose-response models are available for RDX. In this situation, the U.S.
Environmental Protection Agency (EPA) evaluates a range of dose-response models thought to be
consistent with underlying biological processes to determine how best to empirically model the
dose-response relationship in the range of the observed data. Consistent with this approach, EPA
evaluated dose-response information with the models available in EPA's Benchmark Dose Software
(BMDS, versions 2.4 and 2.5). EPA estimated the benchmark dose (BMD) and BMDL using a
benchmark response (BMR) selected for each effect A conceptual model of the analysis approach
used for RDX is provided in Figure 2-1. In this assessment, points of departure (PODs) are
identified through BMD modeling (preferred) or identification of a NOAEL, and followed by animal-
to-human extrapolation through the use of physiologically based pharmacokinetic (PBPK) models
or the application of a dosimetric adjustment factor, depending on the data available.
BMD modeling
(or NOAEL)
Animal PBPK
model
Dosimetric
adjustment
factor (DAF)
Human PBPK
model
POD
Animal internal dose
Animal administered
(external) dose
Human equivalent
(external) dose (HED)
Human equivalent
(external) dose (HED)
PBPK MODELING BW3'4 SCALING
Figure 2-1. Conceptual approach to dose-response modeling for oral exposure.
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Nervous System Effects
Incidence data for convulsions from Grouse etal. (2006) and Cholakis etal. (19801 were
amenable to BMD modeling. For Grouse etal. (20061. statistical analysis conducted by EPA
indicated no significant difference in convulsion rates of male and female rats (Mantel-Haenszel test
for independence; see Table 2-2); thus, combined incidence data from male and female rats were
used for modeling convulsion data from this study. A BMR of 1% extra risk (ER) for convulsions
was used to address the severity of this endpoint; modeling with 5 and 10% ER is provided in
Appendix D (see Section D.1.2, Tables D-3 to D-14) for comparative purposes. In general, for
noncancer effects, severe endpoints are not typically used as the basis of a noncancer risk value
because of relatively high uncertainty in extrapolating to a level of exposure likely to be without
appreciable risk. The use of a 1% ER BMR for convulsions in Grouse etal. (2006) resulted in
extrapolation below the range of the experimental doses. However, the BMD of 3.02 mg/kg-day
was not far below the dose range of 4-15 mg/kg-day used in the study; thus, this extrapolation was
considered moderate. In addition to uncertainty from extrapolation, model uncertainty from the
use of the 1% ER BMR can be a concern. However, the BMDLs from Grouse etal. (2006) ranged
from 0.54 to 2.90, a 5.4-fold difference, which is also not considered large, so the use of a 1% ER
BMR did not result in substantial model uncertainty.
Because incidence data for convulsions were not provided by Levine etal. (1983b). a
NOAEL was used as the POD for this dataset rather than a BMDL.
Table 2-2 summarizes the PODs derived for each data set. More detailed BMD modeling
information is available in Appendix D.
Kidney/Urogenital System Effects
Incidence data on suppurative prostatitis as reported by Levine etal. (1983b) were
amenable to BMD modeling. A BMR of 10% ER was applied under the assumption that it represents
a minimally biologically significant level of change. Table 2-2 summarizes the POD derived using
data on the incidence of suppurative prostatitis. More detailed BMD modeling information is
available in Appendix D.
Male Reproductive Effects
Incidence data on testicular degeneration as reported by Lish etal. (1984) were amenable
to modeling. A BMR of 10% ER was applied under the assumption that it represents a minimally
biologically significant level of change. Table 2-2 summarizes the POD derived using date on the
incidence of testicular degeneration. More detailed BMD modeling information is available in
Appendix D.
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Table 2-2. Summary of derivation of PODs following oral exposure to RDX
Endpoint and
reference
(exposure
duration/route)
Species/sex
Model3
BMR
BMD
(mg/kg-d)
BMDL
(mg/kg-d)
PODhed (mg/kg-d)
Admin-
istered
doseb
RDX
AUC
RDX
Cmax"
Nervous system
Incidence of
convulsions
Crouse et al. (2006)
(90-d/gavage)
Male and
female F344
rat, combined6
Multistage 2°
1% ER
3.02
0.57
0.14
0.28
0.37
Incidence of
convulsions
Cholakis et al. (1980)
(GDs 6-19/gavage)
Female F344
rat
Quantal-
linear
1% ER
0.18
0.12
0.03
0.06
0.08
Incidence of
convulsions
Levine et al. (1983b)
(2-yr/diet)
Male and
female F344
rat
LOAEL = 40 mg/kg-d; NOAEL = 8 mg/kg-df
1.9
3.9
4.3
Kidney/urogenital system
Incidence of
suppurative
prostatitis
Levine et al. (1983b)
(2-yr/diet)
Male F344 rat
LogProbit
10% ER
1.67
0.47
0.11
0.23
0.25
Male reproductive system
Incidence of
testicular
degeneration
Lish et al. (1984)
(2-yr/diet)
Male B6C3Fi
mouse
LogProbit
10% ER
56.0
16.3
2.4
0.08
0.18
aFor modeling details, see Appendix D.
bPOD was converted to an HED using a standard DAF based on BW3/4.
cPOD was converted to an HED based on the equivalence of internal RDX dose (expressed as AUC for RDX
concentration in arterial blood) derived using PBPK models.
dPOD was converted to an HED based on the equivalence of internal RDX dose (expressed as peak RDX
concentration in arterial blood, Cmax) derived using PBPK models.
eExact Mantel-Haenszel test for independence between convulsion incidence and sex, stratified by dose, yielded
p-value >0.05.
'Nervous system effects for male and female rats reported qualitatively; incidence of convulsions and other
nervous system effects was not reported. Therefore, available data do not support BMD modeling.
AUC = area under the curve; BW = body weight; DAF = dosimetric adjustment factor; ER = extra risk; HED = human
equivalent dose
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Human Extrapolation
EPA guidance fU.S. EPA. 20111 describes a hierarchy of approaches for deriving human
equivalent doses (HEDs) from data in laboratory animals, with the preferred approach being PBPK
modeling. Other approaches can include using chemical-specific information in the absence of a
complete PBPK model. In lieu of either reliable, chemical-specific models or data to inform the
derivation of human equivalent oral exposures, a body weight scaling to the % power (i.e., BW3/4)
approach is generally applied to extrapolate toxicologically equivalent doses of orally administered
agents from adult laboratory animals to adult humans.
Candidate PODs for endpoints selected from rat and mouse bioassays were expressed as
HEDs. HEDs were derived using both PBPK modeling (with alternative measures of internal dose),
and a BW3/4 scaling approach. These approaches are outlined in Figure 2-1, and the resulting
PODhed values are presented in Table 2-2.
Extrapolation using PBPK modeling. PBPK models for RDX in rats, humans, and mice have
been published (Sweeney etal.. 2012a: Sweeney etal.. 2012b: Krishnan et al.. 20091 based on RDX-
specific data. EPA evaluated and further developed these models for extrapolating doses from
animals to humans (see Appendix C, Section C.1.5). In general, appropriately chosen internal dose
metrics are expected to correlate more closely with toxic responses than external doses for effects
that are not occurring at the point of contact fMcLanahan etal.. 20121. Therefore, PBPK model-
derived arterial blood concentration of RDX is considered a better dose-metric for extrapolation of
health effects than administered dose when there is adequate confidence in the estimated value.
The PBPK models for RDX were used to estimate two dose metrics: the area under the curve (AUC)
and the peak concentration (Cmax) for RDX concentration in arterial blood. The AUC represents the
average blood RDX concentration for the exposure duration normalized to 24 hours and the Cmax
represents the maximum RDX concentration for the exposure duration.
It appears logical to use RDX concentration levels in the brain as the internal dose metric for
analyzing convulsion data. Nevertheless, the blood concentration of RDX was preferred as the dose
metric due to greater confidence in modeling this variable. This is because of the substantially
greater number of measurements of RDX blood levels used in calibrating model parameters.
Additionally, predictions of RDX concentrations in the brain are highly correlated with RDX blood
concentrations because the brain compartment does not have absorption, metabolism, or
elimination of RDX. Greater confidence was placed in model estimates of blood AUC than peak
blood concentrations because, as discussed in Appendix C, Section C.1.5, the rate constant for oral
absorption (KAS) is uncertain, and peak concentrations are more sensitive to variations in this
parameter than average values. RDX-induction of convulsions and seizures appears to be more
strongly correlated with dose than exposure duration, which might argue for use of peak blood
concentration as an appropriate dose metric; however, biological support for blood AUC, rather
than peak blood concentration, comes from mechanistic information on RDX binding at the
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picrotoxin convulsant site of the gamma-amino butyric acid (GABA) channel. There is evidence
from examination of picrotoxin binding to GABAAthat a resulting period of elevated neuronal
activity post-exposure could result in increased likelihood of seizures developing over time or other
longer-term effects on normal brain function (see Section 1.2.1 for further discussion). Therefore,
the AUC for RDX concentration in arterial blood was selected as the internal dose metric for
analyzing dose-response data for convulsions. Nevertheless, the PODhed values based on both
blood AUC and peak blood concentration (Cmax) are presented in Table 2-2 for completeness.
The rodent PBPK model was applied to the BMDLs generated from BMD modeling to
determine the animal internal dose, expressed as the AUC of RDX blood concentration, and
representing the cross-species toxicologically equivalent (internal) dose. The human PBPK model
was then applied to derive the corresponding HEDs (see Figure 2-2). Because the AUC is linear
with exposure level, at least in the exposure range of interest, the value of the HED would be the
same whether the rat or mouse PBPK model is applied before or after BMD modeling is performed.
Because the sequence of the calculation does not influence the results, applying the PBPK model
after BMD modeling is more efficient—BMD modeling would not have to be redone if there were
changes to the PBPK model, and it is easier to evaluate and show two dose metrics (as discussed
above). Because of relatively high confidence in the PBPK models developed for the rat and human,
these models were used to derive reliable internal dose metrics for extrapolation. For datasets
selected from the rat bioassays, the candidate oral values were calculated assuming cross-species
toxicological equivalence of the AUC of RDX blood concentration derived from PBPK modeling. A
published PBPK model for the mouse was evaluated (Sweeney etal.. 2012b): however, major
uncertainties were identified in this model. The mouse model was based on fitting both the
absorption and metabolic rate constants to a single set of blood concentration measurements. In
this study, the lowest dose at which RDX was detected was 35 mg/kg, an exposure level high
enough to manifest some toxicity in the chronic mouse bioassay, and except for measurements from
a single animal, all other data points were non-detects or excluded as outliers f Sweeney etal..
2012b). The type of additional data that increased confidence in the rat and human models (e.g., in
vitro measurements of RDX metabolism and RDX elimination data) are not available for mice.
Consequently, confidence in the mouse model parameter values is low. Further, there are no data
to enable characterizing the fraction of RDX that is metabolized in the mouse; this is problematic
considering evidence that indicates that the role of metabolism in RDX toxicity may differ across
species (e.g., mice may have more efficient or higher expression of the cytochrome P450 [CYP450]
enzymes). Given the high sensitivity of the model to the metabolic rate constant, the uncertainty in
mouse toxicokinetics significantly decreases confidence in using the mouse PBPK model for
predicting mouse blood RDX concentrations. (See Summary of Confidence in PBPK Models for RDX
in Appendix C, Section C.1.5 for further discussion of confidence in the mouse model.) Given the
low confidence in the mouse PBPK model, the preferred approach for determining candidate oral
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values for endpoints selected from the mouse bioassay is that based on the administered dose of
RDX extrapolated to humans using allometric BW3/4 scaling.
Extrapolation using BW3/4 scaling. HEDs were also calculated using a BW3/4 scaling
approach consistent with EPA guidance (U.S. EPA. 20111. PODs (BMDLs or NOAELs) based on the
RDX dose administered in the experimental animal study were adjusted by a standard dosimetric
adjustment factor (DAF) derived as follows:
DAF = (BWa'/'VBWi,1/4},
where
BWa = animal body weight
BWh = human body weight
Using BWa values of 0.25 kg for rats and 0.035 kg for mice and a BWh of 70 kg for humans
(U.S. EPA. 19881. the resulting DAFs for rats and mice are 0.24 and 0.15, respectively. Applying the
DAF to the POD identified for effects in adultrats or mice yields a PODhed as follows (see Table 2-2):
PODhed = laboratory animal dose (mg/kg-day) x DAF
Further details of the BMDL modeling, BMDS outputs, and graphical results for the best fit
model for each dataset included in Table 2-2 can be found in Appendix D, Section D.l. Details of the
PBPK model evaluation used for extrapolation from BMDL values can be found in Appendix C,
Section C.1.5. Table 2-2 summarizes the results of the BMD modeling and the PODhed for each data
set discussed above.
2.1.3. Derivation of Candidate Values
Under EPA's A Review of the Reference Dose and Reference Concentration Processes (U.S. EPA.
20021 (Section 4.4.5), and as described in the Preamble, five possible areas of uncertainty and
variability were considered when determining the application of UFs to the PODs presented in
Table 2-2. An explanation follows:
An intraspecies uncertainty factor, UFh, of 10 was applied to all PODs to account for
potential differences in toxicokinetics and toxicodynamics in the absence of information on the
variability of response in the human population following oral exposure to RDX. The available
human pharmacokinetic data are not sufficient to inform human kinetic variability and derive a
chemical-specific UF for intraspecies uncertainty.
An interspecies uncertainty factor, UFa, of 3 (101/2 = 3.16, rounded to 3) was applied to all
PODs to account for uncertainty in characterizing the toxicokinetic and toxicodynamic differences
between rodents and humans. For the testicular degeneration dataset from the mouse bioassay,
mouse to human extrapolation was accomplished using BW3/4 scaling (see rationale in
Section 2.1.2—Human Extrapolation), which addresses predominantly toxicokinetic and some
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toxicodynamic aspects of cross-species extrapolation; residual uncertainty in toxicokinetic and
toxicodynamic extrapolation remains. In the absence of chemical-specific data to quantify this
uncertainty, EPA's BW3/4 guidance fU.S. EPA. 20111 recommends use of an uncertainty factor of 3.
For datasets from the rat bioassays, a PBPK model was used to convert internal doses in rats to
external doses in humans (see rationale in Section 2.1.2—Human Extrapolation). This reduces
toxicokinetic uncertainty in extrapolating from the rat to humans, but does not account for
interspecies differences due to toxicodynamics. A UFAof 3 was applied to account for this
remaining toxicodynamic and any residual toxicokinetic uncertainty not accounted for by the PBPK
model.
A subchronic to chronic uncertainty factor, UFS, of 1 was applied to all PODs. This is because
(1) in studies of subchronic or gestational exposure used to derive a POD, effects were seen at
lower doses in the studies of shorter duration than in chronic studies, and (2) other studies used to
derive a POD were of 2-year duration. Although EPA guidance recommends a default UFs of 10 on
the assumption that effects in a subchonic study would occur at approximately 10-fold higher
concentration than in a corresponding (but absent) chronic study fU.S. EPA. 20021. the RDX
database does not support a UFs of 10. As discussed in Section 1.2.1, although Gerkin etal. (20101
introduces the possibility of effects developing over longer-term exposures to RDX, in general,
seizure induction appears to be more strongly correlated with dose level than with exposure
duration. The available bioassays suggest that chronic exposure would not lead to effects at lower
doses than those induced by subchronic exposure. In addition, chronic dietary doses associated
with convulsions were >35 mg/kg-day and were at least fourfold higher than gavage doses that
induced convulsions in 14- and 90-day studies (i.e., 2 mg/kg-day in Cholakis etal. (19801 and
8 mg/kg-dav in Grouse etal. (200611 (also see Table 1-3 and Figure 1-1). This may be due to
differences between dietary and gavage administration (see Sections 2.1.1 and 2.1.7). Nevertheless,
these studies do not support the default expectation of observing effects in chronic studies at
approximately 10-fold lower exposure levels than in subchronic studies. Accordingly, a UFS of 1
was applied to PODs derived from studies of less-than-chronic duration.
A LOAEL to NOAEL uncertainty factor, UFl, of 1 was applied to all POD values because the
POD was a BMDL or a NOAEL. When the POD is a BMDL, the current approach is to address this
factor as one of the considerations in selecting a BMR for BMD modeling. In this case, the BMR for
modeled endpoints was selected under the assumption that the BMR represents a minimal,
biologically significant change for these effects.
A database uncertainty factor, UFd, of 3 was applied to all POD values. The oral toxicity
database for RDX includes subchronic and chronic toxicity studies in the rat and mouse, a 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 monkey. As discussed
below, some uncertainty is associated with characterization of the RDX neurotoxicity.
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Analyses presented in Section 2.1.6 note that reference values derived from mortality data
would be similar to the RfD for RDX based on convulsions. EPA prefers to identify reference values
based on upstream (less severe) effects that would precede frank effects like convulsions and
mortality. Some uncertainty remains in our understanding of RDX-induced neurotoxicity. In part,
this is due to limitations in study design to assess neurotoxicity across the RDX database; the
frequency of animal observations in the available studies raises concerns that there may be
underreporting of the true incidence of convulsions, and in general the reporting of this effect does
not include a measure of the severity at the time of observation. No follow-up studies were
identified that employed more sensitive assays to assess more subtle neurotoxicity. Uncertainties
in the database for RDX neurotoxicity could be addressed by:
• Analysis of "convulsions" using more detailed behavioral scoring methods. In the available
studies, "convulsion" can indicate a range of observable behaviors in response to altered
brain activity, ranging from involuntary limb and facial twitches to tonic-clonic seizures in
which animals exhibit a sustained (seconds to hours) and widespread loss of muscle control
sometimes resulting in respiratory arrest and/or death. As there are studies where
convulsions occur at the same dose as mortality, the convulsive activity in these studies is
interpreted as severe. Scoring methods quantifying the occurrence of different behavioral
aspects of the RDX-induced convulsions, such as the Racine scale (Racine. 1972). employed
in Burdette etal. (1988) would provide a much more accurate, complete, and possibly more
sensitive measure of RDX neurotoxicity.
• Additional electrophysiological measures of epileptiform activity. Well-established and
sensitive methods for evaluating brain activity exist These measures could not only better
describe the profile of RDX-induced convulsant activity, but could also be used to identify
and quantify sub-convulsive effects of RDX exposure (e.g., EEG spiking).
Electrophysiological characterization of the effects of RDX in vitro and in vivo has already
been demonstrated by Williams etal. (2011). Additional studies building on this work,
looking at the effects of different concentrations of RDX, could potentially identify more
sensitive measures of RDX neurotoxicity.
• A FOB conducted by Grouse etal. (2006) provides information on neurobehavioral effects
associated with RDX exposure, yet the results of that study did not identify notable effects
associated with RDX exposure. While some components of the FOB testing conducted by
Grouse etal. (2006) would be expected to give a screening-level evaluation of some stimuli-
induced behaviors that have the potential to be related (e.g., response to handling, touch,
click or open field), additional studies addressing whether RDX exposure alters the
susceptibility to seizures elicited by traditional means could be informative. Burdette et al.
Q9881 examined seizure susceptibility in gavaged male Long Evans rats, but at doses >10
mg/kg. Further evaluation of seizure susceptibility at doses lower than 10 mg/kg, and with
longer exposure durations, may identify additional measures of RDX neurotoxicity.
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• Further evaluation of potential developmental neurotoxicity (and specifically seizure
induction) associated with RDX exposure. Models for examining seizure-related behaviors
during development exist, mainly involving manipulation and analyses in pre-weanling
rodents. Hess-Ruth et al. f20071 reported possible transfer of RDX to offspring during
gestation, as well as the presence of RDX in the milk of dams, indicating a potential for
lactational transfer of RDX to offspring. Although examination of specific developmental
neurotoxicity endpoints has not been conducted in studies of RDX toxicity, the available
testing, including a two-generation reproductive toxicity study in the rat fCholakis etal..
19801. did not report any evidence of neurobehavioral effects in offspring exposed during
gestation or lactation. However, confidence in the observation is reduced as there is a
question if the extent of observation in Cholakis etal. f 19801 was sufficient to accurately
characterize neurobehavioral effects. Additional developmental neurotoxicity studies could
further rule out the possibility that RDX exposure during development might result in
immediate or delayed seizure activity, or predispose animals to developing seizures as
adults.
Overall, while the RDX database adequately covers major systemic effects, including reproductive
and developmental effects, uncertainties in the adequacy of the database were identified in
characterization of the neurotoxicity hazard. There is some concern that additional studies
described above may lead to identification of a more sensitive endpoint or a lower POD.
Accordingly, a UFd of 3 was applied to all derived PODs.
Table 2-3 is a continuation of Table 2-2 and summarizes the application of UFs to each
PODhed to derive a candidate value for each data set The candidate values presented in Table 2-3
are preliminary to the derivation of the organ/system-specific reference values. These candidate
values are considered individually in the selection of a representative oral reference value for a
specific hazard and subsequent overall RfD for RDX.
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1 Table 2-3. Effects and corresponding derivation of candidate values
Endpoint and reference
PODhed3
POD
type
UFa
UFh
UFl
UFs
UFd
Composite
UF
Candidate
value
(mg/kg-d)
Nervous system (rat)
Incidence of convulsions
Crouse et al. (2006)
0.28
BMDLoi
3
10
1
1
3
100
2.8 x 10"3
Incidence of convulsions
Cholakis et al. (1980)
0.06
BMDLoi
3
10
1
1
3
100
6.0 x 10"4
Incidence of convulsions
Levine et al. (1983b)
3.9
NOAEL
3
10
1
1
3
100
3.9 x 10"2
Kidney/urogenital system (rat)
Incidence of prostate
suppurative inflammation
Levine et al. (1983b)
0.23
BMDLio
3
10
1
1
3
100
2.3 x 10"3
Male reproductive system (mouse)
Incidence of testicular
degeneration
Lish et al. (1984)
2.4
BMDLio
3
10
1
1
3
100
2.4 x 10"2
2
3 aPODHED values based on data from the rat were derived using PBPK modeling; the PODhed based on data from the
4 mouse was derived using BW3/4 adjustment (see Section 2.1.2 and discussion of the PBPK models above and in
5 Appendix C, Section C.l.5).
6 Figure 2-2 presents graphically the candidate values, UFs, and PODhed values, with each bar
7 corresponding to one data set described in Tables 2-2 and 2-3.
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Convulsions; Crouse et
al. (2006)
Convulsions; Cholakis et
al. (1980)
Convulsions; Levine et
al. (1983b)
Prostate - suppurative
inflammation; Levine et
al. (1983b)
Testicular
degeneration; Lish etal.
(1984)
~ Candidate RfD
• PODhed
0.0001
0.001
0.01 0.1
1
Composite UF
mg/kg-day
10
1 Figure 2-2. Candidate values with corresponding POD and composite UF.
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2.1.4. Derivation of Organ/System-Specific Reference Doses
Table 2-4 distills the candidate values from Table 2-3 into a single value for each organ or
system. Organ- or system-specific reference values may be useful for subsequent cumulative risk
assessments that consider the combined effect of multiple agents acting at a common site.
Table 2-4. Organ/system-specific RfDs and overall RfD for RDX
Effect
Basis
RfD
(mg/kg-d)
Study exposure
description
Confidence
Nervous system
Incidence of convulsions
(Crouse et al., 2006)
3 x 10"3
Subchronic
Medium
Kidney/urogenital system
Incidence of suppurative prostatitis
(Levine et al., 1983b)
2 x 10"3
Chronic
Low
Male reproductive system
Incidence of testicular degeneration
(Lish etal., 1984)
2 x 10"2
Chronic
Low
Overall RfD
Nervous system
3 x 10"3
Subchronic
Medium
Nervous System Effects
The organ/system-specific RfD for nervous system effects was based on the incidence of
convulsions in F344 rats reported in Grouse etal. (2006). a well-conducted study that used a
99.99% pure form of RDX, five closely-spaced dose groups that provided a good characterization of
the dose-response curve for convulsions, and an endpoint (convulsions) that was replicated across
multiple studies. Although the candidate value derived from the developmental toxicity study in
F344 rats by Cholakis etal. (1980) is lower (by approximately fivefold), there is greater certainty in
the value derived from Grouse etal. (2006). Grouse etal. (2006) was specifically designed to assess
the nervous system effects of RDX (including a functional observational battery), whereas Cholakis
etal. (1980) was designed as a developmental toxicity study with only routine monitoring of
clinical signs (the methods section states that "Dams were monitored daily for toxic signs"). Grouse
etal. (2006) used five dose groups (plus the control) that provided good resolution of the dose-
response curve for RDX-induced convulsions, whereas Cholakis etal. f 19801 used only three dose
group (plus the control) with order of magnitude dose spacing, resulting in a less well defined
characterization of the dose-response curve for this endpoint. Further, Crouse etal. (2006) used a
higher purity test material than did Cholakis etal. (1980) (99.99% versus 88.6%, respectively).
Finally, the Crouse etal. (2006) study used a longer exposure duration (90 days) than did the
Cholakis etal. (1980) study (14 days), and is more representative of a chronic exposure duration.
The lower candidate reference value from the Cholakis etal. (1980) developmental toxicity study
could indicate that pregnant animals are a susceptible population, which could support selection of
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this study as the basis for the RfD; however, as discussed in Section 1.3.3, the available studies in
pregnant and nonpregnants rats cannot be directly compared, and the available information is not
considered sufficient to identify pregnant animals as a susceptible population.
As discussed in Section 2.1.1, the 2-year dietary study by Levine etal. f !983bl was also
considered for RfD derivation because the available oral studies suggest that bolus doses of RDX
received with gavage administration may induce nervous system effects at doses lower than those
resulting from dietary administration (recognizing that differences in particle size and purity of the
test material may confound direct comparisons between gavage and dietary administration).
Convulsion data from Levine etal. (1983b) yielded a PODhed 14-fold higher than the PODhed derived
from Grouse etal. (2006). The POD derived from the Levine etal. Q983bl study is considered less
certain than that derived from Grouse etal. (2006). Levine etal. (1983b) did not provide
information on the incidence of neurotoxic effects, and BMD analysis was thus not supported (i.e.,
the POD was based on a NOAEL). As discussed in Section 1.2.1, the frequency of daily observations
in the Levine etal. (1983b) study may not have been sufficient to provide an accurate measure of
the occurrence of nervous system effects, potentially leading to underestimation of convulsions and
other nervous system effects. For these reasons, and in light of the fact that data from the Levine et
al. fl983bl study yielded a higher POD, Levine etal. Q983bl was not used as the basis for the
organ/system-specific RfD for nervous system effects.
Kidney/Urogenital Effects
A single data set for incidence of suppurative prostatitis in male F344 rats as reported in a
2-year dietary study by Levine etal. (1983b) was brought forward for quantitative analysis as a
sensitive marker for the broader array of RDX-associated effects observed in the urogenital system.
The RfD for kidney and other urogenital effects is based on this dataset.
Male Reproductive Effects
A single dataset for male reproductive effects, specifically the incidence of testicular
degeneration as reported in male B6C3Fi mice exposed to RDX in diet for 24 months (Lish etal..
1984). was brought forward for quantitative analysis. The RfD for male reproductive effects is
based on this dataset.
2.1.5. Selection of the Overall Reference Dose
Multiple organ/system-specific reference doses were derived for effects identified as
potential hazards from RDX exposure, including nervous system effects, kidney and other
urogenital effects, and male reproductive effects. Evidence for nervous system effects, and
specifically convulsions, was observed in multiple studies, in multiple species, and following a range
of exposure durations. In addition, the organ/system-specific RfD for nervous system effects was
the lowest among the organ/system-specific RfDs derived for RDX. Evidence for dose-related
effects on the urogenital system comes primarily from a single 2-year toxicity study in male rats
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(Levine etal.. 1983b). and evidence for male reproductive effects comes primarily from a single
2-year toxicity study in mice (Lish etal.. 19841: neither a second chronic study in the rat that
evaluated prostate histopathology nor a second mouse study was available to validate and replicate
these findings.
The organ/system-specific RfD of 3 x 10~3 mg/kg-day for nervous system effects in the rat
as reported by Grouse etal. (2006) is selected as the overall RfD for RDX given the strength of
evidence for the nervous system as a hazard of RDX exposure, and as the lowest organ/system-
specific RfD. This overall RfD provides an exposure level below which effects associated with RDX
exposure are not expected to occur.
The overall RfD is derived to be protective of all types of effects for a given duration of
exposure, and is intended to protect the population as a whole, including potentially susceptible
subgroups fU.S. EPA. 20021. Decisions concerning averaging exposures over time for comparison
with the RfD should consider the types of toxicological effects and specific lifestages of concern.
Fluctuations in exposure levels that result in elevated exposures during these lifestages could
potentially lead to an appreciable risk, even if average levels over the full exposure duration were
less than or equal to the RfD. In the case of RDX, no specific lifestages have been identified as a
potentially susceptible subgroup.
2.1.6. Comparison with Mortality LDoiS
As previously discussed, mortality was considered in discussions of other organ/system-
specific toxicity (and in particular, effects on the nervous system and kidney). EPA did not develop
a candidate RfC from mortality because EPA generally does not develop reference values based on
frank effects such as mortality, rather, reference values are generally based on earlier (less severe)
upstream events, where possible, in order to protect against all adverse outcomes. Nevertheless,
additional analysis of mortality data was undertaken because some studies (see Table 2-5)
identified mortality at the same RDX dose that induced nervous system effects fCrouse etal. f20061:
Angerhofer etal. (1986): Cholakis etal. (1980): von Oettingenetal. (1949)1.
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Table 2-5. Comparison of dose levels associated with mortality and
convulsions in selected studies
Study
Doses associated with
mortality
Doses associated with
convulsions
Crouse et al. (2006)
Rats, F344,10/sex/group
0, 4, 8,10,12, or 15 mg/kg-d
13 wks/gavage
>8 mg/kg-d
>8 mg/kg-d
von Oettingen et al. (1949)
Rats, sex/strain not specified, 20/group
0,15, 25, or 50 mg/kg-d
13 wks/diet
>25 mg/kg-day
>25 mg/kg-d
Cholakis et al. (1980)
Rats, F344, 24-25 females/group
0, 0.2, 2.0, or 20 mg/kg-d
GDs 6-19/gavage
20 mg/kg-d
Primarily 20 mg/kg-d; 1
convulsion at 2 mg/kg-d
Angerhofer et al. (1986)
Rats, Sprague-Dawley, 39-51 mated females/group
0, 2, 6, or 20 mg/kg-d
GDs 6-15/gavage
Primarily at 20 mg/kg-d, but
one death each at 2 and
6 mg/kg-d
20 mg/kg-d
A discussion of mortality evidence for RDX is presented in Appendix C, Section C.3.1, and the
relationship between mortality and nervous system effects in Sections 1.2.1 and 1.3.1. Unscheduled
deaths were observed as early as day 8 of a 90-day gavage study fCrouse etal.. 20061 and in
development toxicity studies with exposure durations of two weeks fAngerhofer et al. f 19861:
Cholakis etal. f 198011.
Given the proximity in the dose at which mortality and nervous system effects were
observed in several studies, the dose-response relationships for mortality were compared across
studies with similar durations to those in Table 2-5 by comparing the LD0i (the dose expected to be
lethal to 1% of the animals) or NOAELs derived from each study. In addition, these LDoiS and
NOAELs were compare to the BMDoi for convulsions used to derive the RfD.14 Interpretation of
mortality data from chronic exposure studies in mice and rats is complicated by other treatment-
related effects and pathology regularly observed in aging animals (e.g., kidney pathology, neoplastic
lesions), and was not considered in this analysis. Other studies that were less informative and not
considered in this analysis are not presented in Table 2-6.15
14 BMDs were compared, as opposed to BMDLs, because, as stated on p. 20 of the BMD Technical Guidance
fU.S. EPA. 2012bl. "In general, it is recommended that comparisons across chemicals/studies/endpoints be
based on central estimates; this is in contrast to using lower bounds for PODs for reference values..."
15The following less informative studies were not included in the analysis of early mortality:
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Table 2-6. Summary of dose-response evaluation for mortality following oral
exposure to RDX
Reference
(exposure duration/route)
Species/sex
Model3
BMR
LDoi
(mg/kg-d)
LDLoi
(mg/kg-d)
Diet studies
Lish etal. (1984)
(11-week data from 2-yr
study/diet)
Male and female
B6C3Fi mouse
Not amenable to
modeling
NOAEL: 35 mg/kg-day
95% CI for response: 0-4%
Levine et al. (1981a)
(13-wk/diet)
Male and female
F344 rat, combined
Multistage 4°
1% ER
7.8
2.2
von Oettingen et al. (1949)
(13-wk/diet)
Rats, sex/strain not
specified
Not amenable to
modeling
NOAEL: 15 mg/kg-day
95% CI for response: 0-15%
Cholakis et al. (1980)
(2-generation design/diet)
Female CD rat
Not amenable to
modeling
NOAEL: 16 mg/kg-day
95% CI for response: 0-13%
Levine et al. (1983b)
(13-week data from 2-yr
study/diet)
Male and female
F344 rat
NA
(no mortality at
highest dose tested)
NOAEL: 40 mg/kg-day
95% CI for response: 0-4%
Cholakis et al. (1980)
(13-wk/diet)
Male and female
F344 rat
NA
(no mortality at
highest dose tested)
NOAEL: 40 mg/kg-day
95% CI for response: 0-25%
Gavage studies
Crouse et al. (2006)
(90-d/gavage)
Male and female
F344 rat, combined
Multistage 2°
1% ER
2.1
0.46
Cholakis et al. (1980)
(GDs 6-19/gavage)
Female F344 rat
Not amenable to
modeling
NOAEL: 2 mg/kg-day
95% CI for response: 0-12%
Angerhofer et al. (1986)
(GD 6-15/gavage)
Female SD rat
Multistage 3°
1% ER
1.7
0.59
Cholakis et al. (1980)
(GDs 7-29/gavage)
Female New
Zealand white rabbit
NA
(no mortality at
highest dose tested)
NOAEL: 20 mg/kg-day
95% CI for response: 0-22%
13-week dietary study in the mouse by Cholakis etal. (19801. Mortality was observed only in the high-
dose group (257-276 mg/kg-day TWA), and the unusual dosing regimen precluded identification of a
NOAEL or LOAEL.
13-week dietary study in the dog by Hart (19741 and 13-week study in the monkey by Martin and Hart
(19741. Both studies used small group sizes (3 animals/dose group), and no animals died on study
(although one high-dose monkey was euthanized).
6-week dietary study in the dog from the 1949 publication by von Oettineen etal. (19491. This dog study
included only one treatment group and recorded only one death.
30-day gavage study in the rat by MacPhail etal. (19851. The authors did not identify treatment-related
mortality, but reporting was limited.
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Reference
LDoi
LDLoi
(exposure duration/route)
Species/sex
Model3
BMR
(mg/kg-d)
(mg/kg-d)
aFor modeling details, see Appendix D, Section D.1.3.
ER = extra risk
LDoi = dose expected to be lethal to 1% of the animals
LDLoi = lower confidence limit on the LDoi
Of the studies in Table 2-6, dose-response analysis was conducted for all studies that
showed an increased incidence of unscheduled deaths. LD values are provided in Table 2-6, and
detailed modeling results are provided in Appendix D, Section D.1.3. Mortality was observed only
at the highest dose tested at week 11 in the 2-year mouse study by Lish etal. (1984). in the 13-week
rat study by von Oettingen etal. (1949). and in the two-generation reproductive and developmental
toxicity studies by Cholakis etal. (1980). In these cases, data were not amenable to LDoi estimation,
and a NOAEL (with its associated confidence interval, CI) was used in this comparative analysis
instead.
LDoi values for mortality in Table 2-6 range from 1.7 mg/kg-day (10-day gavage exposure
in pregnant rats) to 7.8 mg/kg-day (13-week dietaiy exposure in rats), with the lower values
generally from studies that administered RDX by gavage. These values may be compared to the
BMD for convulsions from Grouse etal. (2006) that was used as basis for the overall RfD for RDX.
The BMD for convulsions of 3.0 mg/kg-day is in the middle of the distribution of calculated LDoiS,
and the lowest LDoi of 1.7 mg/kg-day is within twofold of the convulsion BMD of 3.0 mg/kg-day.
The NOAELs from studies where mortality was observed tend to be higher than the LDs.
However, NOAELs are not directly comparable to BMDoiS for several reasons. Confidence intervals
for the observed responses of 0% characterize some statistical uncertainty for NOAELs from
studies that could not be modeled (note that the upper bound on an observed response of 0% is not
directly comparable to a lower bound on a benchmark dose). The confidence intervals suggest that
comparable 1% levels for these datasets could be somewhat lower than the NOAELs. On the other
hand, dose-spacing can affect the interpretation of NOAELs, such that from the Cholakis et al.
(19801 developmental toxicity study because of the wide (order-of-magnitude) spacing between
doses in that study (i.e., the reported NOAEL is 10-fold lower than the dose associated with 17%
mortality at 20 mg/kg-day).
In general, this comparison indicates that reference values derived from mortality data
would be similar to the final RfD for RDX based on convulsions, assuming the application of the
same extrapolation procedures and uncertainty factors. However, the similarity in RDX doses
associated with both mortality and nervous system effects should be taken into consideration when
using this RfD, and in particular in evaluating exposures that exceed the RfD in light of the severity
of mortality.
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2.1.7. Uncertainties in the Derivation of the Reference Dose
To derive the RfD, the UF approach fU.S. EPA. 2000a. 19941 was applied to a PODhed based
on nervous system effects in rats exposed to RDX for a subchronic duration. UFs were applied to
the PODhed values to account for uncertainties in extrapolating from an animal bioassay to human
exposure, the likely existence of a diverse human population of varying susceptibilities, and
subchronic to chronic duration. For the most part, these extrapolations are carried out with default
approaches given the lack of data to inform individual steps. One exception is the use of PBPK
modeling to perform interspecies (i.e., rat to human) extrapolation. Uncertainties associated with
the PBPK models are considered in Appendix C, Section C.1.5.
Nervous system effects have been documented in multiple studies and animal species and
strains; however, some uncertainty is associated with the incidence of reported neurological effects
in studies that employed a study design that did not monitor animals with sufficient frequency to
accurately record neurobehavioral effects, including convulsions. In the study used to derive the
RfD (Grouse etal.. 20061. lohnson f201 Sal noted that convulsions were observed infrequently
outside the dosing period; more often, seizures were observed during the 2-hour (gavage) dosing
period, typically within 60-90 minutes of dosing. Similar information was not available for other
studies to assess the likelihood that observations of convulsions were missed. However, animals
were not monitored continuously during the Grouse etal. (20061 study, and investigators reported
that nearly all observed pre-term deaths in rats exposed to the three higher doses were preceded
by signs of neurotoxicity. If an animal died during the study as a result of effects on the nervous
system, convulsions preceding death could have been missed, resulting in an underestimation of
the incidence of convulsions. Conversely, attributing all mortality to neurotoxicity (i.e., all deaths
were preceded by convulsions that may not have been observed) could result in an overestimation
of the incidence of convulsions. A dose-response analysis of the combined incidence of seizures and
mortality from Grouse etal. (20061 was conducted to evaluate the impact of these assumptions, as
the true convulsion incidence would likely fall somewhere between the observed convulsion
incidence and the combined incidence of convulsions and mortality. The PODhed of 0.24 mg/kg-day
for a combined incidence of convulsions and mortality16 was compared to the PODhed of
0.28 mg/kg-day for convulsions alone, indicating that the addition of mortality incidence did not
have a significant impact Therefore, the RfD based on the incidence of convulsions alone does not
appear to underestimate the toxicity associated with RDX.
Some uncertainty is also associated with the influence of the method of oral dosing on the
magnitude of dose required to induce nervous system effects. As noted in Section 1.2.1, gavage
16BMD = 2.56 mg/kg-day; BMDL = 0.49 mg/kg-day (see Appendix D.1.2 for BMD modeling results). The
PODhed value was derived using PBPK modeling (see Section 2.1.2 and discussion of the PBPK models in
Appendix C, Section C.1.5).
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administration generally induced convulsions in experimental animals at lower doses than did
dietary administration, possibly due to the bolus dose resulting from gavage administration that
could lead to comparatively faster absorption and higher peak blood concentrations of RDX. To
some extent, this uncertainty is reflected in the 14-fold difference in the candidate PODhed values
derived from the Grouse etal. (2006) (gavage administration) and Levine etal. Q983b) (dietary
administration) studies. A more rigorous examination of the effect of oral dosing method cannot be
performed because of the differences in test materials and study designs used in the available
gavage and dietary studies that could also have contributed to differences in response (e.g., test
article purity and particle size, number and spacing of dose groups, exposure duration, frequency of
clinical observations, and thoroughness of the reporting of observations).
Although the database is adequate for reference value derivation, uncertainty is associated
with the consistency in toxicity results across studies that used RDX test materials that differed in
purity, formulation, and particle size. There is evidence that differences in test material
formulation and particle size (i.e., the increased surface area associated with finely powdered RDX
allows for increased absorption) can affect oral bioavailability of RDX and subsequent toxicity (see
discussion in Appendix C, Section C.1.5, Absorption of RDX from the GI Tract).
2.1.8. Confidence Statement
A confidence level of high, medium, or low is assigned to the study used to derive the RfD,
the overall database, and the RfD itself, as described in Section 4.3.9.2 of EPA's Methods for
Derivation of Inhalation Reference Concentrations and Application of Inhalation Dosimetry (U.S. EPA.
1994). The overall confidence in this RfD is medium. Confidence in the principal study (Grouse et
al.. 2006) is high. The study was well-conducted, utilized 99.99% pure RDX, and had five closely-
spaced dose groups that allowed characterization of dose-response curves for convulsions in the
dose range of interest One limitation identified by study authors was the limited ability of the FOB
to fully identify neurobehavioral effects at doses >8 mg/kg-day due to the timing of the dosing
procedure and timing of the FOB screening. Confidence in the database is medium. The database
includes three chronic studies in rats and mice; eight subchronic studies in rats, mice, dogs, and
monkeys; two short-term studies; and four reproductive/developmental toxicity studies in rats and
rabbits (including a two-generation reproductive study). Confidence in the database is reduced
largely because of (1) differences in test material used across studies, and (2) uncertainties in the
influence of oral dosing methods. As discussed in Section 2.1.7 and Appendix C, Section C.1.5,
differences in test material formulation and particle size may affect RDX absorption and subsequent
toxicity, which in turn could influence the characterization and integration of toxicity findings
across studies. The available evidence also suggests that bolus dosing of RDX that results from
gavage administration induces neurotoxicity at doses lower than administration in the diet,
although a rigorous examination of these differences cannot be performed with the available
database. To the extent that dietary administration is more representative of potential human
exposures to RDX, the use of toxicity data from a gavage (bolus dosing) study introduces
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uncertainty in the RfD. Reflecting high confidence in the principal study and medium confidence in
the database, overall confidence in the RfD is medium.
2.1.9. Previous IRIS Assessment
The previous RfD for RDX, posted to the Integrated Risk Information System (IRIS) database
in 1988, was based on a 2-year rat feeding study by Levine etal. f !983bl. The no-observed-effect
level (NOEL) of 0.3 mg/kg-day based on suppurative inflammation of the prostate in male F344 rats
from this study was identified as the POD. An RfD of 3 x 10~3 mg/kg-day was derived following
application of an overall UF of 100 (UFa = 10, UFh = 10).
2.2. INHALATION REFERENCE CONCENTRATION FOR EFFECTS OTHER
THAN CANCER
The inhalation reference concentration (RfC, expressed in units of mg/m3) is defined as an
estimate (with uncertainty spanning perhaps an order of magnitude) of a continuous inhalation
exposure to the human population (including sensitive subgroups) that is likely to be without an
appreciable risk of deleterious effects during a lifetime. It can be derived from a NOAEL, LOAEL, or
the 95% lower bound on the benchmark concentration (BMCL), with UFs generally applied to
reflect limitations of the data used.
As discussed in Section 1.3.1, the available inhalation literature does not support
characterization of the health hazards specifically associated with chronic inhalation exposure to
RDX, nor do the studies support quantitative dose-response analysis. Of the available human
epidemiological studies of RDX fWest and Stafford. 1997: Ma and Li. 1993: Hathaway and Buck.
19771. none provided data that could be used for dose-response analysis. The studies by Ma and Li
(1993) of neurobehavioral effects in Chinese workers and West and Stafford (1997) of
hematological abnormalities in ordnance factory workers had numerous methodological
limitations that preclude their use for quantitative analysis (see Literature Search Strategy | Study
Selection and Evaluation). The study by Hathaway and Buck T19771 found no evidence of adverse
health effects in munition plant workers (based on evaluation of liver function, renal function, and
hematology), and therefore does not identify a POD at which there would be an effect from which to
derive an RfC. Multiple case reports provide some evidence of effects in humans associated with
acute exposure to RDX; however, while case reports can support the identification of hazards
associated with RDX exposure, data from case reports are inadequate for dose-response analysis
and subsequent derivation of a chronic reference value because of short exposure durations and
incomplete or missing quantitative exposure information.
As discussed in Literature Search Strategy | Study Selection and Evaluation, a single
experimental animal study involving inhalation exposure was identified in the Defense Technical
Information Center (DTIC) database; the study is not publicly available. However, the study would
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not have provided useful data on responses to inhaled RDX, as it was limited by small numbers of
animals tested, lack of controls, and incomplete reporting of exposure levels.
Therefore, the available health effects literature does not support the derivation of an RfC
for RDX. While inhalation absorption of RDX particulates is a plausible route of exposure, there are
no toxicokinetic studies of RDX inhalation absorption to support an inhalation model. Therefore, a
PBPK model for inhaled RDX was not developed to support route-to-route extrapolation from the
RfD.
2.2.1. Previous IRIS Assessment
An RfC for RDX was not previously derived under the IRIS Program.
2.3. ORAL SLOPE FACTOR FOR CANCER
The oral slope factor (OSF) is a plausible upper bound on the estimate of risk per
mg/kg-day of oral exposure. The OSF can be multiplied by an estimate of lifetime exposure (in
mg/kg-day) to estimate the lifetime cancer risk.
2.3.1. Analysis of Carcinogenicity Data
As noted in Section 1.3.2, there is "suggestive evidence of carcinogenic potential" for RDX.
The Guidelines for Carcinogen Risk Assessment fU.S. EPA. 2005al state:
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 analyses
may be useful for some purposes, for example, providing a sense of the magnitude
and uncertainty of potential risks, ranking potential hazards, or setting research
priorities.
In the case of RDX, there are well-conducted studies that tested large numbers of animals at
multiple dose levels fLish etal.. 1984: Levine etal.. 1983b! making the cancer response suitable for
dose-response analysis. Considering the data from these studies, along with the uncertainty
associated with the suggestive nature of the weight of evidence, quantitative analysis of the tumor
data may be useful for providing a sense of the magnitude of potential carcinogenic risk.
The incidences of liver and lung tumors in female mice from the study by Lish etal. (1984)
were selected for quantitative dose-response analysis. The study by Lish etal. (1984): (1) included
comprehensive histopathological examination of major organs; (2) contained four dose groups and
a control; (3) used adequate numbers of animals per dose group (85/sex/group, with interim
sacrifice groups of 10/sex/group at 6 and 12 months) and a sufficient overall exposure duration
(2 years); and (4) adequately reported methods and results (including individual animal data).
Female mouse liver tissues from the original unpublished study by Lish etal. (1984) were
reevaluated by a pathology working group (PWG) (Parker et al.. 2006) in order to apply more
up-to-date histopathological criteria established by Haradaetal. (1999). The updated liver tumor
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incidences from the PWG reanalysis ofLish etal. (1984) were used for quantitative dose-response
analysis.
In the case of both liver and lung tumors, benign and malignant tumors (i.e., adenomas and
carcinomas) were combined for dose-response analysis because benign and malignant tumors in
both organs develop from the same cell line and there is evidence for progression from benign to
the malignant stage fU.S. EPA. 2005a: McConnell etal.. 19861. Female mouse liver and lung tumor
incidences from the Lish etal. (1984) study are summarized in Appendix D, Table D-23.
The incidence of hepatocellular carcinomas in male F344 rats from the study by Levine et al.
(1983b) and the incidence of alveolar/bronchiolar carcinomas in male B6C3Fi mice from the study
by Lish etal. f19841 were also considered for quantitative dose-response analysis. Both studies
were well conducted, using similar study designs (described above). In both instances, the
response was less robust than the response observed in female mice from the Lish etal. f 19841
study. The hepatocellular carcinoma result in male F344 rats is based on a small number of tumors
(two each at 8 and 40 mg/kg-day, and one in controls). The alveolar/bronchiolar carcinomas in
male B6C3Fi mice showed a positive trend; however, a positive trend was not observed when the
incidence of adenomas and carcinomas was combined. Modeling results are provided in
Appendix D, Section D.2.3 for comparison.
2.3.2. Dose-Response Analysis—Adjustments and Extrapolation Methods
The EPA Guidelines for Carcinogen Risk Assessment (U.S. EPA. 2005al recommend that the
method used to characterize and quantify cancer risk from a chemical be determined by what is
known about the mode of action (MOA) of the carcinogen and the shape of the cancer
dose-response curve. The linear approach is recommended when there are MOA data to indicate
that the dose-response curve is expected to have a linear component below the POD or when the
weight-of-evidence evaluation of all available data are insufficient to establish the MOA for a tumor
site fU.S. EPA. 2005al. In the case of RDX, the mode of carcinogenic action for hepatocellular and
alveolar/bronchiolar tumors is unknown (see discussion of Mechanistic Evidence in Section 1.2.5).
Therefore, a linear low-dose extrapolation approach was used to estimate human carcinogenic risk
associated with RDX exposure.
The survival curves were compared across dose groups in each study to determine whether
time of death should be incorporated in the dose-response analysis of tumors. For female mice in
Lish etal. Q9841. the survival curves were determined to be similar across dose groups after the
dose was reduced in the high-dose group to 100 mg/kg-day (log-rank test, p-value >0.10);
therefore, a time-to-tumor analysis was not necessary for this study. Tumor incidence was
modeled using the multistage-cancer models in BMDS (versions 2.4 and 2.5). A standard BMR of
10% ER was applied to both tumor sites in the mouse.
Given the finding of an association between RDX exposure in the female mouse and
increased tumor incidence at two tumor sites, basing the OSF on only one tumor site could
potentially underestimate the carcinogenic potential of RDX. Therefore, an analysis that combines
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the results from the mouse liver and lung tumor incidence is preferred. The MS-COMBO procedure
(BMDS, version 2.5) extends the multistage-cancer models to the case with multiple tumors
assuming independence between tumor types. There is no known biological relationship between
liver and lung tumors in RDX-exposed mice, and therefore, as noted by the National Research
Council fNRC. 19941. this assumption of independence is considered not likely to produce
substantial error in risk estimates. MS-COMBO analyzes tumor incidence as present if either organ
(or both) has a tumor and as absent otherwise. The procedure derives a maximum likelihood
estimate of the combined risk at a 95% confidence level based on the parameter values obtained for
the individual tumor multistage model fits.
EPA's preferred approach for extrapolating results from animal studies to humans is
toxicokinetic modeling. As described in Appendix C, Section C.1.5, PBPK models for RDX in mice
and humans published by Sweeney etal. (2012 b) were evaluated and further developed by EPA.
Consideration was given to whether the available toxicokinetic information supported using an
internal dose metric derived by PBPK modeling. The available mechanistic data (Section 1.2.5)
point to some evidence, although not conclusive, that RDX-generated metabolites may be
implicated in the observed tumorigenicity in the female mouse. However, there are no data on the
toxicokinetics of RDX metabolites, and metabolism in the liver is the only route of elimination of
RDX in the PBPK model. In this case, as is to be expected from mass balance principles, the PBPK
modeling provides no further information; the HED obtained from the model-estimated amount of
total RDX metabolites scaled by BW3/4 was equal to that calculated using administered dose scaled
by BW3/4 In addition to the lack of data on metabolism, other major uncertainties were identified
in the mouse PBPK modeling; EPA's evaluation of these uncertainties is discussed in Summary of
Confidence in PBPK Models for RDX in Appendix C, Section C.1.5. Therefore, the PBPK model
developed for the mouse was not used, and consistent with the EPA's Guidelines for Carcinogen Risk
Assessment (U.S. EPA. 2005a). the preferred approach for calculating an HED from the mouse
tumors is adjustment of the administered dose by allometric scaling to achieve toxicological
equivalence across species.
As discussed in Section 2.1.1, the administered dose in animals was converted to an HED on
the basis of (body weight)3/4 (U.S. EPA. 1992). This was accomplished by multiplying administered
dose by (animal body weight in kg/human body weight in kg)1/4 (U.S. EPA. 1992). where the body
weight for the mouse is 0.035 kg and the reference body weight for humans is 70 kg fU.S. EPA.
1988). Details of the BMD modeling can be found in Appendix D, Section D.2.
2.3.3. Derivation of the Oral Slope Factor
The lifetime cancer OSF for humans is defined as the slope of the line from the BMR (10%
ER) at the BMDL to the estimated control response at zero (OSF = 0.1/BMDLio-hed). This slope, a
95% upper confidence limit on the true slope, represents a plausible upper bound on the true slope
or risk per unit dose. The PODs estimated for each mouse tumor site are summarized in Table 2-7.
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1 Using linear extrapolation from the BMDLio-hed, human equivalent OSFs were derived for each
2 tumor site individually and both sites combined and are listed in Table 2-7.
3 Table 2-7. Model predictions and OSFs for hepatocellular and alveolar/
4 bronchiolar adenomas or carcinomas in female B6C3Fi mice administered
5 RDX in the diet for 2 years fLish et al.. 19841
Tumor type
Selected
model
BMR
BMD
(mg/kg-d)
BMDL
(mg/kg-d)
BMDio-hed3
(mg/kg-d)
POD =
BMDLio-hed13
(mg/kg-d)
OSFc
(mg/kg-d)1
Hepatocellular
adenomas or
carcinomasd
Multistage 1°
10% ER
64.2e
32.6e
9.56
4.89
0.020
Alveolar/bronchiolar
adenomas or
carcinomas
Multistage 1°
10% ER
52.8
27.7
7.92
4.16
0.024
Liver + lung tumors
Multistage 1°
(MS-COMBO)
10% ER
29.0e
17.7e
4.35
2.66
0.038
6
7 aBMDio-HED = BMDio x (BWa1/4/BWh1/4), where BWa = 0.035 kg, and BWh = 70 kg.
8 "BMDLio-hed = BMDLio x (BWa1/4/BWh1/4), where BWa = 0.035 kg, and BWh = 70 kg.
9 cOSF = BMR/BMDLio-hed, where BMR = 0.1 (10% ER).
10 incidences of female mouse liver tumors from Lish et al. (1984) are those reported in the PWG reevaluation
11 (Parker et al., 2006).
12 eData for hepatocellular adenomas and carcinomas and for liver and lung tumors combined were remodeled using
13 the original sample sizes provided in Lish et al. (1984), which were slightly different for two groups than those
14 reported in Parker et al. (2006). The resulting BMDs and BMDLs from the remodeling were 64.8 and
15 32.8 mg/kg-day, respectively, for hepatocellular adenomas and carcinomas and 29.1 and 17.7 mg/kg-day,
16 respectively, for liver and lung tumors combined. See Table D-23 and the subsequent MS-COMBO results for
17 details.
18 An OSF was derived from the BMDLio-hed based on a significantly increased trend in the
19 incidence of hepatocellular and alveolar/bronchiolar adenomas or carcinomas in female B6C3Fi
20 mice (i.e., the Liver + Lung BMDLio-hed from MS-COMBO). The OSF of 0.04 (mg/kg-day)i is
21 calculated by dividing the BMR (10% ER) by the Liver + Lung BMDLio-hed and represents an upper
22 bound on cancer risk per unit dose associated with a continuous lifetime exposure:
OSF = 0.10 4- (Liver + Lung) BMDLio-hed = 0.10 4 2.66 mg/kg-day
= 3.8 x 10"2 (mg/kg-day)-1
= 4 x 10"2 (mg/kg-day)-1, rounded to one significant figure
23 The slope of the linear extrapolation from the central estimate of exposure associated with
24 10% extra cancer risk (BMDio-hed) from the same data sets is given by:
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Slope of the linear extrapolation from the central estimate
= 0.10 -7- (Liver + Lung) BMDio-hed = 0.10 -f- 4.35 mg/kg-day
= 2.3 x 10"2 (mg/kg-day)-1
= 2 x 10"2 (mg/kg-day)-1 (rounded to one significant figure)
1 The OSF for RDX should not be used with exposures exceeding the POD (2.66 mg/kg-day),
2 because above this level, the fitted dose-response model better characterizes what is known about
3 the carcinogenicity of RDX.
2.3.4. Uncertainties in the Derivation of the Oral Slope Factor
4 A number of uncertainties underlie the cancer unit risk for RDX. Table 2-8 summarizes the
5 impact on the assessment of issues such as the use of models and extrapolation approaches
6 particularly those underlying the Guidelines for Carcinogen Risk Assessment fU.S. EPA. 2005al. the
7 effect of reasonable alternatives, the approach selected, and its justification.
8 Table 2-8. Summary of uncertainty in the derivation of the cancer risk value
9 for RDX
Consideration and impact on
cancer risk value
Decision
Justification
Selection of study
The cancer bioassay in the rat
(Levine et al., 1983b) would provide
a lower estimate of the OSF
Lish et al. (1984) as
principal oral study to
derive the human
cancer risk estimate
Lish et al. (1984) was a well-conducted studv;
five dose levels (including control) used, with a
sufficient number of animals per dose group (at
terminal sacrifice, n = 62-65 female mice/dose
group except highest dose where n = 31).
Tumor data from the mouse provided a
stronger basis for estimating the OSF than rat
data. Confidence in the OSF based on rat data
was low because of the small numbers of
tumors.
Species/sex
Use of data sets from the male
mouse or male rat would provide a
lower OSF
OSF based on tumors
in female B6C3Fi
mouse
It is assumed that a positive tumor response in
animal cancer studies indicates that the agent
can have carcinogenic potential in humans in
the absence of data indicating that animal
tumors are not relevant to humans (U.S. EPA,
2005a). As there are no data to inform
whether the response in any given
experimental animal species or sex would be
most relevant for extrapolating to humans,
tumor data from the most sensitive species and
sex were selected as the basis for the OSF.
Other data sets would provide smaller OSF
values, and are not considered any more or less
relevant to humans than data from the female
mouse (i.e., 0.017 per mg/kg-day based on
hepatocellular carcinomas in male F344 rats,
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Consideration and impact on
cancer risk value
Decision
Justification
and 0.018 per mg/kg-day based on
alveolar/bronchiolor carcinomas in male
B6C3Fi mice; see Appendix D, Section D.2).
Combined tumor types
Human risk would 4^ if OSF was
based on analysis using only a
single tumor type
OSF based on liver and
lung tumors in female
B6C3Fi mouse
Basing the OSF on one tumor site could
potentially underestimate the carcinogenic
potential of RDX, so an analysis that included
data from the two tumor sites was chosen to
calculate the combined risk. Because there is
no known biological dependence between the
liver and lung tumors, independence between
the two tumor sites was assumed. This is not
likely to produce substantial error in the risk
estimates (NRC, 1994).
Selection of dose metric
PBPK models are available for the
rat, mouse, and human, and using
an appropriate internal metric can
1" accuracy in human extrapolation
Mouse liver and lung
tumors: administered
dose used
EPA evaluated a published PBPK model in the
mouse (Sweeney et al., 2012b); major
uncertainties associated with limited
toxicokinetic data in the mouse and unknown
differences in metabolism across species were
identified. Althought EPA's preferred approach
for extrapolating results from animal studies to
humans is toxicokinetic modeling, the
uncertainties associated with use of the mouse
PBPK model for RDX were considered higher
than use of administered dose.
Cross-species scaling
Alternatives could 4^ or T* OSF
(e.g., 3.5-fold 4/ [scaling by body
weight] or T* 2-fold [scaling by
BW2/3])
BW3/4 scaling (default
approach)
There are no data to support alternatives.
Because the dose metric was not an AUC, BW3/4
scaling was used to calculate equivalent
cumulative exposures for estimating equivalent
human risks. While the true human
correspondence is unknown, this overall
approach is not expected over- or
underestimate human equivalent risks.
BMD model uncertainty
Alternative models could 4/ or T*
OSF
Use multistage model
to derive a BMD and
BMDL for combined
tumor incidence
No biologically based models for RDX are
available, and there is no a priori basis for
selecting a model other than the multistage.
The multistage model has biological support
and is the model most consistently used in EPA
cancer assessments (Gehlhaus et al., 2011).
Low-dose extrapolation approach
4/ cancer risk would be expected
with the application of nonlinear
extrapolation
Linear extrapolation
from the POD
Where the available information is insufficient
to establish the MOA for tumors at a given site,
linear extrapolation is recommended because
this extrapolation approach is generally
considered to be health-protective (U.S. EPA,
2005a). Because the MOA for RDX-induced
liver and lung tumors has not been established,
linear low-dose extrapolation was applied,
consistent with EPA guidance.
Statistical uncertainty at the POD
4/ OSF by 1.6-fold if BMD used as
the POD rather than the BMDL
BMDL (default
approach for
Lower bound is 95% confidence interval (CI) on
administered exposure at 10% ER of liver and
lung tumors.
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Consideration and impact on
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Decision
Justification
calculating plausible
upper bound OSF)
Sensitive subpopulations
1" OSF to an unknown extent
Considered
qualitatively
There is little information on whether some
subpopulations may be more or less sensitive
to the potential carcinogenicity of RDX (i.e.,
because of variability in toxicokinetics or
toxicodynamics for RDX). The mode of
carcinogenic action for liver and lung tumors in
experimental animals is unknown, and little
information is available on RDX metabolites or
variation in metabolic rates that could be used
to evaluate human variability in cancer
response to RDX.
2.3.5. Previous IRIS Assessment: Oral Slope Factor
The previous cancer assessment for RDX was posted to the IRIS database in 1990. The OSF
in the previous cancer assessment was based on the bioassay by Lish etal. (1984) and analysis of
data for hepatocellular adenomas or carcinomas in female mice. An OSF of 1.1 x 10"1 (mg/kg-day)"1
was derived using a linearized multistage procedure (extra risk) and scaling by body weight to the
2/3 power for cross-species extrapolation. In addition, the previous assessment dropped the high-
dose group because the dose was reduced at week 11 to address high mortality.
2.4. INHALATION UNIT RISK FOR CANCER
The carcinogenicity assessment provides information on the carcinogenic hazard potential
of the substance in question, and quantitative estimates of risk from oral and inhalation exposure
may be derived. Quantitative risk estimates may be derived from the application of a low-dose
extrapolation procedure. If derived, the inhalation unit risk (IUR) is a plausible upper bound on the
estimate of risk per |J.g/m3 air breathed.
An IUR value was not calculated because inhalation carcinogenicity data for RDX are not
available. While inhalation absorption of RDX particulates is a plausible route of exposure, there
are no toxicokinetic studies of RDX inhalation absorption to support an inhalation model.
Therefore, a PBPK model for inhaled RDX was not developed to support route-to-route
extrapolation of an IUR from the OSF.
2.5. APPLICATION OF AGE-DEPENDENT ADJUSTMENT FACTORS
As discussed in the Supplemental Guidance for Assessing Susceptibility from Early-Life
Exposure to Carcinogens (U.S. EPA. 2005b). either default or chemical-specific age-dependent
adjustment factors (ADAFs) are recommended to account for early-life exposure to carcinogens
that act through a mutagenic MOA. Because no chemical-specific data on lifestage susceptibility for
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1 RDX carcinogenicity are available, and because the MOA for RDX carcinogenicity is not known (see
2 Section 1.2.5), application of ADAFs is not recommended.
3
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This document is a draft for review purposes only and does not constitute Agency policy.
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