## DRAFT DELIBERATIVE: FOR INTERAGENCY REVIEW ONL Y. DO NOT DISTRIBUTE OUTSIDE YOUR AGENCY

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**Intera gency** 

**R eview**  **D raft**  **www.epa.gov/iris**



## T **oxicological R**  **eview of**  tert

**- B**  **utyl Alcohol (** t ert

- **Butanol )** 

(CAS No. 75

- 65 -0)

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

September 201 4

This document is an **Interagency Science Consultation**  **Review draft** 

**.** This information is distributed solely for the purpose of pre

- dissemination peer review under applicable information quality guidelines. It has not been formally disseminated by EPA. It does not represent and should not be construed to represent any Ag ency 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 **EPA/635/R -1**  **4/ 378** Toxicological Review of tert

- Butyl Alcohol

**DISCLAIMER**

This document is a preliminary draft for review purposes only. This information is distributed solely for the purpose of pre

- dissemination peer review under applicable information quality guidelines. It has not been formally disseminated by EPA. It does not represent and should not be construed to represent any Agency determination or policy. Mention of trade names or commercial products does not constitute endorsement or recommendation for use.

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# CONTENTS

AUTHORS | CONTRIBUTORS | REVIEWERS

................................................................................................ . viii

PREFACE

................................................................................................ ........................................................ x

PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS

........................................................................................... xvii

EXECUTIVE SUMMAR Y

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

- 16 LITERATURE SEARCH STRATEGY | STUDY SELECTION AND EVALUATION

............................................... LS -1

1. HAZARD IDENTIFICATION

.................................................................................................................... 1-1

1.1. PRESENTATION AND SYNTHESIS OF EVIDENCE BY ORGAN/SYSTEM

............................................ 1-1

1.1.1. Kidney Effects

.................................................................................................................... 1-1

1.1.2. Thyroid Effects

................................................................................................................. 1- 27

1.1.3. R eproductive and Developmental Effects

....................................................................... 1- 32

1.1.4. Carcinogenicity (other than in the kidney or thyroid)

..................................................... 1- 46

1.1.5. Other Toxicological Effects

.............................................................................................. 1- 46

1.2. INTEGRATION AND EVALUATION

............................................................................................... 1- 60

1.2.1. Effects Other Than Cancer

............................................................................................... 1- 60

1.2.2. Carcinogenicity

................................................................................................................ 1- 61

1.2.3. Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes

.............. 1- 63

2. DOSE

- RESPONSE ANALYSIS

................................................................................................................. 2-1

2.1. ORAL REFERENCE DOSE FOR EFFECTS OTHER THAN CANCER

...................................................... 2-1

2.1.1. Identification of Studies and Effects for Dose

- Response Analysis

.................................... 2-1

2.1.2. Methods of Analysis

.......................................................................................................... 2-2

2.1.3. Derivation of Candidate Values

......................................................................................... 2-4

2.1.4. Derivation of Organ/System

- Specific Reference Doses

.................................................... 2-8

2.1.5. Selection of the Proposed Overall Reference Dose

.......................................................... 2-8

2.1.6. Confidence Statement

....................................................................................................... 2-9

2.1.7. Previous IRIS Assessment

.................................................................................................. 2-9

2.2. INHALATION REFERENCE CONCENTRATION FOR EFFECTS OTHER THAN CANCER

....................... 2-9

2.2.1. I d entification of Studies and Effects for Dose

- Response Analysis

.................................... 2-9

2.2.2. Methods of Analysis

........................................................................................................ 2- 11

2.2.3. Derivation of Candidate Values

....................................................................................... 2- 15

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2.2.4. Derivation of Organ/System

- Specific Reference Concentrations

................................... 2- 18

2.2.5. Selection of the Proposed Overall Reference Concentration

......................................... 2- 18

2.2.6. Confidence Statement

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

2.2.7. Previous IRIS Assessment

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

2.2.8. Uncertainties in the Derivation of the Reference Dose and Reference Concentration

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

2.3. ORAL SLOPE FACTOR FOR CANCER

............................................................................................. 2- 20

2.3.1. Analysis of Carcinogenicity Data

..................................................................................... 2- 20

2.3.2. Dose

- Response Analysis

— Adjustments and Extrapolations Methods

........................... 2- 21

2.3.3. Derivation of the Oral Slope Factor

................................................................................. 2- 22

2.3.4. Uncertainties in the Derivation of the Oral Slope Factor

................................................ 2- 23

2.3.5. Previous IRIS Assessment: Oral Slope Factor

.................................................................. 2- 24

2.4. INHALATION UNIT RISK FOR CANCER

......................................................................................... 2- 24

2.4.1. Analysis of Carcinogenicity Data

..................................................................................... 2- 25

2.4.2. Dose Response Analysis – Adjustments and Extrapolation Methods

............................. 2- 25

2.4.3. Inhalation Unit Risk Derivation

........................................................................................ 2- 26

2.4.4. Uncertainties in the Derivation of th e Inhalation Unit Risk

............................................ 2- 27

2.4.5. Previous IRIS Assessment: Inhalation Unit Risk

............................................................... 2- 29

2.5. APPLICATION OF AGE

- DEPENDENT ADJUSTMENT FACTORS

...................................................... 2- 29

REFERENCES

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

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# TABLES

Table P

- 1. Physicochemical properties and chemical identity of tert

- butanol

.......................................... xiii Table ES

- 1. Summary of reference dose (RfD) derivation

...................................................................... ES -2 Table ES

- 2. Summary of reference concentration (RfC) derivation

........................................................ ES -3 Table LS

- 1. Details of the search strategy employed for tert

- butanol

.................................................... LS -5 Table LS

- 2. Summary of additional search strategies for tert

- butanol

................................................... LS -6 Table LS

- 3. Questions and relevant experimental information for evaluation of experimental animal studies

................................................................................................................ LS -7 Table 1

- 1. Changes in kidney weight in animals following exposure to tert

- butanol

............................... 1-3 Table 1

- 2. Changes in kidney histopathology in animals following exposure to tert

- butanol

................. 1-5 Table 1

- 3. Changes in kidney tumors in animals following exposure to tert

- butanol

.............................. 1-9 Table 1

- 4. Additional kidney effects potentially relevant to mode of action in animals following exposure to tert

- butanol

............................................................................................... 1- 14 Table 1

- 5. Summary of data on the α 2u - globulin process in male rats exposed to tert

- butanol

............. 1- 16 Table 1

- 6. Summary of additional data informing the contribution of the α 2u - globulin process on the renal tumor development in male rats exposed to tert

- butanol

............................ 1- 19 Table 1

- 7. Evidence pertaining to thyroid effects in animals following oral exposure to tert

butanol

........................................................................................................................... 1- 29 Table 1

- 8. Evidence pertaining to reproductive effects in animals following exposure to tert

butanol

........................................................................................................................... 1- 34 Table 1

- 9. Evidence pertaining to developmental effects in animals following exposure to tert

butanol

........................................................................................................................... 1- 37 Table 1

- 10. Evidence pertaining to neurodevelopmental effects in animals following exposure to
tert

- butanol

.................................................................................................................... 1- 44 Table 1

- 11. Evidence pertaining to effects on body weight in animals following exposure to tert

butanol

........................................................................................................................... 1- 49 Table 1

- 12. Changes in liver weight in animals following exposure to tert

- butanol

.............................. 1- 52 Table 1

- 13. Changes in liver histopathology in animals following exposure to tert

- butanol

................. 1- 54 Table 1

- 14. Changes in urinary bladder histopathology in animals following oral exposure to
tert

- butanol

.................................................................................................................... 1- 56 Table 2

- 1. Summary of derivations of points of departure

....................................................................... 2-4 Table 2

- 2. Effects and corresponding derivation of candidate RfDs

......................................................... 2-6 Tab le 2

- 3. Organ/system

- specific RfDs and proposed overall RfD for tert

- butanol

.................................. 2-8 Table 2

- 4. Summary of derivation of PODs following inhalation exposure

............................................ 2- 13 Table 2

- 5. Summary of derivation of inhalation points of departure derived from route

- to

- route extrapolation from oral exposures

................................................................................ 2- 14 Table 2

- 6. Effects and corresponding derivation of candidate values

.................................................... 2- 16 Table 2

- 7. Organ/system

- specific RfCs and proposed overall RfC for tert

- butanol

................................ . 2- 18 Table 2

- 8. Summary of the oral slope factor derivations

........................................................................ 2- 22 Table 2

- 9. Summary of uncertainties in the derivation of cancer risk values for tert

- butanol

............... 2- 23 Table 2

- 10. Summary of the inhalation unit risk derivation

.................................................................... 2- 27 Table 2

- 11. Summary of uncertainties in the derivation of cancer risk values for tert

- butanol

............. 2- 28

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

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# FIGURES

Figure LS

- 1. Study selection strategy.

..................................................................................................... LS -4 Figure 1

- 1. Exposure response array for kidney effects following oral exposure to tert

- butanol.

......... 1- 10 Figure 1

- 2. Exposure

- response array of kidney effects following subchronic inhalation exposure
to tert

- butanol (no chronic studies available).

.............................................................. 1- 11 Figure 1

- 3. Exposure

- response array for components of α 2u - globulin nephropathy and renal tumors in male rats after oral exposure to tert

- butanol.

.............................................. 1- 17 Figure 1

- 4. Exposure

- response array for components of α 2u - globulin nephropathy and renal tumors in male rats after inhalation exposure to tert

- butanol.

.................................... 1- 18 Figure 1

- 5. Exposure

- response array of thyroid follicular cell effects following chronic oral exposure to tert

- butanol.

.............................................................................................. 1- 31 Figure 1

- 6. Exposure

- response array of reproductive and developmental effects following oral exposure to tert

- butanol.

.............................................................................................. 1- 41 Figure 1

- 7. Exposure

- response array of reproductive and developmental effects following inhalation exposure to tert

- butanol.

............................................................................. 1- 42 Figure 1

- 8. Exposure

- response array of other effects following oral exposure to tert

- butanol.

........... 1- 58 Figure 1

- 9. Exposure

- response array of other effects following inhalation exposure to tert

butanol.

.......................................................................................................................... 1- 59 Figure 2

- 1. Candidate RfD values with corresponding POD and composite UF.

....................................... 2-7 Figure 2

- 2. Candidate RfC values with corresponding POD and composite UF.

..................................... 2- 17

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# ABBREVIATIONS

α2u -g alpha 2u

- globulin ACGIH American Conference of Governmental Industrial Hygienists AIC Akaike’s information criterion ALD approximate lethal dosage ALT alanine aminotransferase AST aspartate aminotransferase
atm atmosphere ATSDR Agency for Toxic Substances and Disease Registry BMD benchmark dose BMDL benchmark dose lower confidence limit BMDS Benchmark Dose Software BMR benchmark response BUN blood urea nitrogen BW body weight CA chromosomal aberration CASRN Chemical Abstracts Service Registry Number CBI covalent binding index CHO Chinese hamster ovary (cell line cells) CL confidence limit CNS central nervous system CPN chronic progressive nephropathy CYP450 cytochrome P450 DAF dosimetric adjustment factor DEN diethylnitrosamine DMSO di methylsulfoxide DNA deoxyribonucleic acid EPA Environmental Protection Agency FDA Food and Drug Administration FEV 1 forced expiratory volume of 1 second GD gestation day GDH glutamate dehydrogenase GGT γ

- glutamyl transferase GSH glutathione GST glutathione -S- transferase Hb/g

- A animal blood:gas partition coefficient Hb/g -H human blood:gas partition coefficient HEC human equivalent concentration HED human equivalent dose i.p. intraperitoneal IRIS Integrated Risk Inform ation System IVF in vitro fertilization LC 50 median lethal concentration LD 50 median lethal dose LOAEL lowest

- observed

- adverse

- effect level MN micronuclei MNPCE micronucleated polychromatic erythrocyte MTD maximum tolerated dose NAG N- acetyl

- β -D- glucosaminidase NCEA National Center for Environmental Assessment NCI National Cancer Institute NOAEL no

- observed

- adverse

- effect level NTP National Toxicology Program NZW New Zealand White (rabbit breed) OCT ornithine carbamoyl transferase ORD Office of Research and Development PBPK physiologically based pharmacokinetic PCNA proliferating cell nuclear antigen POD p oint of departure POD [ADJ] duration

- adjusted POD QSAR quantitative structure

- activity relationship RDS replicative DNA synthesis RfC inhalation reference concentration RfD oral reference dose RGDR regional gas dose ratio RNA ribonucleic acid SAR structure activity relationship SCE sister chromatid exchange SD standard deviation SDH sorbitol dehydrogenase SE standard error SGOT glutamic oxaloacetic transaminase, also known as AST SGPT glutamic pyruvic transaminase, also known as ALT SSD systemic scleroderma TCA trichloroacetic acid TCE trichloroethylene TWA time

- weighted average UF uncertainty factor UF A animal

- to

- human uncertainty factor UF H human variation uncertainty factor UF L LOAEL

- to

- NOAEL uncertain factor UF S subchronic

- to

- chronic uncertainty factor UF D database deficiencies uncertainty factor U.S. United States of America

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### 1 

# AUTHORS | CONTRIBUTORS | REVIEWERS 2

**Assessment Team** Janice S. Lee, Ph.D. (Chemical Manager) Keith Salazar, Ph.D*

Chris Brinkerhoff, Ph.D. U.S. EPA O ffice of R esearch and D evelopment National Center for Environmental Assessment Research Triangle Park, NC *Washington, DC

ORISE Postdoctoral Fellow at U.S. EPA/ORD/NCEA Currently with U.S. EPA, Office of Chemical Safety and Pollution Prevention, Office of Pollution Prevention and Toxics Washington, DC 

### 3 **C**



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**E**



### 7 **I ontributors**  Andrew Hotchkiss , Ph.D. Channa Keshava, Ph.D. Amanda Persad, Ph.D. U.S. EPA O ffice of R esearch and D evelopment National Center for Environmental Assessment Research Triangle Park, NC

**roduction Team** Maureen Johnson Vicki Soto U.S. EPA O ffice of R esearch and D evelopment National Center for Environmental Assessment Washington, DC

**ontractor Support** Robyn Blain, Ph.D. Michelle Cawley* William Mendez, Jr., Ph.D

. Pam Ross ICF International Fairfax, VA *Research Triangle Park, NC

**xecutive Direction** Kenneth Olden, Ph.D., Sc.D., L.H.D. (Center Director) John Vandenberg, Ph.D, (National Program Director, HHRA) Lynn Flowers, Ph.D., DABT (Associate Director for Health) Vincent Cogliano, Ph.D. (IRIS Program Director

— acting) Samantha Jones, Ph.D. (IRIS Associate Director for Science) Weihsueh A. Chiu, Ph.D. ( Branch Chief ) U.S. EPA /ORD/NCEA Washington, DC

**nternal Review Team** General Toxicology Workgroup Inhalation Workgroup Neurotoxicity Workgroup PBPK Workgroup Reproductive and Developmental U.S. EPA O ffice of R esearch and D evelopment National Center for Environmental Assessment Washington, DC

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Toxicology Workgroup Statistical Workgroup Toxicity Pathways Workgroup



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4 **Reviewers**

This assessment was provided for review to scientists in EPA’s Program and Region Offices. Comments were submitted by:

Office of Children’s Health Protection, Washington, DC Office of Solid Waste and Emergency Response, Washington, DC Region 2, New York, NY Region 8, Denver, CO

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# PREFACE

This Toxicological Review critically reviews the publicly available studies on tert

- butyl alcohol ( t ert

- b utanol ) in order to identify its adverse health effects and to characterize exposure

response relationships

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

. T h e assessment covers an oral RfD, an inhalation RfC, and a cancer weight of evidence descriptor. The Toxicological Reviews for ethyl tert

- bu t yl ether 

### ( ETBE ) and tert

- butanol are being developed s i multaneously because they have a number of overlapping scientific issues:

• t ert

- B utanol is a metabolite of ETBE, so some of the toxicological effects of ETBE may be due

to t ert

- bu tanol. Therefore, data on tert

- butanol may inform hazard identification and dose

response assessment of ETBE, and vice versa.

• Th e scientific literature for chemicals include data on α 2u -globulin- related nephropathy. Therefore, a common approach was employed to evaluate those data as they relate to the mode of action for kidney e f fects.

• A combined PBPK model for ETBE and t ert

- butanol in rats was developed to support the
dose- r esponse assessments for these chemicals.
Th is assessment was conducted in accordance with EPA guidance, which is cited and
summarized in the Preamble to IRI S Toxicological Reviews. This is the first IRIS assessment for this compound. Th e findings of this assessment and related documents produced during its development are available on the IRIS Web site ( http://www.epa.gov/iris ). A public meeting was held in December 2013 to o b tain input on preliminary materials for

tert

- b utanol , i ncluding draft literature searches and associated search strategies, evidence tables, and exposure

- r es ponse 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.r egulations.gov (Docket ID No. EPA

- HQ

- OR D

- 2 013 -0 111 ). In April 2011, the National Research Council (NRC) released its Re v iew of the Environmental Protection Agency’s Draft IRIS Assessment of Formaldehyde

. I n addition to offering comments specifically about EP A ’s draft formaldehyde assessment, the NRC made several recommendations to EPA for improving the development of IRIS assessments. EPA agreed with the recommendations and is implementing them consistent with the Panel’s “Roadmap for Revision,” which vie w ed the full implementation of their recommendations by the IRIS Program as a multi

- y ear process. In response to the NRC’s 2011 recommendations, the IRIS Program has made changes to streamline the assessment development process, improve transparency, and create efficiencies in

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

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t he Program. The NRC has acknowledged EPA’s successes in this area. In May 2014, the NRC 

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38 released their report Review of EPA’s Integrated Risk Information System Process reviewing the IRIS assessment development process and found that EPA has made substantial improvements to the IRIS Program in a short amount of time.
The draft tert

- butanol assessment represents a significant advancement in implementing the NRC recommendations. This assessment is streamlined, and uses tables, f i gures, and appendices to increase transparency and clarity. It is structured to have distinct sections for the literature search and screening strategy, study selection and evaluation, hazard identification, and
dose- r esponse assessment. The assessment includes a comprehensive, systematic, and documented literature search and screening approach, provides the database search strategy in a table (databases, keywords), visually represents the inclusion and exclusion of studies in a flow diagram, and all of the references are integrated within the Health and Environmental Research Online (HERO) database. A study evaluation section provides a systematic review of methodological aspects of epidemiology and experimental animal studies, including study design,

c o nduct, and reporting, that was subsequently taken into consideration in the evaluation and synthesis of data from these studies. The evidence is presented in standardized evidence tables, and exposure

- r esponse arrays. The hazard identification and dose

- response sections include subsections based on organ/system

- sp ecific effects in which the evidence is synthesized within and integrated across all evidence for each target organ/systems.
In the draft tert

- b utanol assessment, the IRIS Program has attempte d to transparently and uniformly identify strengths and limitations that would affect interpretation of results. All animal
studies of tert

- bu tanol that were considered to be of acceptable quality, whether yielding positive, negative, or null results, were considered in assessing the evidence for health effects associated with chronic exposure to t ert

- butanol . These studies were evaluated for aspects of design, conduct, and reporting that could affect the interpretation of results and the overall contribu ti on to the synthesis of evidence for determination of human hazard potential using the study quality considerations outlined in the Preamble. A brief summary of the evaluation is included in the section on methods for study selection and evaluation. Info rmation on study features related to this evaluation is reported in evidence tables and documented in the synthesis of evidence. Discussion of study strengths and limitations (that ultimately supported preferences for the studies and data relied upon) wer e included in the text where relevant. In this assessment, the IRIS Program is using existing guidelines to systematically approach the integration of noncancer human, animal, and mechanistic evidence. In conducting this analysis and developing the synt h esis, the IRIS Program evaluates the data for the: strength of the relationship between the exposure and response and the presence of a dose

- r esponse relationship; specificity of the response to chemical exposure and whether the exposure precedes the effe ct; consistency of the association between the chemical exposure and response; and biological plausibility of the response or effect and its relevance to humans. The IRIS Program uses this

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w eight

- of

- evidence approach to identify the potential human hazard s associated with chemical 

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- butanol assessment provides a streamlined presentation of information, integrated hazard identification of all toxic effects, and derivation of organ/system

- sp ecific reference values. Additionally, consist ent with the goal that assessments should provide a scientifically sound and transparent evaluation of the relevant scientific literature and presentation of the analyses performed, this assessment contains an expanded discussion of study selection and

ev alu ation, as well as increased documentation of key assessment decisions. For additional information about this assessment or for general questions regarding IRIS, please contact EPA’s IRIS Hotline at 202

- 5 66

- 1676 (phone), 202

- 566

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

.

**Chemical Pro**  **perties**

t e rt

- Butanol is a white crystalline solid or colorless liquid (above 77°F) with a camphor

- like
odor t hat is highly flammable ( NIOSH, 2005 ; IPCS, 1987a )

. t ert

- Butanol contains a hydroxyl chemical functional group and is miscible with alcohol, ether, and other organic solvents and
soluble in water ( IPCS, 1987a ) . Select ed chemical and physical properties of t ert

- butanol are presented in Table P -1

.

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**Table P -1**  **. Physicochemical properties and chemical identity of** tert

**- butanol**  

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Chemical name tert

- Butanol HSDB (2007) Synonyms/Trade Names t

- butyl alcohol; tert

- Butanol ; tert

- butyl alcohol; t- Butyl hydroxide; 1,1

- Dimethylethanol; NCI

- C55367; 2

- Methyl -2- propanol; tert iary butanol; Trimethyl carbinol; Trimethyl methanol, t

- butyl alcohol, TBA HSDB (2007) IPCS (1987b)

Chemical Formula C 4 H 10 O HSDB (2007)

CASRN 75

- 65

- 0 HSDB (2007) Molecular weight 74.12 HSDB (2007) Melting point 25.7°C HSDB (2007) Boiling point 82.41°C HSDB (2007)
Vapor pressure 40.7 mm Hg @ 25°C HSDB (2007) Density/Specific Gravity 0.78581 HSDB (2007) Flashpoint 11 °C (closed cup) HSDB (2007) Water solubility at 25° C 1 x 10 6 mg/L HSDB (2007) Octanol/Water Partition Coefficient (Log K OW ) 0.35 HSDB (2007)

Henry’s Law Constant 9.05 x 10 - 6 atm

- m 3 /mole HSDB (2007)

Odor threshold 219 mg/m 3 HSDB (2007)

Conversion factors 1 ppm = 3.031 mg/m 3 1 mg/m 3 = 0.324 ppm IPCS (1987b)

Chemical structure HSDB (2007)

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— DO NOT CITE OR QUOTE **Uses**

t ert

- B utanol is primarily an anthropogenic substance that is produced in large quantities ( HSDB, 2007 ) from a number of precursors, including 1

- butene, isobutylene, acetyl chloride and dimethylzinc, and te rt

- b utyl hydroperoxide. The domestic production volume of t ert

- butanol, including imports, was a pp r oximately four billion pounds in 2012 ( U.S. EPA, 2014 )

. t ert

- B utanol has been used as a fuel oxygenate, an octane booster in unleaded gasoline, and

a denaturant for ethanol. From 1997 to 2005, the a nnual t ert

- butanol volume found in gasoline ranged from approximately 4 million to 6 million gallons. During that time, larger quantities were used to make methyl tert

- bu t yl et her (MTBE) and ETBE. MTBE and ETBE are fuel oxygenates thatToxicological Review of tert

- Butyl Alcohol

w ere used in the U.S. prior to 2007 at levels of more than 2 billion gallons annually. Current use 

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. t e rt

- Butanol has been used for a variety of other purposes including as a dehydrating agent and solvent. As such, it is added to lacquers, paint removers, and nail enamels and polishes.

tert

- B utanol is also used to manufacture methyl methacrylate p lastics and flotation devices.
Cosmetic and food

- r elated uses include the manufacture of flavors and, because of its camphor

- like aroma, it is also used to create artificial musk, fruit essences, and perfume ( HSDB, 2007 ) . It is also used in coatings on metal and paperboard food containers ( Cal/EPA, 1999 ), industria l cleaning compounds

, an d can be used for chemical extractions in pharm aceutical application ( HSDB, 2007 )

.
**Fate and Transport**

Soil The mobility of t e rt

- butanol in soil is expected to be high due its low affinity for soil organic matter. Rainwater or other water percolating through soil is expected to dissolve and transport

most t e rt

- butanol present in soil, potentially lead ing to groundwater contaminati on. Based on

t e rt

- butanol’s vapor pressure, volatilization from soil surfaces is expected to be an important dissipation process ( HSDB, 2007 )

. t ert

- Butanol is a tertiary alcohol and this class of alcohols generally d egrades more slowly in the environment compared to primary ( e.g. , ethanol) or secondary ( e.g

. , isopropanol) alcohols. In anoxic soil conditions, the half

- life of t ert

- butanol is estimated to be on the order of months (approximately 200 days). Microbial degradation rates are increased in soils supplemented with nitrate and sulfate nutrients ( HSDB, 2007 )

.

Water

t e rt

- Butanol is expected to volatilize from water surfaces within 2 to 29 days and does not readily adsorb to suspended solids and sediments in water ( HSDB, 2007 ) . Biodegradation in aerobic water is on the magnitude of weeks to months and in anaerobic aquatic conditions, the biodegradation rate decreases. Bioconcentration of t e rt

- butanol in aquatic organis ms is low ( HSDB, 2007 ).

Air

t e rt

- Butanol exists primarily as a vapor in the ambient atmosphere. Vapor

- phase t ert

butanol is degraded in the atmosphere b y reacting with photochemically

- produced hydroxyl
radicals w ith a half

- life of 14 days ( HSDB, 2007 ).

**Occurrence in the Environment**

The Toxics Release I nventory (TRI) Program National Analysis Report estimated that over one million pounds of t e rt

- butanol has been released into the soil from landfills, land treatment,

2 http://www.ihs.com/products/chemical/planning/ceh/gasoline

- octane

- improvers.aspx

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u nderground injection, surface impoundments, and other land disposal sources. The TRI program 

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- butanol was released into the atmosphere from fugitive emissions and point sources ( U.S. EPA, 2012c ) . In California, air emissions of t ert

- butanol from stationary sources are estimated to be at least 27,000 pounds per year, based on data reported by the state’s Air Toxics Program ( Scorecard, 2014 ) . The TRI National Analysis Report esti mated 7,469 pounds of t e rt

- butanol was released into surface waters from point and nonpoint sources in
2011 ( U.S. EPA, 2012c )

. t e rt

- Butanol has been identified in drinking wells throughout the United States ( HSDB,
2007 ). California’s Geotracker Database 2 lists 3496 detections of t ert

- butanol in groundwater

assoc iated with contaminated sites in that state since 2011. t e rt

- Butanol has also been detected in drinking water wells in the vicinity of landfills ( U.S. EPA, 2012c ). Additionally, t ert

- Butanol leaking
from u nderground storage tanks may be a product of MTBE and ETBE, which can degrade to form

t e rt

- butanol in soils ( HSDB, 2007 ) . The industrial chemical tert

- butyl acetate also can degrade to
form t e rt

- butanol in animals and in the environment.
Ambient outdoor air concentrations of t e rt

- butanol vary, seemingly according to proximity
to urban areas ( HSDB, 2007 ).

**General Population Exposure**

t e rt

- Butanol exposure can occur in many different settings. Contamination resulting from leaking underground storage tanks co uld potentially result in exposure to a large number of people who get their drinking water from wells. Due to its high environmental mobility and resistance to biodegradation, t e rt

- butanol has the potential to contaminate and persist in groundwater and soil; therefore, exposure through ingestion of contaminated drinking water is likely occurring ( HSDB,
2007 )

. Contaminated food can also contribute to t e rt

- butanol ingestion through its use as a coating in metallic and paperboard food containers ( Cal/EPA, 1999 )

. t ert

- Butanol has been detected in food, namely beer and chickpeas, and identified in mother’s milk ( HSDB, 2007 ) . Indirect exposure

to t e rt

- butanol may also occur as a result of ingestion of MTBE or ETBE, as t ert

- butanol is a

me t abolite of these compounds ( NSF International, 2003 )

. Alternate human expo s ure pathways of t ert

- butanol include inhalation and, to a lesser extent, dermal contact. t er t

- Butanol inhalation exposure can occur due to the chemical’s volatility and release from industrial processes, consumer products and contaminated sites ( HSDB, 2007 )

. Dermal contact is a viable route of exposure through handling consumer products containing

t e rt

- butanol ( NSF International, 2003 )

.

2 http://geotracker.waterboards.ca.gov/

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

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**A ssessments by Other**  **Federal, State and Inter**  **national Health Agencies** 

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11

12 Toxicity information on tert

- butanol has been evaluated by the American Conference of Governmental Industrial Hygienists ( ACGIH, 2012 )

, the National Institute f or Occupational Safety
and Health ( NIOSH, 2007 ) , the Occupational Safety and Health Administration ( OSHA, 20 06 ) , and Food and Drug Administration. The results of these assessments are presented in Appendix A of the Supplemental Information

. The California EPA carried out an expedited risk assessment for

t e rt

- butanol in drinking water and calculated a cancer slope factor based on rat kidney tumors observed in the NTP bioassays. It is important to recognize that these earlier assessments were prepared for different purposes using different methods and could consider only the studies that
were available at the ti me.

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# PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS

3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 

## 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 safe ty 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 che micals 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 rela tionships. In terms set forth by the National Research Council ( NRC, 1983 ) , IRIS
assessments cov er 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 agenci es that evaluate risk in specific populations and exposure scenarios. IRIS

ass essments are distinct from and do not address political, economic, and technical considerations that influence the design and selection of risk management alternatives. An IRIS assessment may cover a single chemical, a group of structurally or toxicologically related chemicals, or a complex mixture. These agents may be found in air, water, soil, or sediment. Exceptions are chemicals currently used exclusively as pesticides, ioniz ing and non

- ionizing radiation, and criter ia air pollutants listed under S ection 108 of the Clean Air Act (carbon monoxide, lead, nitrogen oxides, ozone, particulate matter, and sulfur oxides). Periodically, the IRIS Program asks other EPA programs and regions, other federal agencies, state health agencies, and the general public to nominate chemicals and mixtures for future assessment or
reassessment. Agents may be considered for reassessment as significant new studies are published . Selection is based on program and regional office priorities and on availability of adequate information to evaluate the potential for adverse effects. Other agents
may also be assess ed in response to an urgent public health need. 

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oses on

xvii 

## 2. Process for developing and peer

reviewing IRIS assessments
The process for developing IRIS assessments (revised in May 2009 and enhanced in July 2013 ) involves critical analysis of the pertinent studies, opportunities for public input, and multiple levels of scientific review. The EPA revises draft as sessments after each review, and external drafts and comments become part of the public record ( U.S. EPA, 2009 ). Before beg inning an assessment, the IRIS Program discusses the scope with other EPA programs and regions to ensure that the assessment will meet their needs. Then a public meeting on problem formulation invites discussion of the key issues and the studies and analyt ical 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 heal th effects, identify mechanistic events and pathways,

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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 justifie d by the evidence. The peer review draft, written public comments, and peer review report become part of the public record. 

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21 22 23 24 25 26 27 28 29 30 31 32 33 34 35 36 37

38 39 40 41 42 43 44 45 46 47 **St ep**  **5 . Revision of draft Toxicological**  **Review and development of draft IRIS summary**  . The draft assessment is revised to reflect the peer review comments, public comments, and newly published studies that are critical to the conclusions of the assessment. The disposition of peer review comments and public comments becomes part of the public record. **St ep**  **6. Final EPA review and interage ncy**  **science discussion with other federal agencies and the Executive Offices of**  **the President** The draft assessment and summary are revised to address the EPA and interagency comments. The science discussion draft, written interagency comments, and EPA’s response to major comments become part of the public record.
**Step 7. Completion and posting**  . The Toxicological Review and IRIS summary are posted on the IRIS website ( http://www.epa.gov/iris/ ). 

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ses on

2 The remainder of this Preamble addresses step 1, the development of a draft Toxicological Review. IRIS assessments follow standard practices of evidence evaluation and peer review, many of which are discussed in EPA guidelines ( U.S. EPA, 2005a

, b

, 2000b

, 1998b

, 1996
, 1991b

, 1986a
, b ) and other methods ( U.S. EPA, 2012a

, b

, 2011

, 2006a

, b

, 2002

, 1994 ) . Transparent application of scientific judgment is of paramount importance. To provide a harmonized approach across IRIS assessments, this Preamble summarizes concepts from these guidelines and emphasizes principles of general applicability.



## 3. Identifying and selecting pertinent studies

**3.1. Identifying studies**  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://hero.epa.gov/ ). Searches for information on mechanisms of toxicity are inherently specialized and may include

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studies on other agents that act through related mechanisms. Each assessment specifies the search strategies, keywords, and cut

- off dates of its literature searches. The EPA posts the results of the literature search on the IRIS web site and requests information from the public on additional studies and ongoing research. The EPA also considers studies received through th e IRIS Submission Desk and studies (typically unpublished) submitted under the Toxic Substances Control Act or the Federal Insecticide, Fungicide, and Rodenticide Act. Material submitted as Confidential Business Information is considered only if it include s 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 o ther 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 a bility to reflect interactions among components. The literature search seeks, in decreasing order of preference ( U.S. EPA, 2000b, §2.2 ; 1986b, §2.1 ): 

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ses onlo 

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35

36 37 38 39 40

41 42 43 44 – 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 the re 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

- cont rol studies, and some population

- based surveys (for example, NHANES) provide the strongest epidemiologic evidence, especially if they collect information about individual exposures and effects.

– E cological studies (geographic correlation studies) relate exp osures 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.

– C ase reports of high or ac cidental 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 cas e 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 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

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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.

– S tudies 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 ( U.S. EPA, 2006b

, 1998b

, 1996
, 1991 b ). 

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## 4. Evaluating the quality of individua l 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 docu mentation of each study, but not whether the results are positive, negative, or null. The objective is to identify the stronger, more informative studies based on a uniform evaluation of quality characteristics across studies of similar design. 

### 4.1. **Evaluating the quality of epidemiologic studies**  The assessment evaluates design and methodological aspects that can increase or decrease the weight given to each epidemiologic study in the overall evaluation ( U.S. EPA, 2005a

, 1998b

, 1996
, 1994
, 1991b ):

– Documentation of study design, methods, population characteristics, and results.

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

– Ascertainment of exposure to the chemical or mixture.

– A scertainment of disease or health effect.

– D uration of exposure and follow

- up and adequacy for assessing the occurrence of effects.

– Cha racterization of exposure during critical periods.

– S ample s ize and statistical power to detect anticipated effects.

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

– Measurement error (can lead to misclassification of exposure, health outcomes, and other factors) and other types of information bias.

– Potential confounding and other sources of bias addressed in the study design or in the analysis of results. The basis for consideration of confounding is a reasonable expectation that the confounder is related to both exposure and 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

, 1998b

, 1996
, 1991b ).

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### 4.2. **Evaluating the quality of experimental studies** 

The assessm ent evaluates design and methodological aspects that can increase or decrease the weight given to each experimental animal study, in

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

, 1998b

, 1996
, 1991b ) . Research involving human subjects is considered only if conducted according to ethical principles. 

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oses onl – D ocumentation 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.

– S ample 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 o f a response is a trend test or comparison of outcomes in the exposed groups against those of concurrent controls. In some situations, examination of historical control data from the same laboratory within a few years of the study may improve the analysis. For an uncommon effect that is not statistically significant compared with concurrent controls, his torical controls may show that the effect is unlikely to be due to chance. For a response that appears significant against a concurrent control response tha t is unusual, historical control s may offer a different interpretation ( U.S. EPA, 2005a, §2.2.2.1.3 ). For developmental toxicity, reproductive toxicity, neurotoxicity, and cancer there is further guidance on the nuances of evaluating experimental studies of t hese effects ( U.S. EPA, 2005a

, 1998b

, 1996
, 1991b ) . In multi

generation studies, age nts that produce developmental effects at doses that are not toxic to the maternal animal are of special concern. Effects that occur at doses associated with mild maternal toxicity are not assumed to result only from maternal toxicity. Moreover , maternal e ffects may be reversible, while effects on the offspring may be permanent ( U.S. EPA, 1998b, §3.1.2.4.5.4 ; 1991b, §3.1.1.4 ),.



### 4.3. **Reporting study results**

The assessment uses evidence tables to present the design and key results of pertinent studies. There may be separate tables for each site of toxicity or type of study. If a large n umber of studies observe the same effect, the assessment considers the study quality characteristics in this section to identify the strongest studies or types of study. The tables present details from these studies, and the assessment explains the

reasons for not reporting details of other studies or groups of studies that do not add new information. Supplemental information provides references to all studies considered, including those not summarized in the tables. The assessment discusses strengths and l imitations that affect the interpretation of each study. If the interpretation of a study in the assessment differs from that of the study authors, the assessment discusses the basis for the difference. As a check on the selection and evaluation of pertine nt studies, the EPA asks peer reviewers to identify studies that were not adequately considered.

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## 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 pertine nt 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 develo ped frameworks for causal inference, among them the U.S. Surgeon General ( CDC, 2004 ; HEW, 1964 ) , the International Agency for Research on Ca ncer ( IARC, 2006 ) , the Institute of Medicine ( IOM, 2008 ) , and the EPA ( 2010, §1.6 ; 2005a, §2.5 )

. Although developed for different purposes, the frameworks are similar in nature and provide an established structure and language for causal inference. Each considers aspects of an association that suggest causation, discussed by Hill ( Hill, 1965 ) and elaborated by Rothman and Greenland ( Rothman and Greenland, 1998 ) , and U.S. EPA ( 2005a, §2.2.1.7 ; 1994, Appendix C ). 

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40 41 42 43 44 45 **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. **Consistency of association:** An inference of causation is strengthened if elevated risks are observed in independent studies of different populations and exposure scenarios. Reproducibility of findings constitutes one of the strongest arg uments for causation. Discordant results sometimes reflect differences in study design, exposure, or confounding factors. **Specificity of association:** As originally intended , this refers to one cause associated with one effect. Current understanding that many agents cause multiple effects and many effects have multiple causes make this a less informative aspect of causation, unless the effect is rare or unlikely to have multi ple causes. **Temporal relationship:** A causal interpretation requires that exposure precede development of the effect. 

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oses onl **Biologic gradient (exposure**

**- response**  **relationship):** Exposure

- response relationships strongly suggest causation. A monotonic increase is not the only pattern consistent with causation. The presence of an exposure

- response gradient also weighs against bias and confounding as the source of an association. **Biologic plausibility:** An inference of causation is strengthened by data demonstrating plau sible biologic mechanisms, if available. Plausibility may reflect subjective prior beliefs if there is insufficient understanding of the biologic process involved.

**Coherence:** An inference of causation is strengthened by supportive results from animal exper iments, toxicokinetic studies, and short

- term tests. Coherence may also be found in other lines of evidence, such as changing disease patterns in the population. **“Natural experiments”:** A change in exposure that brings about a change in disease frequency pr ovides 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

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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

e ffect is large, if there is evidence of an
exposure

- response relationship, or if an association was observed and the plausible biases would tend to decrease the magnitude of the reported effect. Confidence is decreased for study limitations, inconsistency of results, indirectness of evidence, imprecision, or reporting bias ( Guyatt et al., 2008b ; Guyatt et al., 2008a ).



### 5.2. **Evaluating evidence in humans**
For each effect, the assessment evaluates the evidence from the epidemiologic studies as a whole. The objective is to determine whether a credible association has been observed and, i f 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 obse rved association. 

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39 40 41 42 43 44 45 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 explanations such as chance, bias, and confounding can be ruled out with

r easonable confidence. Su ggestive epidemiologic evidence of an

association consistent with causation: The evidence suggests a causal association but chance, bias, or confounding cannot be ruled out as explaining the association. Inadequate epidemiologic evidence to infer a causal association: The available studies do not permit a conclusion regarding the presence or absence of an association. 

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87 88 89 Epidemiologic evidence consistent with no causal association: S everal adequate studies covering the full range of hum an exposures and considering susceptible populations, and for which alternative explanations such as bias and confounding can be ruled out, are mutually consistent in not finding an association.



### 5.3. **Evaluating evidence in animals**
For each effect, the assessment evaluates the evidence from the animal experiments as a whole to determine the extent to which they indicate a potential for effects in humans. Consistent results across various species and strains increase confidence that similar results would occur in humans. Several concepts discussed by Hill ( Hill, 1965 ) are pertinent to the weight of experimental

resu lts: consistency of response, dose

response relationships, st rength of response, biologic plausibility, and coherence ( U.S. EPA, 2005a, §2. 2.1.7 ; 1994, Appendix C ). In weighing evidence from multiple experiments, U.S. EPA ( 2005a, §2.5 ) distinguishes : Conflicting evidence ( that is, mixed positive and negative results in the same sex and strain using a similar study protocol) from Differing results (that is, positive results and negative results are in different sexes or strains or use different study protocols).

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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

, 1998b

, 1996
, 1991b ) . Accordingly, the assessment determines whe ther 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, chromo some aberrations, or aneuploidy in humans or experimental mammals ( in vivo ) provides the strongest evidence.

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

– Negative results carry less we ight, partly because they cannot exclude the possibility of effects in other tissues ( IARC, 2006 ). 

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39

40 41

42 43

44 45 For germ

- cell mutagenicity, The EPA has defined categories of evidence, ranging from positive results o f human germ

- cell mutagenicity to negative results for all effects of concern ( U.S. EPA, 1986a, §2.3 ).



### 5.4. **Evaluating mechanistic data**

Mechanistic data can be useful in answering several questions.

– The biologic plausibility of a causal interpretation of human studies.

– The generalizability of animal studies to h umans. – Th e susceptibility of particular populations or lifestages. The focus of the analysis is to describe, if possible, mechanistic pathways that lead to a health effect. These pathway s encompass:

– Toxicokinetic processes of absorption, distribution, meta bolism, and elimination that lead to the formation of an active agent and its presence at the site of initial biologic interaction.

– Toxicodynamic processes that lead to a health effect at this or another site (also known as a mode of action ). 

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77 78 79 80 81 82 83 84 85 86 87 88 89 90 91 For each effe ct, the assessment discusses the available information on its m odes of action and associated key events ( key events being empirically observable, necessary precursor steps or biologic markers of such steps; mode of action being a series of key events involving interaction with cells, operational and anatomic changes, and resulting in disease). Pertinent information may also come from studies of metabolites or of compounds that are structurally similar or tha t act through s imilar mechanisms. Information on mode of action is not required for a conclusion that the agent is causally related to an effect ( U.S. EPA, 2005a, §2.5 ). The assessment addresses several questions about each hypothesized mode of action ( U.S. EPA, 2005a, §2.4.3.4 ). 1) **Is the hypothesized mode of action sufficiently supported in test animal**  **s?** Strong support for a key event being necessary to a mode of action can come from experimental challenge to the hypothesized mode of action, in which studies that suppress a key event observe suppression of the effect. Support for a mode of action is mea ningfully strengthened by consistent results in different experimental models, much more so than by replicate experiments in the same model. The assessment may consider various aspects of causation in addressing this question.

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2) **Is the hypothesized mode of a ction**  **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 mod es 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
assessment s 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. 

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27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 The assessment discusses the likelihood that an agent operates through multiple mode s o f action. An uneven level of support for different modes of action can reflect disproportionate resou rces spent investigating them ( U.S. EPA, 2005a, §2.4.3.3 )

. It should be noted that in clinical reviews, the credibility of a series of studies is reduced if evidence is limited to studies fund ed by one interested sector ( Guyatt et al., 2008a ). For cancer, the assessment evaluates evidence of a mutagenic mode of action to guide extrapolation to lower doses and consideration of susceptible lifestages. Key data include the ability of the agent or a metabol ite to react with or bind to DNA, positive results in multiple test systems, or similar properties and structure

- activity relationships to mutage nic carcinogens ( U.S. EPA, 2005a ,§2.3.5 ). 

### 5.5. **Characterizing the overall weight of the evidence**  After evaluating the human, animal, and mechanistic evidence pertinent to an effect, the assessme nt answers the question: Does the agent cause the adverse effect? ( NRC, 2009

, 1983 ) . In doing this, the assessment develops a narrative that integrates the evidence pertinent to causation. To provide clarity and consistency, the narrative includes a standard hazard descriptor. For example, the f ollowing standard descriptors co mbine epidemiologic, experimental, and mechanistic evidence of carcinoge nicity ( U.S. EPA, 2005a, §2.5 ). Carcinogenic to humans: There is convincing epidemiologic evidence of a causal association (that is, there is reasonable confidence that the association cannot be fully explained by chance, bias, or confounding); or there is strong human evidence of cancer or its precursors, extensive animal evidence, identification of key precursor events in animals, and strong evidence that they are anticipated to occur in humans.

L ikely to be carcinogenic to humans: The evidence demonstrates a potential hazard to humans but does not meet the criteria for carcinogenic . There may be a plausible association in humans, multiple positive results in animals, or a combination of human, animal, or other experimental evidence. Suggestive evidence of carcinogenic potential: T he evidence raises concern for effects in humans but is not sufficient for a stronger conclusion. This descriptor covers a range of evidence, from a positive result in the only available study to a single positive result in an extensive database that includes negative results in other species. Inadequate information to assess carcinogenic potential: No other descriptors apply. Conflicting evidence can be classified as inade quate information if

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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 hu mans: 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 eviden c e that carcinogenic effects differ by dose range or exposure ro ute ( U.S. EPA, 2005a, §2.5 ). Another example of sta ndard descriptors

come s from the EPA’s Integrated Science Assessments, which evaluate causation for the effects of the criteria polluta nts in ambient air ( U.S. EPA, 2010, §1.6 ). Causal relationship: Sufficient evidence to conclude that there is a causal relationship. Observational studies cannot be explained by plausible alternatives, or they are supported by other lines of evidence, for example, animal studies or mechanistic information. 

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45 46 Li kely 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 o r 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 sho ws 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 exposur e 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. 

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## 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 experimen tal 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:

– E pidemiologic studies are preferred over animal studies, i f 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.

– S tudies by a route of hum an environmental exposure are preferred, although a validated 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

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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

- mo notonic exposure

response relationships are not necessarily excluded from the analysis. A diminished effect at higher exposure levels may be satisfactorily expla ined 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 t his 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 n ot adequately considered. 

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## 7. Deriving toxicity values

**7.1. General framework for dose** 

**** **response analysis**

Th e EPA uses a two

- step approach that distinguishes analysis of the observed dose

response data from inferenc es about lower doses ( U.S. EPA, 2005a, §3 ). Within the observed range, the preferred approach is to use modeling to incorporate a wide range of data into the analysis. The modeling yields a point of departure (an exposure level near the lower end of the observed range, without significant extrapolation to lower doses) (S ections 7.2

- 7.3). Ex trapolation to lower doses considers what is known about the modes of action for each effect ( S ections 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

o f uncertainty and variability (S ection 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 dos eresponse analysis is conducted for one well

studied index chemical in the group, then the potencies of other members are expressed in relati ve terms based on relative toxic effects, relative absorption or metabolic rates, quantitative structure

- activity relationships, or receptor binding characteristics ( U.S. EPA, 2005a, §3.2.6 ; 2000b, §4.4 ). 

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11 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** 

T he 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 cl ose, reliable surrogate measure. The active agent may be the administered chemical or a metabolite. Confidence in the use of a toxicokinetic model depends on the robustness of its validation process and on the results of sensitivity analyses ( U.S. EPA, 2006a ; 2005a, §3.1 ; 1994, §4.3 ). Because toxicokinetic modeling can require many parameters and more data than are typically available, the EPA has developed standard approaches that can be applied to typic al data sets. These standard approaches also facilitate comparison across exposure patterns and species.

– I ntermittent study exposures are standardized to a daily average over

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over the lifespan. Exposures during a critical period, ho wever, are not averaged over a longer duration ( U.S. EPA, 2005a, §3.1.1 ; 1991b, §3.2 ).

– Doses are standardized to equivalent human terms to facilitate comparison of results from different species.

– O ral doses are scaled allometrically using mg/kg 3/4 -d ay as the equivalent dose metric across species. Allometric scalin g pertains to equivalence across species, not across lifestages, and is not used to scale doses from adult humans or mature animals to infants

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

– Inhalation exposures are scaled using dosimetry models that apply species

specific physiologic and anatomic factors and consider whether the ef

fect occurs at the site of first contact or after systemic circulation ( U.S. EPA, 2012a ; 1994, §3 ).

It can be informative to convert doses across exposure routes. If this is done, the assessment descri bes the underlying data, algorithms, and assumptions ( U.S. EPA, 2005a, §3.1.4 ). In the absence of study

- specific data on, for example, intake rates or body weight, t he EPA has developed recommended values for use in dose

- response analysis ( U.S. EPA, 1988 ). 

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36 37

38 39 40 41 42 43 44 45 46 **7.3. Modeling response in the range of**  **observation** Toxicodynamic (“biologically based”) modeling can incorporate data on biologic processes leading to an effe ct. 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 lowe r doses, critical biologic parameters should be measured from laboratory studies, not by model fitting. Confide nce in the use of a toxicodynamic model depends on the robustness of its validation process and on the results of sensitivity analyses. P eer revi ew of the scientific basis and performance of a model is essential ( U.S. EPA, 2005a, §3.2.2 ). Because toxicodynamic modeling can require many parameters and more knowledge and data than are typically avai lable, the EPA has developed a standard set of empirical (“curve

- fitting”) models ( http://www.epa.gov/ncea/bmds/ ) that can be applied to typical data sets, including those that are nonlinear. The EPA has also developed guidance on modeling dose

response data, assessing model fit, selecting 

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ses onl suitable model s, and reporting modeling results ( U.S. EPA, 2012b ) . Additional judgment or alternative analyses are used if the procedure fails to yield reliable results, for example, if the fit is poor, modeling may be restricted to the lower doses, especially if there is competing toxicity at higher doses ( U.S. EPA, 2005a, §3.2.3 ). M odeling i s used to derive a point of departure ( U.S. EPA, 2012b ; 2005a, §3.2.4 ). (See S ection 7.6 for alternatives if a point of departure cannot be derived by modeling.) :

– If li near extrapolation is used, selection of a response level
correspo nding 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.

– F or dichotomous data, a response level of 10% extra risk is generally used for minimally adverse effects, 5% or lower for more severe effects.

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– 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 mor e 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 belo w the observed data. Low

- dose
extr apolation, typically used for cancer data

, considers what is known about modes of action ( U.S. EPA, 2005a, §3.3.1 and §3.3.2 ). 1) If a biologically based model has been developed and validated for the agent, extrapolation may use the fitted mod el 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: 

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10 11 12

13 14

15 16 17 18 19 20

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27 28 29 30

31 32 33 34 35 36 37

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41 42 43 44 45 – 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 ar e 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

- r esponse curve at lower doses or concentration s, respectively. 3) N onlinear models are used for extrapolation if there are sufficient data to ascertain the mode of action and to conclude that it is not linear at lower doses, and the agent does not demonstrate mutagenic or other activity consistent with l inearity 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. 

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ses on If linear extrapolation is u sed, 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 assessme nt 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 S ection 6, and the possibility of basing a more robust result on multiple data set s.



### 7.5. **Considering susceptible populations and lifestages** 
The assessment analyzes the available information on populations and lifestages that may be particularly susceptible to each effect. A tiered approach is used ( U.S. EPA, 2005a, §3.5 ). 1) If an epidemiologic or experimental study reports quantitative results for a susceptible population or lifestage, these data are analyzed to derive separate toxicity values for susceptible individuals.

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2) If data on risk

- relat ed 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 epidemi ologic studies and cancer bioassays do not include early

- life exposure. T o address the potential for early- life susceptibility, the EPA

recom mends ( U.S. EPA, 2005b, §5 ):

– 10

- fold adjustment for exposures before age 2 years.

– 3

- fold adjustment for exposures between age s 2 and 16 years.



### 7.6. **Reference values and uncertainty factors**  1 2 3 4 5 6

7 8 9 10 11 12 13 14 15 16 17 18

19 20 21 22

23 24

25 26 27 28 29 30 31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46 47 An oral reference dose or an inhalation reference concentration is an estimate of an
exposur e (including in susceptible subgroups) that is likely to b e without an appreciable risk of adverse health effects over a lifetime ( U.S. EPA, 2002, §4.2 ). Reference values are typically calculated for effects other than cancer and for suspected carcinogens if a well characterized mode of action indicates that a necessary key event does not occur below a specific dose. Reference values provide no information about risks at higher exposure levels. The assessment characterizes effects that form the basis for reference values as adverse, considered to be adverse, or a precursor to an a dverse effect. For developmental toxicity, reproductive toxicity, and neurotoxicity there is guidance on adverse effects and their biologic markers ( U.S. EPA, 1998b

, 1996
, 1991b ). 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 dep arture. 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 uncert ainty.

**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 gener ally 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 susc eptible and non

susceptible individuals) ( U.S. EPA, 2002, §4.4.5 ; 1998b, §4.2 ; 1996, §4 ; 1994, §4.3.9.1 ; 1991b, §3.4 ).

**Animal**

**- to** 

**- human extrapolation.**  If animal results ar e 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 10 1/2 (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 diff erences across species, this factor is not used. In most other cases, a factor of 10 is applied ( U.S. EPA, 2011 ; 2002, §4.4.5 ; 1998b, §4.2 ; 1996, §4 ; 1994, §4.3.9.1 ; 1991b, §3.4 ).

**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 

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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 ma y be used, depending on the magnitude and nature of t he res ponse and the shape of the dose

response curve ( U.S. EPA, 2002, §4.4.5 ; 1998b, §4.2 ; 1996, §4 ; 1994, §4.3.9.1 ; 1991b, §3.4 ).

**Su bchronic** 

**- 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 subch ronic 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
stu dies and the nature of the response ( U.S. EPA, 2002, §4.4.5 ; 1998b, §4.2 ; 1994, §4.3.9.1 ). **Incomplete database.** If an incomplete database raises concern that further studies might identify a more sensitive effect, organ system, or lifestage, the 

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44 45 46 47 48 as sess ment may apply a database uncertainty factor ( U.S. EPA, 2002, §4.4.5 ; 1998b, §4.2 ; 1996 , §4 ; 1994, §4.3.9.1 ; 1991b, §3.4 ). The size of the factor depends on the nature of the database deficiency. For example, the EPA typically follows the suggestion that a factor of 10 be applied if both a prenatal toxicity study and a two

- generation reproduction s tudy are missing and a factor of 10 1/2 if either is missing ( U.S. EPA, 2002, §4.4.5 ). In this way, the assessment derives candidate values for each suitable data set and effect that is credibly associated with the agent. These results are arrayed, using

common d ose metrics, to sh ow where effects occur across a range of exposures ( U.S. EPA, 1994, §4.3.9 ). The assessment derives or selects an
organ - or system

- specific reference value for each organ or system affected by the agent. The assessment explain s 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 facilita te subsequent cumulative risk assessments that consider the combined effect of multiple agents acting at a common site or through common mechanisms ( NRC, 2009 ). The assessment then selects an overall reference dose and an overall reference concentration for the agent to represent 

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15 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 characteri ze 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 comp leteness of the database, with more weight given to the latter. The level of confidence is increased for reference values based on human data su pported by animal data ( U.S. EPA, 1994, §4.3.9.2 ). **High confidence:** The reference value is not likely to change with further testing, except for mechanistic studies that might affect the interpretation of prior test results.

**Medium confidence:** This is a matter of judgment, between high and low confidence.

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**Low confidence:** The reference value is 

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2

29

8 e specially vulnerable to change with further testing.

These criteria are consistent with guid e lines for systematic reviews that evaluate the quality of evidence. These also focus on whether further research would be likely to change confidence in the estima te of effect ( Guyatt et al., 2008b ). All asses sments discuss the significant uncertainties encountered in the analysis. The EPA provides guidance on characterization of uncertainty ( U.S. EPA, 2005a, §3.6 ). For example, the discussion 

### 14 15 16 17 18 19 20 21 22 23 24 25 26 27 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).

Augus t 2013



# E XECUTIVE SUMMARY

Occurrence and Health Effects 

### 30

31 32 33 34 35 36 37 38 39 40 41 42 43 44 45 46

47

48
49
50
51
52
53
54
55 tert

- Butanol does not occur naturally, but it is produced by humans for multiple purposes

, such as a solvent for paints, a denaturant for ethanol and several other alcohols, an octane booster in gasoline, a dehydrating agent, and the manufacture of flotation agents, fruit essences, and perfumes

. tert

- Butanol is also a primary metabolite of methyl tert

- butyl ether (MTBE) and ethyl tert

- butyl ether (ETBE)

. Exposure to tert

- butanol primarily occurs through breathing air containing
tert

- butanol vapors , as well as consuming contaminated water (or breast milk) or foods. Exposure may also occur through direct skin contact. Animal studies demonstrate that chronic oral exposure to tert

- butanol is associated with kidney and thyroid effects

. D evelopmental effects (e.g., reduced fetal viability) have been observed in short

- term exposure to high levels of tert

- butanol (via oral or inhalation exposure) in animals. No chronic inhalation exposure studies have been conducted. T here is suggest ive evidence that tert

- butanol is c arcinogenic to humans based on predominantly benign renal tumors in male rats and benign thyroid t umors in female mice

.

**Effects Other Than Cancer Observed Following Oral Exposure**  EPA identified kidney effects as a hazard of tert

- buta n ol exposur e , with k idney toxicity observed after oral exposure in two strains of rats

, one s tr ain of mice

, and in both sexes

. In mice
, the only kidney effect observed was an increase in kidney weight (absolute and/or relative) in both

sexes of mice in the 13

- week s tudy , but no treatment

- related histopathological lesions were reported in the kidneys of mice at 13 weeks or 2 year s . Ab s olute and relative kidney weights were
increased in both male and female r at s after 13 week s and 15 months of treatment.

H istopathological examination also indicated kidney toxicity in both male and female rat s, w i th
increased incidence of ne phr o pathy after 13 weeks of oral exposure and transitional epithelium

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hyperplasia observed after 2 years of oral exposure. Additionally, increased suppurative 

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4

5

6
7
8
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10
11
12
13
14
15
16
17
18

19 inflammation was noted in fema les after 2 years of oral exposure . Mode of action analysis
deter mined that male rat kidney effects were not mediated by α 2u -globulin, and these effects are
concluded to be relevant for human health hazard assessment.

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

Kidn ey t oxicity, represented by kidney transi tio n al epithelial hyperplasia , was chosen as the basis for the proposed overall oral reference dose (RfD) (see T ab le ES

- 1), as it was the only

no n cancer endpoint for which there is credible evidence of an association with tert

- buta no l exposure. The chronic stud y by NTP (1995) and the observed kidney effects were used to derive
the R fD. The endpoint of transitional epithelial hyperplasia was selected as the critical effect due to

its con s istency in b o th sexes, its specific ity and its sensitiv ity as an indicator of kidney toxicity , and
the observed dose

- respo n se relationship of effects across dose groups. Benchmark dose (BMD) modeling was utilized to derive the BMDL 10% of 16 mg/kg -day . The BMDL was converted to a human equivalent dose using body weight 3/4 scaling

, and this value of 3.84 mg/kg -d ay was used as the point of departure (POD) for RfD derivation ( U.S. EPA, 2011 ). The proposed overall RfD was calculated by dividing the POD for kid n ey transitional epithelial hyperplasia by a co mp osite uncertainty fa ctor (UF) of 3 0 to account for the extrapolation from animals to humans ( 3 ) and f or in terindividual differences in human susceptibility (10).

**Table ES -1** 

**. Summa**  **r y of r**  **eference d**  **ose (RfD) derivation** 

Effect Basis RfD (mg/kg

- day) Exposure description Confidence

Kidney toxicity Increased incidence of kidney transitional epithelial hyperplasia 1 × 10 - 1 Chronic HIGH

Proposed overall RfD Increased incidence of kidney transitional epithelial hyperplasia 1 × 10 -1 Chronic HIGH



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21

22
23
24
25
26
27 **Effects Other Than Cancer Observed Following Inhalation Exposure**  EPA identified kidney effects as a human hazard of tert

- butan o l exposure. Both absolute and relative kidney weights were increased in male and female rats. There was an increase in nephropa thy severity in male rats

, wh ic h supported the increase in kidney weights. No available human studies evaluated the effects of inhalation exposure. Mode o f action analysis determined that male rat kidney effects were not mediated by α 2u -globulin, and these effects are concluded to
be relevant for human health hazard assessment.

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**Inhalation Reference Concentration (Rf**  **C ) for Effects Other Than Cancer**  

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12 Kidney toxicity, represented by transitional epithelial hyperplasia, was chosen as the basis for the proposed inh alat ion reference concentration (RfC) (see Table ES

- 2), as it was the only noncancer endpoint for which there is credible evidence of an association with t ert -b ut a nol exposure. The chronic ora l exposure study in rats ( NTP, 1995 ) was used to derive the overall RfC.
A PBPK model for tert

- buta no l in rats was developed internally, and route

- to

- route e xtrapolation was used to derive equivalent inhalation PODs. The P OD adjusted for the human equivalent concentration (HEC) was 26.1 mg/m 3 and based on transitional epithelial hyperplasia

. The RfC was calculated by dividing the POD by a composite UF of 3 0 to ac c ount for toxicodynamic differences betw een animals and humans (3) and interindividual differences in human susceptibility (10)

.

**Table ES**  **-2**

**. S**  **umma ry of r**  **eference concentration**  **(Rf C)**  **derivation**

Effect Basis Rf C (mg/ m 3 ) Exposure description Confidence

Kidney toxicity Increased incidence of kidney transitional epithelial hyperplasia 9 × 10 - 1 Chronic HIGH

Proposed overall Rf C Increased incidence of kidney transitional epithelial hyperplasia 9 × 10 -1 Chronic HIGH



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29 **Evidence for Human Carcinogenicity**

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

tert

- butanol pr o vides “ suggestive evidence of carcinogenic potential .” H uman data are not available

to assess the carcinogenic potential of tert

- but a nol

. In 2

- year stud ies in F344 rats and B6C3F 1 mice
,

m ale rats exhibited dose

- relat ed increases in renal tubule adenoma and combined renal tubule
adenoma or carcinoma. Although dat a support α 2u - globulin deposition in the kidney of male rats, there is insufficient evidence to support this as the only or prima ry m e chanism for renal tumor development in male rats . T herefo r e, the renal tumors are considered relevant to humans. However, the observed renal tumors were predominantly benign, only occurred in a single sex/species combination, and were not observed in stud ie s that exposed the same strain of rat to ETBE, which is rapidly metabolized to tert

- buta no l. In addition, a statistically significant increase in the incidence of thyroid follicular cell adenoma was observed in a 2

- year d rinking water study in
fema le mice ( NTP, 1995 )

. These tumors were all benign and only a single sex/species combination
was aff ected. Th er e are no studies examining the carcinogenic potential of tert

- butanol after inhalation exposure in animals. H owe v er, internal tumors developed after oral exposure and may occur regardless of exposure route

, as b l ood concentrations were found to be similar after oral or

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inhalation exposures

. Genotoxicity data for tert

- butanol are inconclusive. tert

- Butanol was negative 

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37 in a variety of genotoxicity assays in different cell systems including gene mutations, sister chromatid exchanges, micronucleus formation and chromosomal aberrations . However, DNA adducts in male Kunming mice and DNA damage in human HL

- 60 leukemia cells have been
observed. Overall , the cancer descriptor “ suggestive evidence of carcinogenic potential ” is plausible , as some concern is raised by the positive evidence of predominantly benign renal tumors in male rats and benign thyroid tumors in female mice

.

**Quantitative Es**  **t imate**  **of Carcinogenic Risk**  **from Oral Exposure**  Lifetime oral exposure to tert

- but a nol has been associated with i ncreased renal tubule
adenomas and car cin oma in male F344 rats , increased thyroid follicular cell adenomas in female
B6C3F 1 mice, and increased thyroid follicular cell adenomas and carcinomas in male B6C3F 1 mice

. The NTP (1995) study in rats and mice was the only available study for dose

- response analys is. The study included histological examinations for tumors in many different tissues, contained three exposure levels and controls, contained adequate numbers of animals per dose group (~50/sex/group), treated animals for up to 2 years, and included detai l e d reporting of methods
and results. Although tert

- b u tanol was considered to have “ suggestive evidence of carcinogenic potential ,” EPA concluded that the main study was well

- c o nducted and quantitative analysis may be useful for providing a sense of the m agn i tude of potential carcinogenic risk. For renal tumors, t wo slope factors were derived for this endpoint from the NTP (1995) bioassay : one based on the original reported incidences and one based on the Hard et al. (2011) reanalysis. The two estimates
diffe r ed by less than 20%, and round ed t o the same number at one significant figure . However, the Hard et al. (2011) reanalysis is considered preferable, as it is based on a Patholo gy Working Group ( PWG ) an alysis. A slope factor was also derived for thyroid tumors in female mice. The modeled

tert

- butanol P O Ds were scaled to HEDs according to EPA guidance by converting the BMDL 10 on the basis of (body weight) 3/4 scaling ( U.S. EPA, 2011

, 2005a )

. Using linear extrapolation from the

BMDL 10 , a hum an equivalent oral slope factor w as derived (slope factor = 0.1/BMDL 10 ). The more sensitive endpoint of renal tumors was used bec au se there is no data to support n either renal nor thyroid tumors most relevant to humans. The oral slope fa c tor of **1 × 10**  **-2 per mg/kg** 

**- day**  , based
on the renal tubule tumor response in male F 344 rats

.

**Quantitative Est**  **i mate**  **of Carcinogenic Risk**  **from Inhalation Exposure** 

No chronic i n halation exposure stud ies to tert

- butanol are available . However, through the oral route of exposure, l if et ime exposure has bee n associated with increased renal tubule adenomas
and carcinoma i n male F344 rats, increased thyroid follicular cell adenomas in female B6C3F 1 mice,
and increased th y roid follicular cell adenomas and carcinomas in male B6C3F 1 mice . The NTP (1995) study in rats and mice was the only available study for dose

- response analysis. The study included histological examinations for tumors in many different tissues, contained three exposure

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levels and controls, contained adequate numbers of animals per dose group (~50/sex/group), 

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37 treated animals for up to 2 years, and included detailed reporting of methods and results. Although t ert

- b utanol was considered to have “suggestive evidence of carcinogenic potential,” EPA c oncluded that the main study was well

- conducted and quantitative analysis may be useful for providing a sense of the magni tu d e of potential carcinogenic risk. Since the available
evidence for tert

- but a nol carcinogenicity is from a 2 year oral exposure, route

- to

- route extrapolation of the oral BMDL was performed to derive an inhalation equivalent BMCL

. Th e BMCL
was then conve rt ed to a human equivalent concentration (HEC) a ccording to the RfC guidelines ( U.S. EPA, 1994 ) by multiplying the BMCL by the blood:gas partition coefficient ratio. Using linear extrapolation from the resulting BM C L 10

- HEC , a hum an equivalent inhalation unit risk w as derived ( inhalation unit risk = 0 .1/BM CL 10

- HEC ). Extrapolation from the oral study results for renal tubule adenoma or carcinoma in male F44 rats gives a unit risk of **2 × 10**  **-3 per mg/**  **m 3**  , associated with lifetime inhalation exposure to tert

- but a nol

.
**Susceptible Populations and Lifestages fo**  **r Cancer and Noncancer**  No data were identified to indicate susceptible populations or lifestages.

**Key Issues Addressed in Assessment** 

D ue to the observation of kidney tumors and noncancer toxicity following chronic exposure

to tert

- b utanol, an evaluation of whether t ert

- butanol cause d α 2u - globulin nephropathy was performed. T he p resence of α 2u - globulin in the hyaline droplets was confir med in male rats by α 2u immunohistochemical staining . Lin ear mineralization and tubular hyperplasia were reported in male rats, though only in the chronic study. Ot h er subsequent steps in the pathological sequence, including necrosis, exfoliation, and gr anu lar casts, were either absent or not consistently observed
across subchronic or chronic studies. N on e o f the observed effects occurred in female rats or in
either sex of mice. B e cause the available data support s the occurrence of at least two of the

s ubsequent steps in the pathological sequence, these data are sufficient to conclude that

α 2u - globulin nephropathy is occurring in the kidney of male rats following tert

- butanol exposure. Thus, the noncancer lesions associated with α 2u - globulin nephropathy are not considered relevant
to humans. However, tumors develop at doses lower than some precursors of α 2u -globulin nephropathy, such as granular casts and tubular hyperplasia ( Hard et al., 2011 ; NTP, 1995 ).
Therefor e, there is insufficient evidence to support a conclusion that α 2u - globulin nephropathy is the sole or primary contributor to renal tumor development. Be c ause carcinogenic processes other
than α 2u - globulin nephropathy cannot be ruled out, the renal tumors are considered relevant to
humans. In addition, s o m e o f the observed renal lesions in rats following exposure to tert

- butanol are effects commonly associated with chronic progressive nephropathy (CPN), an age

- relat ed renal disease of laboratory rodents t ha t occurs spontaneously. While it has been argued that CPN in rats

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is not relevant to humans, it is acknowledged that the mechanism regulating CPN in rats is not 

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11 understood. Moreover, no key events for the exacerbation of CPN have been identified, so no mode of action analysis can be performed. Therefore, kidney effects from tert

- butanol exposure
associated with CPN are considered relevant to humans. Sufficient data were available to develop a PBPK model in rats for both oral and inhalation

e x posure in order to perform route

- to

- rout e extrapolation, so rat studies from both routes of exposure were considered for dose

- res p onse analysis. The only endpoin t available from the
subchronic inh a lation study ( NTP, 1997 ) was increased kidney weights

, which is a less

- specific endpoint compared to other endpoints available fo r analysis from the oral study ( NTP, 1995 )

. In regards to the carcinogenic effects, the 2

- ye a r oral study ( NTP, 1995 ) was the only study to evaluate lifetime carcinogenic effects and was selected for route

- to

- ro u te extrapolation.

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38 

# LITERATURE SEARCH STRATEGY | STUDY SELECTION AND EVALUATION

A literature search and screening strategy were used to identify literature characterizing
the health effects of tert

- b u t anol

. This strategy consisted of a broad search of online scientific databases and other sources in order to identify all potentially pertinent studies. In subsequent steps, references were screened to exclude papers not pertinent to an assessment of the health
effects of tert

- b uta n ol , and remaining references were sorted into categories for further evaluation. This section describes the literature search and screening strategy in detail.
The chemical

- sp ecif i c search was conducted in four online scientifi c databases , including PubMed, Toxline, Web of Science, and TSCATS thr oug h Ap ril 2014 , using the keywords and limits
described in Table LS

- 1. T h e overall literature search approach is shown graphically in Figure LS -1

. An additional 7 cit at ions were obtained using additional search strategies described in Table LS

- 2. After electronically eliminating duplicates from the citations retrieved through these databases, 2,532 unique citations were identified.
Th e r esulting 2,532 cita t ions were screened into categories as presented in Figure LS

- 1 using the title, abstract, and/or full text for pertinence to examin e the h e alth effects of tert

- butanol

exposure.

• 12 references were identified as potential “S o ur ces of H ealth E ffects D ata ” and we re considered for data extraction to evidence tables and exposure

- res p onse arrays.

• 19 6 references were identified as “ S upporting S tudies ;” these included 3 9 studies describing physiologically

- bas e d pharmacokinetic (PBPK) models and other toxicokinetic in formation, 70 st udies providing genotoxicity and ot her mechanistic information

, 1 human case report, 7 3 n o t relevant exposure paradigm s (including acute, dermal, eye irritation, and injection studies)

, 6 pr eliminary toxicity studies, and 7 physical dependency studies

. While still considered sources of health effects information, studies investigating the effects of acute and direct chemical exposures are generally less pertinent for characterizing health hazards associated w ith chronic oral and inhalation exposure. Therefore, information from these studies was not considered for extraction into evidence tables. Never t heless, these studies were still evaluated as possible sources of supporting health effects information.

• 63 references w er e identified as secondary sources of health effects information (e.g., reviews and other agency assessments); these references were kept as additional resources for development of the Toxicological Review.

• 2,2 61 r eferences were identified as n ot being pertinent to an evaluation of the health effects of tert

- butanol an d were excluded from further consideration (see Figure LS

- 1 for exclusion categories).

The complete list of references and t he sorting of these materials can be found on the

HE RO website at http://hero.epa.gov

.
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**Selection of Critical Studies for Inclusion in Evidence Tables** 

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37 E ach study retained after the literature search and screen wa s evaluated for aspects of its design or conduct that could affect the interpretation of results and the overall contribution to the evidence for determination of hazard potential. Some general questions that were co n sid ered in evaluating e xp er imental animal studies are presented in Table LS

- 3. Much of the key information for conducting this evaluation can generally be found in the study’s methods section and in how the study results are reported. Importantly, the evaluation at this stage does not consider t h e direction or magnitude of any reported effects.
To facilitate th is evalu at ion , evidence tables we re constructed that systematically summarize the important information from each study in a standardized tabular format as recommended by the NRC (2011)

. T welve studies identified as “Sources of Health Effects Data ”

were c onsidered for extraction i n to evidence tables for hazard identification in Chapter 1. Initial
review of studies fo un d two studies to be publications of the NTP (1995) data prior to the release of
the final N TP report ( Cirvello et al., 1995 ; Lindamood et al., 1992 ). One publication in the “Supporting Studies” category also was based on data from the NTP report ( Takahashi et al., 1993 )

. T here were differences between the published reports and the final N T P report; therefore, the
finalized NTP (1995) report was included in evidence tables. Data from the remaining 10 studies in the “Sources of Health Effects Data” category were extracted into evidence tables. Supporting s tu d ie s that contain pertinent information for the toxicological review and
augment hazard identification co n clusions, such as genotoxic and mechanistic studies, studies describing the kinetics and disposition of tert

- but a nol absorption and metabolism, pilot studies,

s hort t e rm or acute studies, w ere not included in the evidence tables. Such s upporting studies may
be discussed in the narrative sections of Chapter 1, or presented in Appendices, if they provide
additional or c orroborating information.

**Database E**  **va lu**  **ation**

The database for tert

- but a nol is comprised of animal toxicity studies containing one 2

- year bioassay that employ s or a l exposures in rats and mice ; two oral subchronic studies in rats and one

in mice ; one inhalation subchronic study in rats and mice ; a re

- evaluation of the NTP (1995) rat
data ; two oral developmental studies ; tw o inhalation developmental studies ; and one one

generation reproductive study that also evaluate s ot h er systemic effects . Several acute and short
term studies (i nc luding an 18

- day inhalation study and a 14

- day study by NTP) using oral and
inhalation e xp o sures were pe rformed mostly in rats, but were grouped as supporting studies since
the database of chronic and subchronic rat studies was considered sufficient. No cohort studies

,

case

- co ntrol studies, o r ecological studies exist in the published literature. There was one case report available. Health effect studies of gasoline and tert

- but a nol mixtures were not considered pertinent to the assessment since the separate effects of the ga s oline component s could not be determined; thus, these studies were excluded d u ring the manual screen.

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The “Sources of Health Effects Data ” were comprised entirely of studies performed in rats 

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10 and mice with drinking water, oral gavage, and inhalation exposures to tert

- butanol

. T hese 12 sources were conducted according to OECD Good Laboratory Practice (GLP) guidelines, presented extensive histopathological data, and /or clear ly present ed their methodology; thus, these are considered high quality. Preliminary, acute

, and short

- term studies contained information that supported and did not differ qualitatively from the results of the ≥3 0 day exposure studies; thus,
these studies are not included in the evidence tables. Some of these shorter duration studies are presented in the text of the Toxicological Review and are used in section s such as “Mechanistic Evidence” to augment the discussion. A more detailed discussion of methodological concerns that were identified will precede each endpoint evaluated in the hazard identification section.

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### 1

2 **Figure LS -1.**  **Study selection strategy.**Toxicological Review of tert

- Butyl Alcohol

**Table LS -1**  **. Details of the search strategy employed for** tert

**- butanol**  

### 1

Database (Search Date) Keywords Limits PubMed (12/20/2012) (4/17/2014) tert

- butanol OR 75

- 65

- 0[rn] OR “t

butyl hydroxide” OR “2

- methyl -2- propanol” OR “trimethyl carbinol” OR “t

- butyl alcohol” OR tert

- butanol OR

" tert

- butyl alcohol" OR tert

- butyl alcohol[mesh] None

Web of Science (12/20/2012) (4/17/2014) Topic = ( tert

- butanol OR 75

- 65

- 0 OR "t

- butyl hydroxide" OR "2

- methyl -2- propanol" OR "trimethyl carbin ol" OR "t

- butyl alcohol" OR " tert

- butanol " OR " tert

- butyl alcohol") Refined by: Research Areas = (cell biology OR respiratory system OR microscopy OR biochemistry molecular biology OR gastroenterology hepatology OR public environmental occupational health OR oncology OR physiology OR cardiovascular system cardiology or tox icology OR life sciences biomedicine other topics OR hematology OR pathology OR neurosciences neurology OR developmental biology)

Toxline (includes TSCATS) (1/11/2013) (4/17/2014) tert

- butanol OR 75

- 65

- 0 [rn] OR t

butyl hydroxide OR 2

- methyl -2- propanol OR trimethyl carbinol OR t

butyl alcohol OR tert

- butanol OR tert

butyl alcohol OR tert

- butyl alcohol Not PubMed

TSCATS2 (1/4/2013) (4/17/2014) 75

- 65

- 0 None



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**Table LS**  **-2**

**. Summary of additional search strategies for**  tert

**- butanol**  

### 1

Approach used Source(s) Date performed Number of additional references identified Manual search of citations from reviews Review article: Mcgregor (2010)

. Tertiary

- butanol: A toxicological review. Crit Rev Toxicol 40(8): 697

- 727. 1/2013 5

Review article: Chen (2005) . Am ended final report of the safety assessment of t

- butyl alcohol as used in cosmetics.” Int J Toxicol 24(2): 1

- 20. 1/2013 2

Manual search of citations from reviews conducted by other international and federal agencies IPCS (1987a)

. Butanol s: Four isomers: 1- butanol, 2

- butanol, ter t

- butanol
, isobutanol [WHO EHC]. Geneva, Switzerland: World Health Organization. 1/2013 None

OSHA (1992) . Occupational safety an d health guideline for tert

- butyl alcohol. Cincinnati, OH: National Institute for Occupational Safety and Health. 1/2013 None



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**Table LS**  **-3**

**. Questions and relevant experimental information for evaluation**  

### 1 2 **of experimental animal studies**

Methodological fe ature Question(s) considered Examples of relevant information extracted
Test animal Based on the endpoint(s) in question, are concerns raised regarding the suitability of the species, strain, or sex of the test animals on study? Test animal species, strain, sex

Experimental setup Are the timing, frequency and duration of exposure, as well as animal age and experimental group allocation procedures/ group size for each endpoint evaluation, appropriate for the assessed endpoint(s)? Age/lifestage of test animals at exposure and all endpoint testing timepoints

Timing and periodicity of exposure and endpoint evaluations; duration of exposure

S ample size for each experimental group (e.g., animals; litters; dams) at each endpoint evaluation
Exposure Are th e exposure conditions and controls informative and reliable for the endpoint(s) in question, and are they sufficiently specific to the compound of interest? Exposure administration techniques (e.g., route; chamber type)

Endpoint evaluation procedures Do the procedures used to evaluate the endpoint(s) in question conform to established protocols, or are they biologically sound? Are they sensitive for examination of the outcome(s) of interest? Specific methods for assessing the effect(s) of exposure, includ ing related details (e.g., specific region of tissue/organ evaluated)

Endpoint evaluation controls, including those put in place to minimize evaluator bias

Outcomes and data reporting Were data reported for all pre

- specified endpoint(s) and study groups, or were any data excluded from presentation/ analyses? Data presentation for endpoint(s) of interest

Note: “Outcome” refers to findings from an evaluation (e.g., hypertrophy ), whereas “endpoint” refers to the evaluation itself (e.g., liver histopatholog y). 

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# 1. HAZARD IDENTIFICATION

**1.1. PRESENTATION AND**  **SYNTHESIS OF EVIDENCE BY ORGAN/SYSTEM**  1

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32 **1.1.1. Kidney**  **Effects**

Synthesis of Effects in Kidney This section reviews the studies that investigated whether exposure to tert

- but a nol can cause kidney effects in hu ma ns or animals. The data base examining kidney effects following

tert

- butanol exp o sure contain s no human data

, six studies performed in rats or mice

, and one re

evaluation of the rat data from NTP (1995)

. Studies employing short

- term and acute exposures that examined kidney effects are not included in the evidence tables; however, they are discussed in the text if they provide data to support mode of action or hazard identification . No methodological
concerns were identified that would lead one or more studies to be considered less informative for assessing human health hazard. A pa t hology working group ( Hard et al., 20 11 ) re- examine d kidney histopathology from the NTP (1995) 13

- week and 2

- year studi es in rats to evaluate questions involving MOAs for renal tubule development. All slides were analyzed in a blinded manner. Hard et al. (2011) did report different incidences of adenomas or carcinomas compared with the original

NTP (1995) study; thus, these data were presented separately. Histopathological results from both
Hard and NTP will be considered for hazard identification.

t ert

- Butanol exposure resulted in a number of kidney effects after both oral (drinking

w a ter) and in h alation exposure in both sexes of rats and mice. Kidney effects observed after oral

exposure ( Tabl e 1- 1 ; Table 1- 2 ; Figure 1-1 ) include increased kidney weight in female rats and in male and female mice (13

- week ex p osure), and kidney inflammation, kidney transitional epithelial hyperplasia, and increased incidence and/or severity of kidney nephropathy in female rats (2

- yea r exposure) ( NTP, 1995 ). In a 2

- year oral exposure study in male rats, increased kidney weight, increased hyaline droplets, kidney transitional epithelial hyperplasia, kidney mineralization, renal tubule hyperplasia, and increased incidence and/or severity of kidney nephropathy were observed, with some of these effects seen at earlier time periods ( NTP, 1995 ) . Other kidney effects in male
rats were observed in a 10

- week o r al exposure study ( Acharya et al., 1997 ; Acharya et al., 1995 )

. No changes in clinical chemistry that would typically be indicative of kidney damage have been
observed with t e rt

- bu t anol exposure. Although there were some changes in urinalysis parameters (e.g., decreased urine volume and increa sed sp ecific gravity), this was accompanied by reduced water consumption and may not be related to an effect of kidney function. The kidney is also the target organ for cancer effects ( Tabl e 1- 3 ; Figure 1-1 )

. Male F344 rats
had an increased incidence of renal tubule adenomas and combined renal tubule adenoma or

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carcinoma in a 2

- year oral bioassay ( Hard et al., 2011 ; NTP, 1995 ). The highest exposure group had 

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3 5 an increase in mortality, which may in part explain the apparent non

- mo n otonicity in the observed
dose- response, in which the highest exposur e group had a lower incidence of tumors than the middle exposure group. An Independent Pathology Working Group (PWG), sponsored by Lyondell Chemical
C o mpany, re

- ev a luated the kidney changes in the NTP 2

- year study ( Hard et al., 2011 ) . The PWG consisted of senior pathologists with experience in chemically

- ind u ced nephrotoxicity and renal neoplasia. In all cases, PWG members were blinded to treatment groups to preclude any possible
bias

, and u s ed guidelines published by the Society of Toxicologic Pathology . T he PWG confirmed the NTP findings of atypical tubule hyperplasia and renal tubule tumors in male rats at 2

- year s

. In particular, they reported very similar overall tu m or incidences in the exposed groups. However, the PWG evaluation of the control groups reported fewer renal tubule adenomas and carcinomas than the original NTP study. As a result, based on the PWG evaluation, all treated groups had statistically significant increases in rena l t ubule adenomas and carcinomas (combined) as compared to controls . Ad d itionally, the PWG considered fewer of the tumors to be carcinomas as compared to the original NTP study. No chronic (2

- year ) inhalation exposure study is available, but minimal kidney effects were observed in rats (mainly the males) after t ert

- b u tanol exposure by inhalation for 13 weeks at concentrations ranging from 406

– 6

, 368 mg/m 3 ( NTP, 1997 ) ( Table 1-1 ; Ta ble 1- 2 ; Figure 1- 2 ). Absolute kidney weights were elevated ( 9. 8

– 11%) in male rats exposed at ≥3,27 4 mg/m 3 (not dose

dependent) ; r elative kidney weights were statistically elevated (~9%) in males at ≥3,27 4 mg/m 3 and females at 6,36 8 m g /m 3 . Male rats exhibited an increase in the severity of chronic nephropathy (characterized as number of foci of regenerative tubules). Alt h ough the kidney effects were less severe after inhalation exposure, a direct comparison can onl y b e made on the basis of internal
dose

. ARCO (1983) found that blood levels of tert

- butanol and its metabolites are equivalent after a single oral dose of 350 mg/kg compared to a single 6

- ho u r inhalation exposure to 6

, 164 mg/m 3

. That would indicate , b a sed on bolus exposure s, that the inhalation exposures used in the NTP (1997) study were in the range of the lower doses used in the NTP (1995) oral study. On the other hand, based on PBPK modeling, chronic exposure in the range of the NTP (1995) bioassay doses of
90

– 4 20 mg/kg

- d ay lead to the same average blood concentration of tert

- butanol as 6

- hour/day, 5 day/week inhalation exposure s to 860

– 4500 mg/m 3 , suggesting that the oral and inhalation exposures in NTP (1995) and NTP (1997) , respectively, overlap on the basis of internal dose. Finally, the lack of eithe r m ortality or changes in body weight (both observed with oral exposure) observed after the inhalation exposure sug g ests that a direct comparison cannot be made.

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- Butyl Alcohol

**Table 1-1.**  **Changes in kidney weight in animals following exposure**  **to** 

### 1 2 tert

**- butanol** 

Reference and s tudy d esign Results
Kidney weight (percent change as compared to control)

Lyondell Chemical Co. (200 4)

Sprague

- Dawley rat; 12/sex/treatment
G avage 0, 64, 160, 400, or 1,000 mg/kg -d Males: 9 weeks beginning 4 weeks prior to mating Females: 4 weeks prior to mating through PND21 Males

Dose (mg/kg

- d) Left absolute

w eight Left relative

w eight Right absolute

w eight Right relative

w eight

0 0 0 0 0

64 +6 +8 +6 +8

160 +9 +14* +6 +11*

400 +12* +14* +14* +17*

1,000 +18* +28* +20* +31*

Females

Dose (mg/kg

- d) Left absolute

w eight Left relative

w eight Right absolute

w eight Right relative

w eight

0 0 0 0 0

64

− 1

− 2 +2 0

160 0 0 +1 0

400 +3 +2 +4 +2

1,000 +4 0 +7 +2
NTP (1995)

F344/N rat; 10/sex/treatment Drinking water 0, 2.5, 5, 10, 20, 40 mg/mL M: 0, 230, 490, 840, 1,520, 3,610 a mg/kg -d F: 0, 290, 590, 850, 1,560, 3,620 a mg/kg -d 13 weeks Males Females

Dose (mg/kg

- d) Absolute

w eight Relative

w eight Dose (mg/kg

- d) Absolute

w eight Relative

w eight

0 0 0 0 0 0

230 +12* +19* 290 +19* +17*

490 +17* +26* 590 +16* +15*

840 +16* +32* 850 +29* +28*

1,520 +26* +54* 1,560 +39* +40*

3,610 All dead All dead 3,620 +36* +81*


### 3

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- Butyl Alcohol

**Table 1** 

**- 1** 

**. Changes in kidney weight in**  **animals following exposure to**  tert

**- butanol**  **(** continued **)**

Reference and s tudy d esign Results

NTP (1995)

B6C3F 1 mouse; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, 40 mg/mL) M: 0, 350, 640, 1,590, 3,940, 8,210 a mg/kg -d F: 0, 500, 820, 1,660, 6,430, 11,620 a mg/kg -d 13 weeks Males Females

Dose (mg/kg

- d) Absolute weight Relative weight Dose (mg/kg

- d) Absolute weight Relative weight

0 0 0 0 0 0

350 +1 +1 500 0

− 3

640 +3 +2 820

− 3

− 1

1,590 +2 +8 1,660 +1 0

3,940 +6 +22* 6,430 +6 +15*

8,210 0 +48* 11,620 +12* +35*

NTP (1995)

F344/N rat; 60/sex/treatment (10/sex/treatment evaluated at 15 months) Drinking water ( 0, 1.25, 2.5, 5, or 10 mg/mL) M: 0, 90, 200, or 420 a mg/kg -d F: 0, 180, 330, or 650 a mg/kg -d 2 years Males Females

Dose (mg/kg

- d) Absolute weight Relative weight Dose (mg/kg

- d) Absolute weight Relative weight

0 0 0 0 0 0

90 +4 +8 180 +8* +14*

200 +11 +15* 330 +18* +21*

420 +7 +20* 650 +22* +42*
Only animals sacrificed at 15 months were evaluated for organ weights. Organs were not weighed in the 2

- year mouse study
NTP (1997)

F344/N rat; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1,643, 3,273 or 6,368 mg/m 3 ) (dynamic whole

- body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical c oncentration and method were reported Males Females

Concentration (mg/m 3 ) Absolute weight Relative weight Absolute weight Relative weight

0 0 0 0 0

406 +1 +1

− 4

− 1

824

− 2

− 1 0 +1

1,643 +3 +2 +4 +4

3,273 +11* +8* +2 +2

6,368 +9.8* +9* +4 +9*

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- Butyl Alcohol

**Table 1** 

**- 1** 

**. Changes in kidney weight in**  **animals following exposure to**  tert

**- butanol**  **(** continued **)**

Reference and s tudy d esign Results

NTP (1997)

B6C3F 1 mouse; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1,643, 3,273 or 6,368 mg/m 3 ) (dynamic whole

- body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical concentration and method were reported Males Females

Concentration (mg/m 3 ) Absolute weight Relative weight Absolute weight Relative weight

0 0 0 0 0

406

− 6

− 4 +1

− 3

824

− 1 +3 +5 +9

1,643 +4 +3 +1

− 2

3,273

− 10

− 3 0 +7

6,368 +3 +6 +3 +15*



### 1 2 3 4 5 6

7 8 a The high

- dose group had an increase in mortality. * Statistically significant p ≤ 0 .0 5 as determined by the study authors. Percentage change compared to control = (treated value – co nt rol value) ÷ control value × 100. Conversions from drinking water concentrations to mg/kg

- d perf o rmed by study authors. Conversion from ppm to mg/m 3 is 1 ppm = 3.031 mg/m 3

.

**Table 1-2.**  **C hanges in kidney histopathology in**  **animals following exposure to**  tert

**- butanol** 

Reference and s tudy d esign Results

Acharya et al. ( 1997 ; 1995 )

Wistar rat; 5

– 6 males/treatment Drinking water (0 or 0.5%), 0 or
575 mg/kg -d 10 weeks ↑ tubular degeneration, degeneration of the basement membrane of the Bowman’s capsule, diffused glomeruli, and glomerular vacuolation (no incidences reported)

↓ kidney glutathione (~40%)*

Lyondell Chemical Co. (2004)

Sprague

- Dawley rat; 12/sex/treatment Gavage 0, 64, 160, 400, or 1,000 mg/kg -d F0 males: 9 weeks beginning 4 weeks prior to mating F0 females: 4 weeks prior to mating through PND21 There were no changes in kidney histopathology observed.



### 9

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- Butyl Alcohol

**Table 1** 

**- 2** 

**. Changes in kidney histopathology in**  **animals following exposure to**  tert

**- butanol**  **(** continued **)**

Reference and s tudy d esign Results

NTP (1995)

F344/N rat; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, or 40 mg/mL) M: 0, 230, 490, 840, 1,520, 3,610 a mg/kg -d F: 0, 290, 590, 850, 1,560, 3,620 a mg/kg -d 13 weeks Incidence (severity):

Males Females

Dose (mg/kg

- d) Mineralization Nephropathy Dose (mg/kg

- d) Mineralization Nephropathy

0 0/10 7/10 (1.0) 0 10/10 (1.7) 2/10 (1.0)

230 0/10 10/10 (1.6*) 290 10/10 (2.0) 3/10 (1.0)

490 2/10 (1.5) 10/10 (2.6*) 590 10/10 (2.0) 5/10 (1.0)

840 8/10*(1.4) 10/10 (2.7*) 850 10/10 (2.0) 7/10* (1.0)

1,520 4/10*(1.0) 10/10 (2.6*) 1,560 10/10 (2.0) 8/10* (1.0)

3,610 4/10*(1.0) 7/10 (1.1) 3,620 6/10 (1.2) 7/10* (1.0)

NTP (1995)

B6C3F 1 mouse; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, or 40 mg/mL ) M: 0, 350, 640, 1,590, 3,940, 8,210 a mg/kg -d F: 0, 500, 820, 1,660, 6,430, 11,620 a mg/kg -d 13 weeks H istopathology data for the 13

- week study were not provided , but the kidney was evaluated indicating that no changes in kidney histopathology were observed in the 13

- week study.

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- Butyl Alcohol

**Table 1** 

**- 2** 

**. Changes in kidney histopathology in**  **animals following exposure to**  tert

**- butanol**  **(** continued **)**

Reference and s tudy d esign Results

NTP (1995)

F344/N rat; 60/sex/treatment (10/sex/treatment evaluated at 15 months) Drinking water ( 0, 1.25, 2.5, 5, 10 mg/mL) M: 0, 90, 200, 420 a mg/kg -d F: 0, 180, 330, 650 a mg/kg -d 2 years Incidence (severity): Males

Dose (mg/kg

- d) Mineralization ( interim ) Mineralization ( terminal ) Linear mineralization ( terminal )

0 1/10 (1.0) 26/50 (1.0) 0/50

90 2/10 (1.0) 28/50 (1.1) 5/50* (1.0)

200 5/10 (1.8) 35/50 (1.3) 24/50* (1.2)

420 9/10* (2.3) 48/50* (2.2) 46/50* (1.7)

Dose (mg/kg

- d) Renal tubule hyperplasia (extended evaluation) Transitional epithelium hyperplasia Nephropathy severity

0 12/50 (2.3) 25/50 (1.7) 3.0

90 16/50 (2.3) 32/50 (1.7) 3.1

200 14/50 (2.2) 36/50* (2.0) 3.1

420 23/50* (2.8) 40/50* (2.1) 3.3*

Females

Dose (mg/kg

- d) Mineralization b

Interim Mineralization b
Terminal Inflammation (suppurative)

incidence

0 10/10 (2.8) 49/50 (2.6) 2/50

180 10/10 (2.9) 50/50 (2.6) 3/50

330 10/10 (2.9) 50/50 (2.7) 13/50*

650 10/10 (2.8) 50/50 (2.9) 17/50*

Dose (mg/kg

- d) Renal tubule hyperplasia Transitional epithelium hyperplasia Nephropathy severity

0 0/50 0/50 1.6

180 0/50 0/50 1.9*

330 0/50 3/50 (1.0) 2.3*

650 1/50 (1.0) 17/50*(1.4) 2.9*

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- Butyl Alcohol

**Table 1** 

**- 2** 

**. Changes in kidney histopathology in**  **animals following exposure to**  tert

**- butanol**  **(** continued **)**

Reference and s tudy d esign Results

NTP (1995)

B6C3F 1 mouse; 60/sex/treatment Drinking water ( 0, 5, 10, or 20 mg/mL) M: 0, 540, 1,040, or 2,070 a mg/kg -d F: 0, 510, 1,020, or 2,110 mg/kg -d 2 years No changes in kidney related histopathology observed. c

NTP (1997)

F344/N rat; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1,643, 3,273 or 6,368 mg/m 3 ) (dynamic whole

- body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical concentration and method were reported Male

Concentration

(mg/m 3 ) Average severity of chronic nephropathy

0 1.0

406 1.4

824 1.4

1,643 1.6

3,273 1.9

6,36 8 2.0
Severity categories: 1= minimal, 2= mild. No results from statistical tests reported
NTP (1997)

B6C3F 1 mouse; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1,643, 3,273 or 6,368 mg/m 3 ) (dynamic whole

- body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analyt ical concentration and method were reported There were no kidney effects observed.

a The high

- dose group had an increase in mortality. 

### 1 2 3 4 5 6 7 8 b Linear mineralization not observed in female rats.

c Organs were not weighed in mice during the 2

- year study. Relative organ weights refer to relative to body weight * Statistically significant p ≤ 0 .0 5 as determined by the study authors. Conversions from drinking water concentrations to mg/kg

- d perf o rmed by study authors. Conversion from ppm to mg/m 3 is 1 ppm = 3.031 mg/m 3

.

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- Butyl Alcohol

**Table 1-3.**  **Changes in kidney tumors in**  **animals following exposure to**  

### 1 2 tert

**- butanol** 

Reference and s tudy d esign Results

NTP (1995)

F344/N rat; 60/sex/treatment (10/sex/treatment evaluated at 15 months) Drin king water ( 0, 1.25, 2.5, 5, or 10 mg/mL) M: 0, 90, 200, or 420 a mg/kg -d F: 0, 180, 330, or 650 a mg/kg -d 2 years Male
Dose (mg/kg

- d) Renal tubule adenoma

(single) Renal tubule adenoma (multiple) Renal tubule carcinoma Renal tubule adenoma (single or multiple) or

carcinoma

0 7/50 1/50 0/50 8/50

90 7/50 4/50 2/50 13/50

200 10/50 9/50* 1/50 19/50*

420 10/50 3/50 1/50 13/50

Female

Dose (mg/kg

- d) Renal tubule adenoma

(single) Renal tubule adenoma (multiple) Renal tubule carcinoma Renal tubule adenoma (single or multiple) or

carcinoma

0 0/50 0/50 0/50 0/50

180 0/50 0/50 0/50 0/50

330 0/50 0/50 0/50 0/50

650 0/50 0/50 0/50 0/50

Results do not include the animals sacrificed at 15 months.
Hard et al. (2011)

reanalysis of the slides from male
rats in the NTP (1995) study (see above) Male
Dose (mg/kg

- d) Renal tubule adenoma

(single) Renal tubul e adenoma (multiple) Renal tubule carcinoma Renal tubule adenoma (single or multiple) or

carcinoma

0 3/50 1/50 0/50 4/50

90 9/50 3/50 1/50 13/50*

200 9/50 9/50 0/50 18/50*

420 9/50 3/50 1/50 12/50*

NTP (1995)

B6C3F 1 mouse; 60/sex/treatment Drinking water ( 0, 5, 10, or 20 mg/mL) M: 0, 540, 1,040, or 2,070 a mg/kg

d F: 0, 510, 1,020, or 2,110 mg/kg -d 2 years No changes in kidney

- related tumors



### 3 4 5 6 a The high

- dose group had an increase in mortality. * Statistically significant p ≤ 0 .0 5 as determined by the study authors. Conversions from drinking water concentrations to mg/kg

- d perf o rmed by study authors.

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- Butyl Alcohol



### 1 2 3

4 5

This document is a draft for review purposes only and does not constitute Agency policy. 1- 10 DRAFT

— DO NOT CITE OR QUOTE Sources: (A) Acharya et al. ( 1997 ; 1995 ) ; (B ) Hard et a l. (2011) *; (C) Lyondell Chemical Co. (2004) (D) NTP (1995) ; * reanalysis of NTP (1995)
**Figure 1-1**  **. Exposure response array f**  **or kidney**  **effects following oral exposure to**  tert

**- butanol** 

**.**Toxicological Review of tert

- Butyl Alcohol



### 1 2

3 4 Absolute weight; M Rat

Relative weight; M Rat

Absolute weight; F Rat

Relative weight; F Rat

Absolute/relative weight; M Mouse

Absolute weight; F Mouse

Relative weight; F Mouse

100 1,000 10,000 

### Concentration ( mg/m3)  = exposures at which the endpoint was reported statistically significant by study authors  = exposures at which the endpoint was reported not statistically significant by study authors

Source: NTP (1997)
**Figure 1-2**  **. Exposure**

**- response array of kidn**  **ey effects following subchronic inhalation exposure to**  tert

**- butanol**  **(no chronic studies available).**

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- Butyl Alcohol

Mode of Action Analysis

— Kidney E

ffects 

### 1

2

3

4
5
6
7
8
9
10
11
12
13
14
15
16
17
18
19
20
21
22
23
24
25
26
27
28
29

30

31

32 33

34

35 Mode of Action Analysis for α 2u - globulin

- associated nephropathy

Description of the hypothesized MOA

Several studies were identified that evaluated the role of α 2u - globulin in tert

- butanol

induced renal tumor development ( Borghoff et al., 2001 ; Williams and Borghoff, 2001 ; Takahashi et al., 1993 ). α 2u -G lobulin is a member of a large superfamily of lo w- molecular

- weight proteins and
was first c haracterized in male rat urine. Such proteins have been detected in various tissues and
fluids of most mammals ( inc l uding humans ), but the particular isoform of α 2u - globulin commonly detected in male rat urine is considered specific to th e m ale rat. The hypothesized sequence of α 2u - globulin

- associated nephropathy, as described by U.S. EPA (1991a) , is as follows. Chemicals that induce α 2u - globulin accumulation do so rapidly. The
accumulation of α 2u - globu lin in the hyaline droplets results in hyaline droplet deposition in the P2 segment of the proximal tubule within 24 hours of exposure. Hy ali ne droplets are a normal constitutive feature of the mature male rat kidney; they are particularly evident in the P 2 s egment of the proximal tubule and contain α 2u - globulin ( U.S. EPA, 1991a ) . Abnormal increases in hyaline droplets have more than one etiology and can be associated with the accumulation of different proteins. As hyaline dr o p l et deposition continues, single

- cell necrosis occurs in the P2 segment
whic h leads to exfoliation of these cells into the tubule lumen within 5 days of chemical exposure. In

r es ponse to the ce ll loss, cell proliferation is observed in the P2 segment aft er 3 weeks and continues for the duration of the exposure. After 2 or 3 weeks of exposure, the cell debris accumulates in the P3 segment of the proximal tubule to form granular casts. Continued chemical exposure for 3 to 12 months leads to the formation of ca l ci um hydroxyapatite in the papilla which results in linear mineralization. After 1 or more years of chemical exposure, these lesions may
result in the induction of renal adenomas and carcinomas. U.S. EPA (1991a) states that two questions must be addressed to determine the extent to
which α 2u - globulin mediated processes induce renal tumors and nephropathy. First, it must be
determined whether the α 2u - globulin process is occurring in male rats and therefore could be a
factor in renal effects. U.S. EPA (1991a) states that the criteria for answering this question in the
affirmative are as follows:

1) h yaline droplets are increased in size and number in male rats,

2 ) th e protein in the hyaline droplets in male rats is α 2u - globulin, and

3) if several (but not necessarily all) additional steps in the pathological sequence are present in ma le rats, such as:

(a) single

- ce l l necrosis,

(b) exfoliation of epithelial cells into the tubular lumen,

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- Butyl Alcohol

(c) granular casts,

(d) linear mineralization, and

(e) tubule hyperplasia.

Th e available data relevant to this question in male rats are summarized in Table 1-5 and Figure 1- 3 a nd Figure 1 -4 , and will be evaluated below in accordance with the mode of action ( MOA ) fr amework from the EPA cancer guidelines ( U.S. EPA, 2005a )

. If the α 2u - globulin process is operative, then U.S. EPA (1991a) states that a second question
must be answered as to whether the renal eff ects are solely due to the α 2u - globulin process, are a
combination of the α 2u - globulin process and other carcinogenic processes, or are due primarily to other processes. U.S. EPA (1991a) states that the follow ing types of data may be useful for answering this question: 

### 1

2

3

4
5
6
7
8
9
10
11

12

13

14

15

16

17

18

19

20

21
22
23
24
25

26 27 28

29 30 31 1 ) Hypothesis

- tes t ing data

2) Biochemical information

3) S u stained cell division in the proximal tubule of the male rat

4) Stru cture

- activ ity relationships

5) Covalent binding to macromolecules

6) G e notoxicity

7) Nep hrotoxicity

8 ) A nimal bioassay data in other species

- , sex

- comb inations

9) Other information not specifically listed

Th e av ailable data relevant to this question are summarized in Table 1- 6 , and will be evaluated below in accordance with the MOA framework from the EPA cancer guidelines ( U.S. EPA,
2005a ). From these two questions, U.S . EPA (1991a) states that one of three possible conclusions

can be made:

• If renal tumors in male rats are attri butable solely to the α 2u - globulin process, then U.S. EPA (1991a) states that such tumors will not be used for human cancer hazard identification or for dose

- response extrapolations.

• If renal tumors in male rats are not linked to the α 2u - globulin process, then U.S. EPA (1991a) states that such tumors are an appropriate endpoint for human hazard identification and are considered, along with other appropriate endpoints, for quantitative risk estimation.

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- Butyl Alcohol

• I f some renal tumors in male rats are attributable to the α 2u - globulin process and some attributable to other carcinogenic processes, then U.S. EPA (1991a) states that such tumors

rema in relevant for purposes of hazard identification, but a dose

- response estimate based on such tumors in male rats should not be performed unless there is enough information to determine the relative contribution of each process to the overall renal tumor r esponse.
Additionally, U.S. EPA (1991a) states that if the α 2u - globulin process is occurring in male rats, then the associated nephropathy in male rats (described above) would not be an appropriate endpoint to determine noncancer effects occurring in humans. In such a case, the characterization of human health hazard for renal toxicity would need to rely on other types of nephrotoxic effect data in male rats and/or on nephrotoxic effect data in female rats o r other species.

**Table 1-4.**  **Additional kidney effects potentially relevant to mode of action in animals following exposure**  **to** tert

**- butanol**  

### 1 2 3 4 5

6
7
8
9
10

11 12

Reference and s tudy d esign Results

Williams and Borghoff (2001)

F344 rats; 4/sex Single gavage dose: 500 mg/kg Males: ↑ binding of tert

- butanol to α 2u - globulin compared to females* Females: no change in binding observed

NTP (1995)

F344/N rat; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, or 40 mg/mL) M: 0, 230, 490, 840, 1,520, 3,610 a mg/kg -d F: 0, 290, 590, 850, 1,560, 3,620 a mg/kg -d 13 weeks Accumulation of hyaline droplets:

Male
Dose (mg/kg

- d) Hyaline droplet accumulation

0 0/10

230 + e

490 ++

840 ++

1,520 ++

3,610 0/10

No information provided on females. No results from statistical tests reported.

Hard et al. (2011)

Rean alysis of the slides in the NTP (1995) study Males: Confirmed accumulation of hyaline droplets increased with increasing dose

- levels in 13 week study above. No incidence data available.

Females: not evaluated


### 13

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- Butyl Alcohol

**Table 1** 

**- 4** 

**. Additional kidney effects potentially relevant to mode of action in**  **animals following exposure to**  tert

**- butanol**  **(** continued **)**

Reference and s tudy d esign Results

Borghoff et al. (2001)

F344 rat; 5/sex/treatment Analytical concentration:0, 250, 450, 1,750 ppm (0,771, 1,387 or 5,395mg/m 3 ) 6hr/d 10 days Males: positive trend for a ccumulation of protein droplets ( p < 0.05), significant increase in accumulation of α 2u - globulin at 5,39 5 mg/m 3 as compar ed to controls (no incidence data provided) Females: No positive staining for α 2u - globulin was observed in exposed female rats.



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

- dose group had an increase in mortality.
b Linear mineralization not observed in female rats.

c Organs were not weighed in mice during the 2

- year study.
d Standard & extended evaluation combined.

e + or ++ indicated an increased accumulation relative to controls, as reported by the authors; no additional incidence data an d

no results from statistical t e sts available. * Statistically significant p ≤ 0 .0 5 as determined by the study authors. Conversion from ppm to mg/m 3 is 1 ppm = 3.031 mg/m 3

. Percentage change compared to control = (treated value – co nt rol value) ÷ control value × 100. Conversions from dri n king water concentrations to mg/kg

- d performed by study authors.

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- Butyl Alcohol

**Table 1-5**  **. Summary of data on**  **the α**  **2u -**  **globulin process in male rats exposed** 

### 1 2 **to** tert

**- butanol** 

Criterion Duration Results Reference (1) hyaline droplets are increased in size and number 10 d + Borghoff et al. (2001) 13 wk ( + ) NTP (1995) 13 wk

– NTP (1997) 13 wk (+) a Hard et al. (2011) (2) the protein in the hyaline droplets is α 2u - globulin 12 hr + Williams and Borghoff (2001) 10 d + Borghoff et al. (2001) (3) Several (but not necessarily all) additional steps in the pathological sequence are present in male rats, such as: (a) Single

- cell necrosis 13 wk

– NTP (1995) 2 yr

– NTP (1995) 2 yr

– Hard et al. (2011) (b) exfoliation of epithelial cells into the tubular lumen 13 wk

– NTP (1995) 2 yr

– NTP (1995) 2 yr

– Hard et al. (2011) (c) granular casts 13 wk

– NTP (1995) 13 wk (+) a,c Hard et al. (2011) 13 wk

– NTP (1997)

2 yr

– b NTP (1995) (d) linear mineralization 13 wk

– NTP (1995) 13 wk

– NTP (1997)

2 yr + NTP (1995) 2 yr (+) a Hard et al. (2011) (e) tubule hyperplasia 2 yr + NTP (1995) 

### 3 4 5 6 7 8 9 10 + = Statistically significant change reported in one or more treated groups. (+) = Effect was reported in o ne or more treated groups, but statistics not reported.

– = No statistica l l y significant change reported in any of the treated groups.

a Re

- analysis of one control and one treated group from NTP (1995)
b Protein casts reported, not granular casts

c Precursors to granular casts reported

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### 1

2 3 4 **Figure 1-3**  **. Exposure**

**- response array for components of**  **α 2u**  **- globulin**  **nephropathy and renal tumors in male rats after oral exposure to**
tert

**- butanol** 

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- Butyl Alcohol

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### 1

2 3 4 **Figure 1-4**  **. Exposure**

**- response array for components of**  **α 2u**  **- globulin**  **nephropathy and renal tumors in male rats after inhalation exposure to**
tert

**- butanol** 

**.**Toxicological Review of tert

- Butyl Alcohol

**Table 1-6**  **. Summary of additional**  **data informing the contribution of the**  

### 1 2 3 **α 2u**  **- globulin process**  **on the renal**  **tumor development**  **in male rats exposed to** 
tert

**- butanol** 

Type of data Description Reference

(1) Hypothesis

- testing data No data (2) Biochemical information Reversible, non

- covalent binding of tert

- butanol to α 2u - globulin

. Williams and Borghoff (2001) (3) Sustained cell division in the proximal t ubule of the male rat Hyperplasia at 2 yr reported in both male and female rats, attributed to CPN. Hard et al. (2011)

No effect on proliferation at 13 wk NTP (1997) Increased proliferation at 1 3 wk based on PCNA assay NTP (1995) Increased proliferation of the P2 segment at 10 days based on BrdU labeling Borghoff et al. (2001)

(4) Structure

- activity relationships No data (5) Covalent binding to macromolecules No data (6) Genotoxicity Limited database to conclude tert

- butanol is genotoxic or non

genotoxic See Appendix B.3

.
(7) Nephrotoxicity Increased tubular regeneration and intratubule protein cast formation
at 2 yr in males and females, with effects in females occurring at lower dose. NTP (1995)

Increased severity of CPN in male rats after 13 wk inhalation exposure NTP (1997) Increased CPN in male and female F344/N rats following drinking water exposure for 13 wk. NTP (199 5)

Increased CPN in male and female F344/N rats following drinking water exposure for 2 yr

. NTP (1995)

(8) Animal bioassay data in other species

- , sex - combinations Two renal tubular adenocarcinomas (not statistically significant as compared to concurrent controls) reported in male mice following drinking water exposure for 2 yr . These t umors are very rare in mice. NTP (1995)

(9) Other data
Dose

- response and tempor al concordance (see Figures 1

- 3 and 1

- 4). 

### 4
5

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Strength, consistency, specificity of association 

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- butanol ?

The first criterion considered is whether hyaline droplets are increased in size and number
in male rats

. P r otein droplet accumulation was statistically significantly increased in the kidneys of male rats exposed to 5 ,3 05 mg/m 3 tert

- butanol for 6 hr/day for 10 days ( Borghoff et al., 2001 ) . Data from drinking water studies ( NTP, 1995 ; Takahashi et al., 1993 ; Lindamood et al., 1992 )
demonstrated a st a tistically significant increase, except at the highest dose, in hyaline droplet formation and severity in the proximal tubule of male rats following o ral exposure to t e rt

- butanol for 13 weeks. Treated males had large hyaline droplets with crystal accumulation, but the controls
had s mall droplets without crystals. NTP (1997) stained for hyaline droplet formation in male rats exposed to 0, 3,273, or 6,368 m g/ m 3 tert

- butanol via inhalation for 13 weeks

, and there was no
difference between the contro l s and treatment groups

. The second criterion considered is whether the protein in the hyaline droplets in male rats

is α 2u - globulin

. Two studies measured α 2u - globulin immunoreac tivity in the hyaline droplets of the renal proximal tubular epithelium ( Borghoff et al., 2001 ; Williams and Borghoff, 2001 )

. Borghoff et al. (2001) observed α 2u - globulin immunoreactivity present in the hyaline drop lets in the male rats.

No α 2u - globulin immunostaining was observed in the kidneys of the female rats. Williams and Borghoff (2001) found the content of α 2u - globulin statistically significantly elevated in 72 % of the kidneys of male rats treated with tert

- b ut anol compared with controls treated with corn oil.
The third criterion c o n sidered is whether several (but not necessarily all) additional steps in the pathological sequence are present in male rats. S ev eral, but not all, of the subsequent histopathological lesions were observed in the available subchronic or chronic tert

- b ut anol exposure studies. L i n ear mineralization was the lesion most consistently observed in male rats and
was found to be st a tistically significantly increased in male rats after 2 years of oral exposure ( NTP,
1995 ) (see Table 1- 2 ). The 13

- week study in rats ( NTP, 1995 ) reported mineralization, but it was
not characterized as linear mineralization. Additiona l l y, although the inhalation study by NTP (1997) did not report linear mineralization at 13 w ee ks , this may be due to the lower internal dose as compared to the oral studies. Atypical tubule hyperplasia was statistically significantly increased at the highest dose following 2 y ear s of oral exposure ( NTP, 1995 ) . Granular casts were increased at the 13 w ee k time point ( NTP, 1995 ) , though statistical significance was not determined . The reanalysis of the 13

- w e ek data by Hard et al. (2011) concluded that the lesions were precursors to granular casts a n d not the casts themselves. Other studies did not observe granular casts at 13

w ee k s or 2 y ea rs ( NTP, 1997
, 1995 )

. The formation of p rotein casts were observed, but these may be part of the pathology of CPN and are not thought to be related to α 2u -globulin nephropathy ( NTP,
1995 ) . Neither necrosis nor epithelial exfoliation was reported in any study. In summary, the evidenc e s upports the conclusion that tert

- butanol causes increases in the
size and number of hyaline droplets

, a nd the accumulating protein in the hyaline droplets is

α 2u -globuli n. Additionally, several, but not all, of the additional steps in the pathological sequence

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were observed, and not always consistently across studies. Therefore, the overall strength, 

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38 consistency, and specificity of the association between tert

- butanol and the hypothesized key
events is moderate.

Are t he renal effects in male rats expos ed to tert

- butanol solely due to the α 2u - globulin process ?

As summarized in T ab le 1-6 , there are many potential sources of additional data that may
inform the contribution of the α 2u - globulin process on renal tumor development. No hypothesis

testing data, structure

- a ct ivity r elationships, or covalent binding data were located, so these types
of data are not discussed. Additional data related to dose

- r e sponse concordance and temporal relationships ar e d iscussed in subsequent sections. In terms of biochemical information, Williams and Borghoff (2001) report that tert

- butanol reversibly and non

- c o valen tly bind s to α 2u - globulin. This provides additional support to the
evidence that the α 2u - process is occurring, but it does not inform the relative contribution to renal tumor development. Sustained cell division in the proximal tubule of the male rat is consistent with, though not specific to, the α 2u - process. Proliferation of the proximal tubule was significantly increased in male rats after 10 days of inhalation exp o s ure to tert

- butanol at concentrations of 771

– 5,395 mg/m 3 ( Borghoff et al., 2001 ) and after 13 week s of oral exposure to 1520 mg/kg

- day ( NTP, 1995 ; Takahashi et al., 1993 ; Lindamood et al., 1992 ) , but not after 13 week inhalation exposure up to
6368 m g/m 3 ( NTP, 1997 ) . Therefore, it is unclear the extent to which increased cell division is sustained. While hyperplasia was reported in chronic studies ( Hard et al., 2011 ) , it was observed in both male and female rats, and a ttri buted to CPN. There are a limited number of studies available to assess the genotoxic potential of tert

butanol ( see App e ndix B.3 in Supplemental Information for further details). tert

- Butanol was generally negative in a variety of genotoxicity ass a y s and cell systems including Salmonella typhimurium

, E s cherichia coli and Neurospora crassa ( Mcgregor et al., 2005 ; Zeiger et al., 1987 ; Dickey et al., 1949 ) . Studies also demonstrate negative results for gene mutations, sister chromatid exchanges, micronucleus formation

, a n d chromosomal a berrations ( NTP, 1995 ; McGregor et al.,
1988 ) . However, DNA adducts were found in male Kunming mice ( Yuan et al., 2007 ), and DNA damage w as observed i n human HL

- 60 leukemia cells ( Tang et al., 1997 ) . In another study by Sgambato et al. (2009)

, an initial increase in DNA damage was observed as measured by nuclear fragmentation, but the damage declined d r astically following 4 hours of exposure and disappear ed entirely a fter 12 hours of exposure to t ert

- butanol

. In terms of nephrotoxicity, a number of renal effects have been reported in female rats and/or in mice. Kidney transitional epithelial hyperplasia and inflammation were significantly
increased in bo t h male and female F344 rats exposed for 2 years via oral exposure

. F344 rats exposed to tert

- b ut anol at dose ranges of 230

– 1520 mg/kg

- day and 850

– 3,620 mg/kg

- day in males and females, respectively, ex h ibited a statistically significant increase in the incidence of nephropathy co mp ared with controls after 13 weeks of exposure ( NTP, 1995 ). Nephropath y

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severity was also significantly increased at 420 mg/kg

- day in males and 180

- 650 mg/kg

- day in 

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37 females after 2 years of exposure ( NTP, 1995 ). Average severity of chronic n ephropathy was
minimal to mild in males after a 1 3

- w eek inhalation exposure ( NTP, 1997 )

. Female rats also had lesions associated with nephropathy ( NTP, 1995 ) , but none of the lesions were similar to those
observed in the male rat that are associated with α 2u - globulin nephropathy. With respect to renal tumors, no s ta t istically significant increases in renal tumors were reported in t e rt

- butanol

- exposed female rats or mice compared with concurrent controls. T wo renal tubular adenocarcinomas were reported in male mice following drinking water exposure for 2 years ( NTP, 1995 ) (one each in the low and high dose groups). Although s uch tumors are very rare in mice, with h i storical control incidences of 2/1351 (0.15%) in feeding studies and 4/1093 (0.37%) in chamber studies ( Haseman et al., 1998 ) , these data are not sufficient to indicate that the kidney t umors observed in mice exposed to tert

- b utanol are treatment -related

. Overall, the strength, consistency, and specificity of the data supporting a t e rt

- butanol

induced α 2u - globulin process as the sole actor for renal effects in male rats is weak to moderate.

Dose

- re sponse concordance

Is

the α 2u - globulin process occurring in male rats exposed to tert

- butanol ?

As show n i n F igure 1- 3 and Figure 1 -4 , the dose

- response concordance among hypothesized key events is mixed. Borghoff et al. (2001) exposed male and female F344 rats to tert

- butanol at concentrations of 758, 1,364, or 5 ,3 05 mg/m 3 for 6 hr/day for 10 days to assess the role of α 2u - globulin nephropathy and renal cell proliferation . S i gnificant tubular proliferation in males was observed at all exposure levels, but accumulation of α 2u - globulin

- positive hyaline droplets was increased only at the highest dose ( Borghoff et al., 2001 ) . These data suggest that cell proliferation may be related to
the α 2u - globulin process only at the highest exposure concentration.
The d ose

- r esponse relationship s observed after 13 weeks and 2

- years were also only moderately concordant

. D a ta from a drinking water stud y ( NTP, 1995 ; Takahashi et al., 1993 ; Lindamood et al., 1992 ) demonstrated hyaline droplet formation in the proximal tubule of male rats at all tested doses ( ex c ep t at the highest dose where all rats died during weeks 5

- 12) following oral
exposure to tert

- bu t ano l for 13 weeks. PWG reevaluation by Hard et al. (2011) reported observing precursors to granular casts at the only dose level evaluated (1,520 m g/ kg

- day). Spontaneous
minerali z ation was observed, but the linear mineralization characteristic of this MOA was not

o bserved at any dose

. At t he 2

- year timepoint, linear mineralization was observed at all exposure
levels from 90

- 4 2 0 mg/kg -day

, and renal tubular hyperplasia was observed at the highest dose of 420 mg/kg -d ay ( NTP, 1995 ) , consistent with the expected dose

- respon se relationship . Notably, however, g ranular casts w ere not observed. Overall, the dose

- r es ponse concordance of the association between tert

- butanol and the hypothesized key events is moderate.

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Are the renal effects in male rats exposed to tert

- butanol solely due to the α 2u - globulin process ? 

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- response concordance between the hypothesized key events and the occurrence of
renal tumors c a n inform whether carcinogenesis is solely due to the α 2u - globulin process. Male
F344 rats ex h ibited an increased inc idence of renal tubule adenomas and combined renal tubule
adenoma or carcinoma in a 2

- y ear oral bioassay ( Hard et al., 2011 ; NTP, 1995 ) . Increased tumors were observed at 90, 200, and 420 m g/ kg

- day. Although some effect s related to α 2u - globulin nephropathy

, i nc luding hyaline droplets and linear mineralization, were observed at all doses, tubule hyperplasia was n o t observed at doses lower than 420 mg/kg

- day in any study and only precursor granular casts were observed at a much higher dose of 1,520 m g/ kg -day

. Moreover, t he
middle d ose of tert

- butanol induced the greatest incidence of tumors , so increasing the dose from
200 to 420 m g/ kg

- day led to additional markers of α 2u -globulin nephropathy in the form of tubule hyperplasia, bu t without any increase in tumor burden

. Thus

, tert

- butanol induced tumors at lower doses than for other precursor effects such as hyperplasia a n d granular casts

, suggesting a weak dose response c o ncordan ce with the incidence of tumors. Therefore, on the basis of w e ak dose

- response conc ordance, the data suggest that the observed tumors are not solely due to α 2u -globulin and that other processes are primarily responsible for tumors.

Temporal relationship

Is the α 2u - globulin process occurring in male rats exposed to tert

- butanol ?

As shown in Table 1

- 2 and Table 1-4 , hyaline droplets and α 2u - globulin accumu lation we re observed after a s i ngle dose or 10

- day exposure ; precursors to granular casts were observed at 13
weeks ; an d tubular hyperplasia was observed at 2 years

. The observations are consistent with the expected temporal relationship ( Hard et al., 2011 ; Borghoff et al., 2001 ; Williams and Borghoff,
2001 ; NTP, 1995 ) . However, the absence of other key events such as necrosis, exfoliation, and granular casts in most other studies at the a n t icipated time points weaken the case for α 2u - globulin
MOA. Ad ditionally, the NTP’s 13

- w eek study in rat reported kidney mineralization, but it was not
characterized as linear mineralization.

Are t he renal effects in male rats exposed to tert

- b utanol solely due to the α 2u - globulin process ? This question cannot be answered from data on temporality.

Biological plausibility and coherence Both U.S. EPA and IARC have accepted the general biological plausibility and coherence of
role of α 2u - globulin

- mediate d nephropathy in renal tumor induction ( Swenberg and Lehman

McKeeman, 1999 ; U.S. EPA, 1991a ) , and those rationales will not be repeated here. However, a retrospective analysis has suggested that a number of α 2u - globulin inducing chemicals fail to induce many of the pathological sequences in the α 2u - gl obulin pathway ( Doi et al.,

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2007 )

. For instance, dose

- response concordance was not observed for several endpoints such as 

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36 linear mineralization, tubular hyperplasia, granular casts, and hyaline droplets following exposure

to α 2u - globulin

- inducing chemicals such as d -l imonene, decalin, propy lene glycol mon o-t- butyl
ether , and Stoddard solvent IIC (SS IIC) . Although some of these chemicals induced histopathological lesions that exhibited a dose response, all of them failed to induce a dose

response trend for at least one of the endpoints in the sequence. Furthermore, no endpoint in the pathological sequence was predictive for tumor incidence. Tumor incidence did not exhibit a dose

response following either d

- limonene or decalin exposure. Finally, tumor incidence was not predicted by the severity of a particular effect in the α 2u -globulin sequence as demonstrated by SS IIC which induced some of the most severe nephropathy precursors relative to the other chemicals but did not significantly increase kidney tumors ( Doi et al., 2007 )

. Thus, these analyses suggest that another MOA may be operative for inducing tumors. Moreover, renal tumors were not observed following exposure to ETBE, which is rapidly
metabolized to tert

- b ut anol . Specifically

, Suzuki et al. (2012) and Saito et al. (2013) reported no increase in renal tumo r s in both sexes of Fischer 344 rats following 2

- year oral or inhalation exposures to ETBE at doses that yield similar internal concentrations (based on PBPK modeling) of

tert

- b utanol c ompared with the concentrations of the tert

- butanol bioassays. After 1 3 weeks of
exposure to tert

- bu t anol or ETBE, hyaline droplets were increased in a dose

- response manner

. ETBE exposure increased hyaline droplets at lower internal concentrations of tert

- bu t anol than by
direct t ert

- b utanol a dministration. Similar to hyaline droplets, linear mineralization was increased
at an internal tert

- bu t anol concentration approximately tenfold lower following ETBE exposure
than a t ert

- b utanol e xposure. By contrast, t ubule hyperplasia and renal tumors were both observed following a 2

- y ear exposure to tert

- butanol but not following ETBE exposure

. Renal tumors and tubule hyperplasia were not observed following any ETBE exposure despite achieving similar blood
concentrations of t e rt

- butanol as the NTP (1995) study. The failure of internal tert

- butanol concentrations to induce histopathological lesions early in the 

### α 2u 

### - globulin pathological sequence at blood levels that later induced hyperplasia and tumors su g gests a lack of coherence across the
two data s ets

.

C onclusions about the hypothesized MOA for α 2u - globulin

- associated nephropathy
Is the hypothesized MOA sufficie n t ly supported in test animals?

This conclusion is divided into two sub

- q u estions: whether the α 2u - globulin process is occurring in male rats, and whether it is the sole contributor to renal effects in male rats. With respect to the first question, tert

- bu t a nol induced increase s in α 2u -globulin deposition and hyaline droplet accumulation, and several o f t he subsequent steps in the pathological sequence were observed. These data provide sufficient evidence that the α 2u - globulin process can be operating g iven sufficient t ert

- butanol exposure

.

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— DO NOT CITE OR QUOTEIs the hypothesized MOA relevant to humans? Wit h respect to the second question, m ale rats are more sensitive to the kidney effects of

tert

- b utanol

, and t he available data indicate that male rats accumulate α 2u - globulin in the kidney, which is a specific MOA for male rats. Many of the steps in the pathological sequence of lesions
related to α 2u - globulin

- associated nephropathy were observed exclusively in male rats but not in
female rats

, or mice of either sex, and renal tumors occurred only in male rats. However, there is insufficient evidence to support a conclusion that α 2u - globulin nephropathy is the sole or primary contributor to renal tumor development. Given the inconsistencies and limitations of the genotoxicity database, the effect of tert

- butanol with respect to genotoxicity cannot be ruled out. Additionally, tert

- butanol induced tumors at lower doses than other precursor effects such as hyperplasia and granular casts , with no further increase in tumor incidence coinciding with the
additional mark ers of α 2u - globulin nephropathy. Thus, renal tumors observed in male rats are unlikely to be confounded by the presence of α 2u - globulin nephropathy. Therefore, on the basis of a weak dose

- response concordance, the data support a conclusion that processes other than

α 2u - globulin nephropathy are likely responsible for renal tumor development.

Based on the conclusion that processes other than α 2u - globulin nephropathy are likely responsible for renal tumor developme nt induced by tert

- butanol

, U.S. EPA (1991a) states that the following conclusion will be made:

Therefore, kidney tumors a re relevant to humans for purposes of hazar d identification and
dose- r esponse assessment . Because female rats and both sexes of mice do not have 

### α 2u 

### - globulin present, kidney effects in these animals are considered relevant to humans for both hazard
identification and dose

- r esponse.

Which po pulations or lifestages can be particularly susceptible to the hypothesized MOA? This question is not applicable.

Alternative MOA hypotheses with inadequate data for analysis Other nephrotoxic responses, such as exacerbation of CPN, inflammation, transit ional epithelial hyperplasia, and increased kidney weight, are observed in rats and/or mice, suggesting other possible processes are operative. I t has been proposed that enhanced chronic progressive nephropathy (CPN) is a mode of action for chemically

- i nd uced kidney tumors in male rats and that • I f renal tumors in male rats are not linked to the α 2u - globulin process, then ( U.S. EPA, 1 991a ) states that such tumors are an appropriate endpoint for human hazard identification and are considered, along with other appropriate endpoints, for quantitative risk estimation. Toxicological Review of tert

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r enal tubule tumors induced by chemicals that concomitantly exacerbate CPN are not relevant to 

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38 humans ( Hard and Khan, 2004 ). CPN is an age

- r elated renal disease of unknown etiology that occurs spontaneously in rats, especially the F344, Sprague

- Daw ley, and Osborne

- Mendel strains. Additional markers associated
with CPN include elevated protein and albumin in the urine and increased BUN, creatinine, and
cholesterol in the serum ( Hard et al., 2009 ) . CPN is often more severe in males compared with
females. S everal of the CPN pathological effects are similar to and can obscure the lesions
characteristic of α 2u -globulin- related hyaline droplet nephropathy ( Webb et al., 1990 ). Additionally,
renal effects of α 2u - globulin accumulation can exacerbate the effects associated with CPN ( U.S. EPA,
1991a ). However, Webb et al. (1990) suggested that exacerbated CPN was one component of the nephropathy resulting from exposure to chemicals that induce α 2u - globulin nephropathy. Male rat sensitivity has been noted with both CPN and α 2u - globulin nephropathy. Increased severity of CPN occurred in both male and female rats as a result of tert

- bu tanol

exposure. S ome of the observed renal lesions in male rats following exposure to tert

- butanol are effects commonly associated with CPN. Hard et al. (2011) concluded that the observation of transitional epithelial hyperplasia in the 2

- ye ar drinking study conducted by NTP (1995) was associated with CPN, and not a direct effect of t ert

- butanol exposure. However, there was a strong, statistically

- s ignificant, treatment

- related, dose

- response relationship between chronic tert

- butanol exposure and increased incidence of transitional epithelial hyperplasia in both male an d female rats
in the NTP (1995) study. The severity of CPN also increase d with tert

- butanol exposure, although
the dose -r e sponse relationship in males was very weak (only a 10% increase in mean severity at the highest dose). T he very different dose

- response relationships argue against a close association

. Moreover, even if transitional epithelial h yperplasia were associated with CPN, there is no evidence
to support that th e effect is independent of tert

- butanol treatment, given the robust dose

- response relationships. T herefore, the data are insufficient to dismiss transitional epithelial hyperplasia as causally related to tert

- b utanol exposure. Additionally, there have been a few research groups who have discussed the role of CPN and

α 2u - globulin accumulation on the renal tumors observed in male rat s exposed to tert

- butanol. Cruzan et al. (2007) concluded that α 2u - globulin, exacerbation of CPN, or a combination of both were the MOAs for the kidney tumors in males. Hard et al. (2011) also concluded that both α 2u globulin- i nduced nephropathy and exacerbated CPN were MOAs for the kidney tumors observed in
the male rats in the 2

- ye ar drinking study conducted by NTP (1995)

. However, t he underlying mechanisms regulating CPN and its exacerbation are n ot well understood, and to date, there is no
scientific consensus on the relevance of CPN in rats to human health hazard ( Melnick et al., 2012 ; Hard et al., 2009 ) . Moreover, no key events for the exacerbation of CPN have been identified, so no MOA analysis can be performed under the EPA Cancer Guidelines MOA framework ( U.S. EPA,
2005a ). Therefore, kidney effects from tert

- butanol exposure associated with CPN are consider ed
relevant to humans.

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S ummary of kidney toxicity 

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37 Kidney toxicity was consistently observed after oral exposure in two strains of rats and in
one strain of mice and in both sexes

. A bsolute and relative kidney weights also were increased in
male and f emale rats in both the 13

- week and 2

- year studies. In male and female rats, histopathological examination of the kidneys indicated kidney lesions exhibiting a dose

- r esponse trend, i ncreased incidence of nephropathy after 13 weeks and 2 years

, and increa sed transitional epithelium hyperplasia a nd suppurative inflammation ( females only ) after 2 years. In mice

, the only kidney effect observed was an increase in kidney weight (absolute and/or relative) in both

sexes o f mice in the 13

- week study. Organs wer e not weighed in the 2

- year mouse study

, so no determination can be made. Furthermo re , t here were no treatment

- related histopathological lesions in the kidneys of mice at 13 weeks or 2 year s

. Male rats are more sensitive to the kidney effects of t ert

- bu tanol

, and t he available data
indicate that male rats accumulate α 2u - globulin in the kidney, which is a specific MOA for male rats. MOA analysis determined that the renal tumors o bserved in male rats are mediated by other processes besides α 2u -globulin. Therefore, in the absence of a known MOA, EPA considers the male and female kidney effects observed in experimental animals to be relevant to assessing human
health hazard.
EPA identified kidney e ffects as a human hazard of tert

- butanol exposure. Data on kidney
tumors associated with tert

- b utanol exposure are discussed as part of the overall weight of evidence for carcinogenicity in Section 1.2.2.

**1.1.2. T**  **hyroid Effects** 

Synthesis of Effects in Thyroid This section reviews the studies that investigated whether exposure to tert

- b utanol can cause thyroid effects in humans or animals. T he data base examining thyroid effects following

tert

- b utanol exposure contain s no human data and two chronic studies ( one in rats and one in

mice )

. Studies employing short term and acute exposures that examined thyroid effects are not included in the evidence tables; however, they are discussed in the text if they provide data to support mode of action or hazard identificati o n. No methodological concerns were identified that would lead one or more studies to be considered less informative for assessing human health
hazard.
A 2

- y ear inhalation study is not available. Thyroid effects were not observed in studies in

rats ( NTP, 1995 ) . Thyroid toxicity was observed in mice of both sexes after 2 years of oral exposure
via dri nk ing water ( NTP, 1995 ) . Follicular cell hyperplasia, as well as follicular cell adenomas, was present in both male and female mice. The evidence was stronger in females due to the dose

related increase in follicular cell hyperplasia reaching statistical significance in the highest two doses, and the presence of a statistically significant increase i n follicular cell adenomas in the high

dose group. There was also a st atistically significant increase in follicular cell hyperplasia at all

This document is a draft for review purposes only and does not constitute Agency policy. 1- 27 DRAFT

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- Butyl Alcohol

d oses in male mice, but only a marginal increase in follicular cell adenomas in the mid

- dose group. 

### 1
2
3
4
5
6 One high -do se male mouse developed a follicular cell carcinoma. The lower tumor incidence in males may be due to the increased mortality seen in the high

- dose group. NTP (1995) noted that thyroid follicular cell tumorigenesis follows a progression from hyperplasia to adenoma and carcinoma, suggesting that hyperplasia is a preneoplastic lesion in th e thyroid

.

This document is a draft for review purposes only and does not constitute Agency policy. 1- 28 DRAFT

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- Butyl Alcohol

**T able**  **1-7 . Evidence pertaining to thyroid effects in animals following oral**  

### 1 2 **exposure to** tert

**- butanol** 

Reference and s tudy design Results Follicular cell hyperplasia
NTP (1995)

F344/N rat; 60/sex/treatment (10/sex/treatment evaluated at 15 months) Drinking water ( 0, 1.25, 2.5, 5, or 10 mg/mL) M: 0, 90, 200, or 420 a mg/kg -d F: 0, 180, 330, or 650 a mg/kg -d 2 years Incidence b

Males Females

Dose (mg/kg

- d) Follicular cell hyperplasia Dose (mg/kg

- d) Follicular cell hyperplasia

0 3/50 0 0/50

90 0/49 180 0/50

200 0/50 330 0/50

420 0/50 650 0/50
NTP (1995)

B6C3F 1 m ouse; 60/sex/treatment Drinking water ( 0, 5, 10, or 20 mg/mL ) M: 0, 540, 1,040, or 2,070 a mg/kg -d F: 0, 510, 1,020, or 2,110 mg/kg -d 2 years Incidence (severity)

Males Females

Dose (mg/kg

- d) Follicular cell hyperplasia Dose (mg/kg

- d) Follicular cell hyperplasia

0 5/60 (1.2) 0 19/58 (1.8)

540 18/59* (1.6) 510 28/60 (1.9)

1,040 15/59* (1.4) 1,020 33/59* (1.7)

2,070 18/57* (2.1) 2,110 47/59* (2.2)

Tumors
NTP (1995)

F344/N rat; 60/sex/treatment (10/sex/treatment evaluated at 15 months) Dri nking water ( 0, 1.25, 2.5, 5, or 10 mg/mL) M: 0, 90, 200, or 420 a mg/kg -d F: 0, 180, 330, or 650 a mg/kg -d 2 years Incidence b

Dose (mg/kg

- d) Follicular cell adenoma Follicular cell carcinoma

Male

0 2/50 2/50

90 0/49 0/49

200 0/50 0/50

420 0/50 0/50

Female

0 1/50 1/50

180 0/50 0/50

330 1/50 1/50

650 0/50 0/50


### 3

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- Butyl Alcohol

**Table 1** 

**- 7**  **. Evidence pertaining to thyroid effects in animals following oral exposure to**  tert

**- butanol**  **(** continued **)**

Reference and s tudy design Results
NTP (1995)

B6C3F 1 mouse; 60/sex/treatment Drinking water ( 0, 5, 10, or 20 mg/mL ) M: 0, 540, 1,040, or 2,070 a mg/kg -d F: 0, 510, 1,020, or 2,110 mg/kg -d 2 years Incidence

Dose (mg/kg

- d) Follicular cell adenoma Mortality

adjusted rate (%) Follicular cell carcinoma or adenoma Mortality

adjusted

rate (%) Follicular cell carcinoma c

Male

0 1/60 3.6 1/60 3.6 0/60

540 0/59 0.0 0/59 0.0 0/59

1,040 4/59 10.1 4/59 10.1 0/59

2,070 1/57 5.9 2/57 8.7 1/57

Female

0 2/58 5.6 2/58 5.6 0/ 58

510 3/60 8.6 3/60 8.6 0/ 60
1,020 2/59 4.9 2/59 4.9 0/ 59
2,110 9/59* 19.6 9/59* 19.6 0/ 59 

### 1 2 3 4 5 6 a There was a significant decrease in survival in the high

- dose group.
b Results do not include the animals sacrificed at 15 months.

c Mortality

- adjusted rates were not calculated by study authors for follicular cell carcinoma. * Statistically significant p ≤ 0.05 as determined by the study authors. Conversions from drinking water concentrations to mg/kg

- d per formed by study authors.

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- Butyl Alcohol



### 1 2

3 4  = exposures at which the endpoint was reported statistically significant by study authors  = exposures at which the endpoint was reported not statistically significant by study authors

Hyperplasia; M mouse

Hyperplasia; F mouse

Hyperplasia; M rat

Hyperplasia; F rat

Adenoma; M mouse

Adenoma; F mouse

Adenoma; M rat

Adenoma; F rat NONCANCER

CANCER

10 100 1,000 10,000 

### Dose (mg/kg

- day)

This document is a draft for review purposes only and does not constitute Agency policy. 1- 31 DRAFT

— DO NOT CITE OR QUOTE Source: NTP (1995)
**Figure 1-5.**  **Exposure**

**- response array of**  **thyroid follicular cell effects following**  **chronic or**  **al exposure to** tert

**- butanol** 

**.**Toxicological Review of tert

- Butyl Alcohol

M

ode of Action Analysis

— Thyroid E

ffects 

### 1

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7
8
9
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19

20

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29
30

31

32

33
34
35
36 There are inadequate data to determine the MOA for tert

- butanol

- induced thyroid follicular cell lesions. The mechanism of formation of these lesions resulting from t ert

- butanol exposure has not been spe c ifically studied; however, Blanck et al. (2010) conducted a short

- term study examining the hepatic and thyroid effects of tert

- bu tanol exposure to provide additional data on the thyroid tumors observed in the chronic NTP ( 1995) study. tert

- Butanol did not have any effect on liver weight when compared to the control group, but the livers were visibly enlarged

, in some cases accompanied by centrilobular hepatocellular hypertrophy

, i n some treated groups. There

were no trea tment

- r elated histological alterations in the thyroid in tert

- butanol treated mice. Only a slight st atistically nonsignificant increase in thyroid stimulating hormone (TSH) was observed after
3 d ays of exposure, but both thyroxine (T 4 ) and tri i odothyronine (T 3 ) levels were decreased.
S u stained increases in TSH, resulting in s ustained thyroid follicular cell proliferation

, may eventually result in progression of hyperplasia to adenoma and carcinoma ( U.S. EPA, 1998a ) . Based on alterations in hepatic phase I and phase II enzyme activities and gene expression, the data from Blanck et al. (2010) suggest a possible role for increased thyroid hormone clearance in the liver in

tert

- b utanol

- induced thyroid tumors. The available support for this hypothesis, however, is weak. In particular, Blanck et al. (2010) did not find any significant changes in TSH levels, though the study duration was short (≤14 days) , an d there are no data regarding thyroid cell proliferation after
exposure to tert

- bu tanol

.
Summary of thyroid toxicity EPA identified thyroid effects as a potential human hazard of t ert

- butanol exposure

. The thyroid endpoints reported following chronic exposure to tert

- b utanol include follicular cell hyperplasia, follicular cell adenoma, and follicular cell carcinoma. Tog et her with the evidence of significantly increased incidence of thyroid follicular cell adenomas in high

- d ose females, these observations support the finding that the increased hyperplasia in male mice is a preneoplastic effect rather than an adaptive resp o nse

. There is no conclusive MOA for the development of thyroid follicular cell adenomas, although there is some evidence supporting greater clearance of thyroid hormones by the liver causing continual secretion of TSH by the pituitary leading to follicul a r cell hyperplasia and tumors. D ata on thyroid tumors associated with tert

- butanol exposure are discussed further as part of the overall weight of evidence for carcinogenicity in Section 1.2.2.

**1.1.3. Re**  **productive and Developmental Effects** 

Synthesis of reprodu c tive and developmental toxicity This section reviews the studies that investigated whether exposure to tert

- b utanol can cause reproductive and developmental effects in humans or animals. Th is section contain s information on reproductive effects, systemi c developmental effects, as well as neurodevelopmental effects f ollowing tert

- butanol exposure

. The data base examining reproductive

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- Butyl Alcohol

o r developmental effects following tert

- butanol exposure contains no human data and 7 studies 

### 1
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25

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27
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32
33
34 performed in rats and mice. Three studies evaluating reproductive effects includ ed changes in reproductive organs (one on e- generation reproductive study and two subchronic stud ies ) . In addition, there was one two

- year oral study in rats and mice that evaluated reproductive histopath ology and did not find any treatment

- related effects. The studies selected for discussion below exposed animals via oral gavage, drinking water, and inhalation for ≥ 63 days

. The collection of reproductive studies o n tert

- butanol is limited by the absence of any two

- generation reproductive oral or inhalation studies

. Four studies evaluated developmental effects (three developmental studies and one one

- generation reproductive study). As mentioned in the Study Selection, the studies selected for discussion b elow exposed animals to tert

- butanol via inhalation, gavage, and drinking water. No methodological concerns were identified that would lead one or more studies to be considered less informative for assessing human health hazard , but the Faulkner et al. (1989) study did not provide results in the dam that could be used to adequately determine if
fetal effects occurred d ue to maternal toxicity. Three studies evaluated neurodevelopmental effects following t ert

- butanol exposure in rats and mice. These studies selected for discussion below exposed animals v ia liquid diet and inhalation

. The collection of neurodevelopmental studies on

tert

- b utanol is limited in that all studies were conducted prior to Developmental Toxicity Guidelines being available from the U.S. EPA (1991b) and OECD; as such, there are study design
considerations for each of the studies. Daniel and Evans (1982) had a small number of animals per treatment group, presentation of results provided limited use of the data with no comparisons to controls, an d there was no long term neurodevelopmental testing. Nelson et al. (1991) evaluated neurodevelopmental effects after either paternal or maternal exposure . Although th e study authors
used two different exposure c oncentrations, the exposures were not run concurrently nor was there information provided on exposure methods to indicate the studies were conducted similarly.

Reproductive effects

R eproductive endpoints , such as sex organ weights, estrous cycle length, and sperm effects

wer e ex amined following either oral or inhalation exposure in three subchronic studies ( Lyondell Chemical Co., 2004 ; NTP, 1997
, 1995 )( Table 1-8 ; Figure 1-6 ; Figure 1-7 ). The only reproductive
effect noted was an increase in e strous cycle length . Estrous cyc le length was increased (28%
increase, p < 0.01) in female mice orally exposed to 11,620 mg/kg

- day. No significant changes in estrous cycle length were observed following oral exposure in rats, or inhalation exposure in mice

or rats.

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- Butyl Alcohol

**T able**  **1-8 . Evidence pertaining to**  **r eproductive effects**  **in animals**  **following** 

### 1 2 **exposure to** tert

**- butanol** 

Reference and s tudy design Results
Male reproductive effects

Lyondell Chemical Co. (2004)

Sprague

- Dawley rat; 12/sex/treatment Gavage 0, 64, 160, 400, or 1,000 mg/kg -d F0 males: 9 weeks beginning 4 weeks prior to mating F0 females: 4 week s prior to mating through PND21 F0 reproductive effects
No significant effect on weights of male reproductive organs or sperm observed

NTP (1995)

F344/N rat; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, or 40 mg/mL ) M: 0, 230, 490, 840, 1,520, 3,610 a mg/kg -d F: 0, 290, 590, 850, 1,560, 3,620 a mg/kg -d 13 weeks No significant effect on weights of male reproductive organs or sperm observed

NTP (199 5)

B6C3F 1 mouse; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, or 40 mg/mL ) M: 0, 350, 640, 1,590, 3,940, 8,210 a mg/kg -d F: 0, 500, 820, 1,660, 6,430, 11,620 a mg/kg -d 13 weeks No significant effect on weights of male reproductive organs or sperm observed

NTP (1997)

F344/N rat; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1

, 643, 3,273 or 6,368 mg/m 3 ) (dynamic whole body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical concentration and method were reported No significant effect on weights of male reproductive organs or sperm observed

Evaluations were performed only for concentrations ≥542 ppm (1,643 mg/m 3 )

NTP (1997)

B6C3F 1 mous e; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1,643, 3,273 or 6,368 mg/m 3 ) (dynamic whole body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical concentration and method were reported No significant effect on weights of male reproductive organs or sperm observed

Evaluations were performed only for concentrations ≥542 ppm (1,643 mg/m 3 )



### 3

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- Butyl Alcohol

**Table 1** 

**- 8**  **. Evidence pertaining to r**  **eproductive effects in**  **animals following**  **exposure to** tert

**- butanol**  **(** continued **)**

Reference and s tudy design Results
Female reproductive effects

Lyondell Chemical Co. (200 4)

Sprague

- Dawley rat; 12/sex/treatment Gavage 0, 64, 160, 400, or 1,000 mg/kg -d F0 males: 9 weeks beginning 4 weeks prior to mating F0 females: 4 weeks prior to mating through PND21 Pregnancy index 91.7% 91.7% 100% 100% 91.7%

NTP (1995)

F344/N rat; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, or 40 mg/mL ) M: 0, 230, 490, 840, 1,520, 3,610 a mg/kg -d F: 0, 290, 590, 850, 1,560, 3,620 a mg/kg -d 13 weeks No significant effect on female estrous cycle (0, −2, −4, 0, +8 % change relative to control)

NTP (1995)

B6C3F 1 mouse; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, or 40 mg /mL) M: 0, 350, 640, 1,590, 3,940, 8,210 a mg/kg -d F: 0, 500, 820, 1,660, 6,430, 11,620 a mg/kg -d 13 weeks ↑ length of estrous cycle Response relative to control : 0, +5, +5, +5, +6, +28*%

NTP (1997)

F344/N rat; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1

, 643, 3,273 or 6,368 mg/m 3 ) (dynamic whole body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical concentration and method were reported No significant effect on female estrous cycle (0, −4, +2, +4 % change relative to control)

Evaluations were performed only for concentrations ≥542 ppm (1,643 mg/m 3 )

NTP (1997)

B6C3F 1 mouse; 10/sex/treatment Analytical conce ntration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1,643, 3,273 or 6,368 mg/m 3 ) (dynamic whole body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical concentration and method were reported No significant effect on female estrous cycle (0, −3, −9, −5 % change relative to control)

Evaluations were only performed for concentrations ≥542 ppm (1,643 mg/m 3 )



### 1 2 3 4 5 * Statistically significant p ≤ 0.05 as determined by the study authors. Conversions from drinking water concentrations to mg/kg

- d per formed by study authors. Conversion from ppm to mg/m 3 is 1 ppm = 3.031 mg/m 3

. Percentage change compared to control = (treated value – c ontrol value) ÷ control value × 100.

This document is a draft for review purposes only and does not constitute Agency po licy. 1- 35 DRAFT

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### 18
19
20
21
22
23
24 Toxicological Review of tert

- Butyl Alcohol

Developmental effects 

### 1

2
3
4
5
6
7
8
9
10
11
12
13
14
15
16
17 Data from three developmental studies (two oral, one inhalation) suggest that fetal effects are generally observed at doses that cause toxicity in the dams as measured by clinical signs (e.g., decreased body weight gain, and/or food consumption ) ( T able 1-9 ; Figure 1-6 ; Figure 1-7 ). Developmental effects of te rt

- butanol observed after oral exposure (liquid diets or gavage)
in several mouse st rains and one rat strain include measures of fetal loss or viability (e.g., in creased number of resorptions, decreased n umbers of neonates per litter) and decreased fetal body weight ( Lyondell Chemical Co., 2004 ; Faulkner et al., 1989 ; Daniel and Evans, 1982 ). Daniel a nd Evans (1982) also observed decreases in body weight gain during PND 2

– 10; however, data suggest that the effect may be due to maternal behavior or nutritional status. In addition, one study reported
increased incidence of variations of the skull or s ternebrae in two mouse strains, but the difference was not statistically significant ( Faulkner et al., 1989 ) . Similar developmental effects were also
observed after whole

- bo dy inhalation exposure in Sprague

- Dawley rats for 7 hours/day on GD s 1–
19 ( Nelson et al., 1989 ) . Fetal effects included con centration

- related reductions in body weight in
male and female fetuses an d higher incidence of skeletal variations when analyzed on the basis of
individual fetuses ( bu t not on a per litter basis ) . In contrast to the oral exposure studies in mice
and rats , h owever, there was no effect on measures of fetal loss.

This document is a draft for review purposes only and does not constitute Agency policy. 1- 36 DRAFT

— DO NOT CITE OR QUOTEToxicological Review of tert

- Butyl Alcohol

**T able**  **1-9 . Evidence pertaining to**  **developmental effects**  **in animals**  **following** 

### 1 2 **exposure to** tert

**- butanol** 

R eference and study design Results

Daniel and Evans (1982)

Swiss Webster (Cox) mouse; 15 pregnant dams/treatment Liquid diet (0, 0.5, 0.75, 1.0%, w/v) 0 (isocaloric amounts of maltose/dextrin), 3,324, 4,879, 6,677 mg/kg -d GD 6 –20 No statistical analysis was conducted on any of these data

Maternal
Percent change compared to control:

Dose (mg/kg

- d) Food consumption (mean g/animal/day) Body weight gain Number of litters (% pregnant dams)

0 0 0 11 (77%)

3,324 +2

− 3 12 (80%)

4,879

− 3

− 19 8 (53%)

6,677

− 4

− 20 7 (47%)

Authors note that lower food consumption in higher tert

- butanol dose groups reflects problems with pair feeding and maternal sedation.

Fetal

Percent change compared to control:

Dose (mg/kg

- d) Number of neonates/litter Fetal body weight on PND 2

0 0 0

3
, 324

− 1

− 7

4,879

− 29

− 19

6,677

− 49

− 38

Number of stillborn also increased with dose (3, 6, 14, and 20, respectively), but the number of stillborn per litter was not provided. The high dose also caused a delay in eye opening and a lag in weight gain during PND 2

– 10 (information was only provided in text or figures)


### 3

This document is a draft for review purposes only and does not constitute Agency policy. 1- 37 DRAFT

— DO NOT CITE OR QUOTEToxicological Review of tert

- Butyl Alcohol

**Table 1** 

**- 9**  **. Evidence pertaining to developmental**  **effects in**  **animals following**  **exposure to** tert

**- butanol**  **(** cont inued **)**

R eference and study design Results

Faulkner et al. (1 989)

CBA/J mouse; 7 pregnant females in control, 12 pregnant females in treated Gavage ( 10.5 mmoles/kg twice a day) ; 0 (tap water), 1,556 mg/kg

- d GD 6

– 18 Maternal results not reported.

Fetal
Percent change compared to control: Incidence:

Dose (mg/kg

- d) L ive fetuses/litter Fetal weight Stern ebr al variations Skull variations

0 0 0 4/28 1/28

1,556

− 41 *

− 4 7/30 3/30

S ternal variations: misaligned or unossified sternebrae S kull variations: moderate reduction in ossification of supraoccipital bone
Number of total resorptions (10 resorptions/66 implants in controls, 37/94 implants in treated) and resorptions per litter ( +118%) increased ( p < 0.05)
Faulkner et al. (1989)

C57BL/6J mouse; 5 pregnant females in controls, 9 pregnant females treated Gavage ( 10.5 mmoles/kg twice a day)
0 (tap water), 1,556 mg/kg

- d GD 6

– 18 Maternal results not reported.

Fetal

Percent change compared to control: Incidence:

Dose (mg/kg

- d) L ive fetuses/litter Fetal weight Sternal variations Skull variations

0 0 0 5/21 1/21

1,556

− 58%*

− 4 9/16 7/16

Sternal variations: misaligned or unossified sternebrae Skull variations: moderate reduction in ossification of supraoccipital bone
Number of total resorptions (4 resorptions/44 implants in controls, 38/68 implants in treated) and resorptions per litter ( +428%) increased ( p < 0.05)

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

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- Butyl Alcohol

**Table 1** 

**- 9**  **. Evidence pertaining to developmental**  **effects in**  **animals following**  **exposure to** tert

**- butanol**  **(** cont inued **)**

R eference and study design Results

Lyondell Chemical Co. (2004)

OECD guideline 421 study:

Sprague

- Dawley rat; 12/sex/treatment Gavage 0, 64, 160, 400, or 1,000 mg/kg

- d F0 males: 9 weeks beginning 4 weeks prior to mating F0 females: 4 weeks prior to mating through PND21 F1 Males and Females: 7 weeks (throughout gestation and lactation; 1 male and 1 female from each litter was dosed directly from PND 21

- 28) Response relative to control

Dose (mg/kg

- d) 0 64 160 400 1000

Maternal effects
Body weight gain GD 0

- 20

0

− 3

− 4 0

− 16*

Food consumption GD 0

- 20

0 0 0 +4 0
Body weight gain PND 1

- 21

0 +3

− 10 +3 +100*

Food consumption LD1

- 14

0

− 2

− 6 0

− 16
Live pups/litter response relative to control

0

− 9

− 11

− 7

− 33*
Dams dosed with 1000 or 400 mg/kg/d showed CNS effects (e.g., ataxia, lethargy) which became undetectable by 4

- weeks of exposure in animals exposed to 400 mg/kg/d but not those in the higher dose group.

F1 effects

Viability index (pup survival to PND4)

96.4% 98.7% 98.2% 99.4% 74.1%*

Lactation index (pup survival to PND21)

100% 100% 100% 99.2% 98.8%
Sex ratio (% males)

54.4 52.3 50.9 53.4 52.1
Pup weight/ litter PND 1 relative to control (%)

0 +6 +4 +7

− 10
Pup weight PND 28 relative to control (%)

M: 0 +2 0 0

− 12*

F: 0 0

− 4

− 2

− 8
Nelson et al. (1989)

Sprague

- Dawley rat; 15 pregnant dams/treatment Maternal: Unsteady gait (no statistical tests reported) , dose

- dependent ↓ in body weight gain (results presented in figure only), dose

- dependent ↓ in food consumption ranging from 7

– 36% depending on dose and time

Fetal

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

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- Butyl Alcohol

**Table 1** 

**- 9**  **. Evidence pertaining to developmental**  **effects in**  **animals following**  **exposure to** tert

**- butanol**  **(** cont inued **)**

R eference and study design Results

Analytical concentration: 0, 2,200, 3,510, 5,03 0 ppm (0, 6,669, 10,640, 15,248 mg/m 3 ), (dynamic whole body chamber) 7 hr/d
GD 1 –19 Generation method, analytical concentration and method were reported Percent change compared to control:

Dose (mg/m 3 ) Number of live fetuses/litter Resorptions per litter

0 0 0

6,669 0 +9

10,640 +15

− 18

15,248 +8 0
Percent change compared to control: Incidence:

Dose (mg/m 3 ) Fetal weight (males) Fetal weight (females) Skeletal variation by litter Skeletal variation by fetus

0 0 0 10/15 18/96

6,669

− 9*

− 9* 14/17 35/104

10,640

− 12*

− 13* 14/14 53/103*

15,248

− 32*

− 31* 12/12 76/83*

Skeletal variation by litter refers to the number of variations observed in the number of litters examined. Skeletal variation by fetus refers to the number of variations observed in the total number of fetuses examined. Fetuses are not categorized by litter. 

### 1 2 3 4 5

6

7

8

9 * Statistically significant p ≤ 0.05 as determined by study authors. Conversions from diet concentrations to mg/kg

- d p erformed by study authors. Conversion from ppm to mg/m 3 is 1 ppm = 3.031 mg/m 3

. Percentage change compared to control = (treated value – control value) ÷ control value × 100.

This document is a draft for review purposes only and does not constitute Agency policy. 1- 40 DRAFT

— DO NOT CITE OR QUOTEToxicological Review of tert

- Butyl Alcohol

This document is a draft for review purposes only and does not constitute Agency policy. 1- 41 DRAFT

— DO NOT CITE OR QUOTE 

### 1 2

3 4 Sources: (A) Faulkner et al. (1989) ; (B) Lyondell Chemical Co. (2004) ; (C) NTP (1995)
**Figure 1-6.**  **Exposure**

**- response array of reproductive and developmental**  **effects following oral exposure to** tert

**- butanol** 

**.**Toxicological Review of tert

- Butyl Alcohol



### 1

2 3

4 5 6

This document is a draft for review purposes only and does not constitute Agency policy. 1- 42 DRAFT

— DO NOT CITE OR QUOTE Sources: (A) NTP (1997) ; (B) Nelson et al. (1989)
**Figure 1-7**  **. Exposure**

**- response array of**  **reproductive and developmental**  **effects following inhalation exposure to** tert

**- butanol** 

**.**Toxicological Review of tert

- Butyl Alcohol

Neurodevelopmental Effects 

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2 6 In addition to the developmental effects noted above, neurodevelopmental effects also have
been observed. T h is includes changes in rotaro d performance following oral or inhalation

exposures , as well as decreases in open field behavior and cliff avoidance following oral exposure

, and reduced time hanging on wire after inhalation exposure during gestation ( T able 1- 10 ; Figure
1-6 ; Figure 1-7 ).

Rot a ro d performance Looking across studies, not all the findings were consistent. Whi le Daniel and Evans (1982)
found decreased rot a r od performance in mouse pups of dams orally exposed during gestation, Nelson et al. (1991) observed an increase in rot a rod performance in rat pups of dams exposed via inhalation during gestation.

Neurochemical measurements
In addition to behavioral effects, o ne stud y evaluated biochemical or physiological changes in the brain of offspring exposed during gestation or early in the postnata l period. Nelson et al. (1991) found statistically significant changes in neurochemical measurements in the brain in offspring of dams exposed via inhalation during gestation ; h owever, the two concentrations tested

were not run concurr e ntly

, and very little data were provided.

Physiological and psychomotor development

D at a also suggest that neurodevelopmental effects were not solely due to in utero exposure ( Daniel and Evans, 1982 )

. Daniel and Evans (1982) cross

- fostered half of the mouse pups born to treated mother with untreated surrogate females to test the effects of maternal nutrition and behavioral factors on the pups ’ p hysiological and psych omotor development. Results indicated that pups fostered with control dams performed significantly better than those maintained with treated dams ( T able 1- 10 ) ( Daniel and Evans, 1982 ) . Results were only presented in figures and were not compared with controls.

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- Butyl Alcohol

**T able**  **1- 10**  **. Evidence pertaining to neurodevelopmental**  **effects in**  **animals** 

### 1 2 **following exposure to** tert

**- butanol** 

Reference and s tudy design Results

Daniel and Evans (1982)

Liquid diet (0, 0.5, 0.75, or 1.0%, w/v); GD6 –20; Swiss Webster (Cox) mouse; 15 pregnant dams/treatment; after birth half the pups were nursed with their treated dams and the other half were fostered by untreated dams who recently gave birth 0 (isocaloric amounts of maltose/dextrin), 3,324, 4,879, or 6,677 mg/kg

- d • a dose

- dependent increase rig hting reflex time, with more time needed in animals maintained with maternal dams

• a dose

- dependent decrease in open field behavior

, with less activity in pups maintained with maternal dams

• a dose

- dependent decrease in rot a rod performance with the pups from maternal dams having lower performances

• a dose

- dependent decrease in the amount of time the pups were able to avoid a cliff

, with animals maintained with their maternal dams having less avoidance time

Nelson et al. (1991)

Sprague

- Dawley rat; 15 pregnant dams/treatment Analytical concentration: 0, 6,000, or 12,000 mg/m 3 ; (dynamic whole body chamber) 7 hr/d
GD 1 –19 Generation method, analytical concentration and method were reported Data were not presented specifically by dose nor were any tables or figures of the data provided

Maternal toxicity was noted by decreased food consumption and body weight gains

Results in offspring

• increase in rot a rod performance in high

- dose group (16 versus 26 revolutions/min for controls and 12,000 mg/m 3 animals, respectively)

• decreased time held on wire in the performance ascent test in the low

- dose group (16 sec versus 10 sec for controls and 1,750 mg/m 3 animals, respectively)

The following differences in neurochemical measurements in the brain between control and treated offspring were observed
,

• 53% decrease in norepinephrine in the cerebellum at 12,000 mg/m 3

• 57% decrease in met

- enkephalin in the cerebr um at 12,000 mg/m 3 and 83% decrease at 6,000 mg/m 3

• 61% decrease in β

- endorphin in the cerebellum at 12,000 mg/m 3

• 67% decrease in serotonin in the midbrain at 6,000 mg/m 3



### 3

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- Butyl Alcohol

**Table 1** 

**- 10** 

**. Evidence pertaining to**  **neurodevelopmental effects**  **in animals**  **following exposure to** tert

**- butanol**  **(** continued **)**

Reference and s tudy design Results

Nelson et al. (1991)

adult male Sprague

- Dawley rats (18/treatment) mated to untreated females

Analytical concentration: 0, 6
, 000, or 12,000 mg/m 3 ; (dynamic whole body chamber) 7 hr/d for 6 wk Generation method, analytical concentration and method were reported Data were not presented specifically by dose nor were any tables or figures of the data provided

Results (generally only specified as paternally treated versus controls) in offspring indicate

• increase in rot a rod performance (16 versus 20 revolutions/min for controls and 12,000 mg/m 3 animals, respectively)

• decreased time in open field (less time to reach the outer circle of the field, 210 sec versus 115 seconds for controls and 12,000 mg/m 3 animals, respectively) The following differences in neurochemical measurements in the brain between control and treated offspring were observed

• 39% decrease in norepinephrine in the cerebellum at 12,000 mg/m 3

• 40% decrease in met

- enkephalin in the cerebrum at 12,000 mg/m 3 and 75% decrease at 6,000 mg/m 3

• 71% decrease in β

- endorphin in the cerebellum at 12,000 mg/m 3

• 47% decre ase in serotonin in the midbrain at 6,000 mg/m 3



### 1 2 3 4

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21 * Statistically significant p ≤ 0.05 as determined by study authors. Conversions from diet concentrations to mg/kg

- d p erformed by study authors. Percentage change compared to control = (treated value – c ontrol value) ÷ control value × 100.

Mechanistic Evidence

No mechanistic evidence is av ailable for reproductive or developmental effects , including neurodevelopmental effects

.
Summary of Reproductive and Developmental Toxicity

EPA concluded that th e evidence does not support reproductive effects as a potential
human hazard of tert

- butanol exposure

. There are no two

- generation reproductive studies available by oral or inhalation exposure. T wo oral exposure studies ( Lyondell Chemical Co., 2004 ; NTP, 1995 ) and one subchronic inhalation study ( NTP, 1997 ) are available. Overall, reproductive effects observed due to exposure to tert

- bu tanol were limited to altered length of estrous cycle ( NTP, 1995 ), but there is no available information to infer how th is effect may influence reproductive ability. EPA identified suggestive evidence of developmental effects as a potential human hazard of

tert

- b utanol exposure

. Exposure during gestation resulted in increased fetal loss, decreased fetal body weight, and possible increases in skeletal variations in exposed offspring or pups , although effects were not always consistent across exposure routes (oral and inhalation). Dam s had decreased body weight and/or body weight gains, decreased f ood consumption, and/or clinical signs of intoxication at the same doses that t ert

- butanol caus ed fetal effects. Neurodevelopmental

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- Butyl Alcohol

e ffects including decreased brain weight, changes in brain biochemistry, and changes in behavioral 

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33 performances have also been observed. Each of the neurodevelopmental studies, however, had limitations in the study design and/or reporting. In addition, results from the neurodevelopmental studies were not always consistent between studies or across dose.

**1.1.4. Carcinogenicity**  **(other than in the kidney or thyroid)**

Synthesis of C a rcinogenicity D ata ( O ther than in the K idney or T hyroid) This section reviews the studies that investigated whether exposure to tert

- b utanol can cause cancers (other than in the kidney or thyroid) in humans or animals. The data base e xamining carcinogenicity f ollowing tert

- butanol exposure contain s no human data and two chronic studies

, one in rats and one in mice

. As mentioned in the Study Selection, t he s tudies providing data on carcinogenicity e xposed animals via drinking water for ≥3 0 days. Shorter duration studies do not generally evaluate carcinogenicity, but any shorter duration studies that examined carcinogenicity
are discuss ed in the text if they provide data to support mode of action or hazard identification. No methodological concerns were identified that would lead one or more studies to be considered less informative for assessing human health hazard. Kidney a nd thyroid tumors are presented above with the specific organ hazard
identification. No other treatment

- r elated changed in tumors in other organs were noted in the 2

- year oral rat or mouse studies conducted by NTP (1995) , which evaluated a comprehensive set of tissues/organs. There is no 2

- y ear inhalation study

.

Mechanistic Evidence

Available m echanistic evidence was previously discussed in the context of kidney and thyroid tumors (Sections 1.1.1 and 1.1.2)

.

Summary o f Carcinogenicity Evidence

Th er e are limited data available on the carcinogeni city of tert

- butanol

. There are 2

- year oral studies in one strain of rats and one strain of mice, but no 2

- y ear inhalation studies. EPA identified suggestive evidence of kidney and thyroid tumors as a potential human hazard.

**1.1.5. O**  **ther Toxicological Effects**

Synthesis of O th er T oxicity D ata The database for effects other than kidney, thyroid, reproductive, developmental (including neurodevelopmental) and cancer contain only 14 rodent studies. As previously mentioned in the Study Selection, all selected studies employed inhalation, oral gavage, or drinking water exposures
for ≥ 3 0 days. Studies employing short term and acute exposures that examined other toxicological effects are not included in the evidence tables; however, they are discussed in the text if they

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- Butyl Alcohol

p rovide data to support mode of action or hazard identification

. No studies were removed for 

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34 methodological concerns.

tert

- B utanol also has been found to have CNS effects similar to ethanol in terms of animals appearing intoxicated and having wit h drawal symptoms after cessation of oral or inhalation exposure. Severity of CNS symptoms su ch as withdrawal increased with dose and duration of exposure. However, study quality concerns (e.g., short exposure durations, lack of data reporting,
small numbe r of animals per treatment group) associated with all of the studies in the database preclude a clear understanding of potential neurotoxicity following tert

- b utanol exposure, and therefore, CNS studies are not presented in evidence tables.
Effects in ot h er tissues were observed with less consistency. Th ese include decreased body weight, l iver effects, and u rinary bladder effects.

B ody weight Body weight was decreased by >10% in both rats and mice with subchronic and chronic exposure, with males generally more affected than females ( T able 1- 11 ). The concentrations used in the subchronic inhalation study did not decrease body weights. However, a short

- t erm (i.e., 18

- day) inhalation study in rats observed a >10% decrease in body we ight at concentrations about
three f old higher (in mg/m 3 ) than the highest concentration used in the subchronic study. The same concentrations did not have any effect on the body weight in mice with short

- t erm inhalation

exposure.

Liver effects Although s ome rodent studies observed statistically significant changes in relative liver weight with tert

- bu tanol exposure, absolute liver weight was significantly increased only in female rats after subchronic oral exposure ( T able 1- 12 ). The results pertaining to histopathology changes

were inconsistent ( T able 1- 13 ) . The oral NTP (1995) subchronic and chronic studies did not
observe treatment

- r elated effects on liver histopathology in both sexes of F344 rats, but in a 10

- week study in a different rat strain (Wistar rats), several liver lesions (including necrosis) and increased liver gly c ogen were seen in male rats (no females were included in the study) with the only dose used ( Acharya et al., 1997 ; Acharya et al., 1995 )

. T he study did n ot provide any incidence or severity data. The dose used in this study was in the range of t he lower doses used in the NTP (1995) study. In the developmental study by Lyondell Chemical Co. (2004) , the F1 SD rats treated by t ert

- butanol for at least 9 weeks did not show any liver effects. An increased incidence of fatty
liver was observed in the m ale mice of the highest dose group in the 2

- year mouse bioassay, but no histopathologic changes were seen in the subchronic mouse study. No changes in liver
hi s topathology were observed in the NTP (1997) subchronic inhalation study.

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- Butyl Alcohol

Urinary bladder effects 

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16 Several studies also reported effects in the urinary bladder ( Table 1- 14 ). Transitiona l epithelial hyperplasia was observed in m ale rats and mice after 13 weeks of exposure at doses of 3,610 mg/kg

- d ay (male rats) and ≥ 3940 mg/kg

- day (male mice). Male mice exposed at doses of 2,070 mg/kg

- d ay for 2 years of also exhibited transitional epithe lial hyperplasia. Neither female rats nor female mice showed increased incidences of transitional epithelial hyperplasia. Both sexes of mice demonstrated incidence of inflammation in the urinary bladder after both subchronic and chronic exposures, with a g r eater incidence in males compared to females.
An e xposure

- response array of these effects in body weight, liver, and urinary bladder is provided in F igure 1-8 and Figure 1-9 for oral and inhalation studies , respectively

.

Mechanistic Evidence

No mechanistic evidence is av ailable for these other toxicological effects

.
Summary of O th er T oxicity D ata

EPA concluded that t he evidence does not support body weight changes, liver effects, and urinary bladder effects as potential h uman hazard s of tert

- butanol exposure.

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- Butyl Alcohol

**T able**  **1- 11** 

**. Evidence**  **pertaining to e ffects on body weight**  **in animals**  

### 1 2 **following exposure to**  tert

**- butanol** 

Reference and s tudy design Results
Acharya et al. (1995)

Wistar rat; 5

– 6 males/treatment Drinking water (0 or 0.5 %), 0 or 575 mg/kg -d 10 weeks Body weight in treated animals lower than controls by ~7% (p< 0.05); (results only provided in a figure)

Lyondell Chemical Co. (2004)

Sprague

- Dawley rat; 12/sex/treatment Gavage 0, 64, 160, 400, or 1,000 mg/kg -d F0 males: 9 weeks beginning 4 weeks prior to mating F0 females: 4 weeks prior to mating through PND21 Percent change compared to control:

F0 Males F0 Females

Dose (mg/kg

- d) Body weight Dose (mg/kg

- d) Body weight

0 0 0 0

64

− 2 64 0

16 0

− 4 16 0

− 2

40 0 +2 40 0 +1
1,00 0

− 7 1,00 0 +4
NTP (1995)

F344/N rat; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, or 40 mg/mL) M: 0, 230, 490, 840, 1,520, 3,610 a mg/kg -d F: 0, 290, 590, 850, 1,560, 3,620 a mg/kg -d 13 weeks Percent change compared to control:

Males Females

Dose (mg/kg

- d) Body weight Dose (mg/kg

- d) Body weight

0 0 0 0

230

− 4 290 +2

490

− 5* 590 +1

840

− 12* 850 +1
1,520

− 17* 1,560

− 2
3,610 All dead 3,620

− 21*
NTP (1995)

B6C3F 1 mouse; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, or 40 mg/mL) M: 0, 350, 640, 1,590, 3,940, 8,210 a mg/kg -d F: 0, 500, 820, 1,660, 6,430, 11,620 a mg/kg -d 13 weeks Percent change compared to control:

Males Females

Dose (mg/kg

- d) Body weight Dose (mg/kg

- d) Body weight

0 0 0 0

350

− 1 500 +3

640 +1 820

− 1
1,590

− 4 1,660 +4
3,940

− 14* 6,430

− 6
8,210

− 24* 11,620

− 15*
High

- dose females had a significantly lower initial weight, but also had a significantly lower body weight gain indicating that there was some effect of
treatment 

### 3

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- Butyl Alcohol

**Table 1** 

**- 11**  **. Evidence pertaining to effects on body weight in animals following exposure to**  tert

**- butanol**  **(** continued **)**

Reference and s tudy design Results
NTP (1995)

F344/N rat; 60/sex/treatment (10/sex /treatment evaluated at 15 months) Drinking water ( 0, 1.25, 2.5, 5, 10 mg/mL)

M: 0, 90, 200, 420 a mg/kg

- d F: 0, 180, 330, 650 a mg/kg

- d 2 years Percent change compared to control:

Males Females

Dose (mg/kg

- d) Body weight Dose (mg/kg

- d) Body weight

0 0 0 0

90

− 15 180

− 2

200

− 18 330

− 5

420

− 24 650

− 21
Only animals that survived at the end of 2 years were evaluated for body weight. Note: statistical significance not determined by study authors. NTP (1995)

B6C3F 1 mouse; 60/sex/treatment Drinking water ( 0, 5, 10, 20 mg/mL) ;

M: 0, 540, 1,040, 2,070 a mg/kg

- d F: 0, 510, 1,020, 2,110 mg/kg

- d 2 years Percent change compared to control:

Males Females

Dose (mg/kg

- d) Body weight Dose (mg/kg

- d) Body weight

0 0 0 0

540 +1 510

− 2
1,040

− 2 1,020

− 3
2,070

− 1 2,110

− 12 Only animals that survived at the end of 2 years were evaluated for body weight. Note: statistical significance not determined by study authors. NTP (1997)

F344/N rat; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1
, 643, 3,273 or 6,368 mg/m 3 ) (dynamic whole body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical concentration and method were reported Percent change compared to control:

Concentration (mg/m 3 ) Male s Body weight Females Body weight

0 0 0

406

− 1

− 5

824

− 2

− 1
1,643 +2 0
3,273 +3 0
6,368 +2

− 3
NTP (1997)

B6C3F 1 mouse; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1,643, 3,273 or 6,368 mg/m 3 ) (dynamic whole body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical concentration and Percent change compared to control:

Concentration (mg/m 3 ) Males Body weight Females Body weight

0 0 0

406 +4 +3

824

− 2

− 3
1,643 0 +3

3,273

− 4

− 6

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- Butyl Alcohol

**Table 1** 

**- 11**  **. Evidence pertaining to effects on body weight in animals following exposure to**  tert

**- butanol**  **(** continued **)**

Reference and s tudy design Results
method were reported 6,368 0

− 8 

### 1 2 3 4 5 a There was a significant decrease in survival in the high

- dose group. * Statistically significant p ≤ 0.05 as determined by study authors. Conversion s from drinking water concentrations to mg/kg -d performed by study authors. Percentage change compared to control = (t reated value – control value) ÷ control value × 100.

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- Butyl Alcohol

**T able**  **1- 12** 

**. Changes in l**  **iver weight in**  **animals following exposure to**  

### 1 2 tert

**- butanol** 

Reference and s tudy design Results

Acharya et al. (1995)

Wistar rat; 5

– 6 males/treatment Drinking water (0 or 0.5%), 0 or 575 mg/kg -d 10 weeks No significant treatment

- related effects (results were only provided in a figure)

Lyondell Chemical Co. (2004)

Sprague

- Dawley rat; 12/sex/treatment Gavage 0, 64, 160, 400, or 1,000 mg/kg -d Males: 9 weeks beginning 4 weeks prior to mating Females: 4 week s prior to mating through PND21 Percent change compared to control:

Males Females

Dose (mg/kg

- d) Absolute weight Relative weight Dose (mg/kg

- d) Absolute weight Relative weight

0 0 0 0 0 0

64

− 1 0 64

− 4

− 4

16 0

− 3 +1 16 0

− 7

− 5

40 0

− 2

− 1 40 0 +2 +1

1,00 0 +8 +16* 1,00 0 +8 +3
NTP (1995)

F344/N rat; 10/sex/treat ment Drinking water ( 0, 2.5, 5, 10, 20, or 40 mg/mL) M: 0, 230, 490, 840, 1,520, 3,610 a mg/kg -d F: 0, 290, 590, 850, 1,560, 3,620 a mg/kg -d 13 weeks Percent change compared to control:

Males Females

Dose (mg/kg

- d) Absolute weight Relative weight Dose (mg/kg

- d) Absolute weight Relative weight

0 0 0 0 0 0

230

− 2 +4 290 +11* +9*

490 +1 +8* 590 +10* +9*

840 +5 +20* 850 +12* +11*

1,520 +8 +31* 1,560 +15* +16*

3,610 All dead All dead 3,620 +9* +41*
NTP (1995)

B6C3F 1 mouse; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, or 40 mg/mL) M: 0, 350, 640, 1,590, 3,940, 8,210 a mg/kg -d F: 0, 500, 820, 1,660, 6,430, 11,620 a mg/kg -d 13 weeks Percent change compared to control:

Males Females

Dose (mg/kg

- d) Absolute weight Relative weight Dose (mg/kg

- d) Absolute weight Relative weight

0 0 0 0 0 0

350 +2 +3 500

− 1

− 4

640

− 1

− 2 820

− 5

− 3

1,590

− 1 +5 1,660

− 8

− 9*

3,940 0 +14* 6,430

− 2 +6

8,210

− 16 +22* 11,620

− 6 +13*


### 3

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- Butyl Alcohol

**Table 1** 

**- 1**  **2**

**. Changes in l**  **iver weight**  **in animals following exposure**  **to** tert

**- butanol**  **(** continued **)**

Reference and s tudy design Results

NTP (1995)

F344/N rat; 60/sex/treatment (10/sex/treatment evaluated at 15 months) Drinking water ( 0, 1.25, 2.5, 5 or 10 mg/mL) M: 0, 90, 200, or 420 a mg/kg -d F: 0, 180, 330, or 650 a mg/kg -d 2 years Percent change compared to control:

Males Females

Dose (mg/kg

- d) Absolute weight Relative weight Dose (mg/kg

- d) Absolute weight Relative weight

0 0 0 0 0 0

90 +2 +7 180

− 14*

− 8

200 +8 +11 330

− 3

− 1

420 +1 +14* 650

− 6 +9*

Only animals sacrificed at 15 months were evaluated for organ weights. Organ weights were not measured in the 2

- year mouse study
NTP (1997)

F344/N rat; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1

, 643, 3,273 or 6,368 mg/m 3 ) (dynamic whole body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical concentration and method were reported Percent change compared to control:

Males Females

Concentration

(mg/ m 3 ) Absolute weight Relative weight Absolute weight Relative weight

0 0 0 0 0

406

− 8

− 8 0 +3

824

− 2

− 1 0 0

1,643 +1

− 1 +3 +2

3,273 +10 +7 +9 +9*

6,368 +5 +5 +4 +8*

NTP (1997)

B6C3F 1 mouse; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1,643, 3,273 or 6,368 mg/m 3 ) (dynamic whole body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical concentration and method were reported Percent change compared to control:

Males Females

Concentrati

on (mg/ m 3 ) Absolute weight Relative weight Absolute weight Relative weight

0 0 0 0 0

406

− 1 0 +1

− 4

824 +4 +9 +1 +5

1,643 +7 +5 +5 +1

3,273

− 8

- 2 +2 +9*

6,368 +5 +7 +8 +21* 

### 1 2 3 a The high

- dose group had an increase in mortality. * Statistically significant p ≤ 0.05 as determined by study authors. Conversions from drinking water concentrations to mg/kg

- d per formed by study authors.

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- Butyl Alcohol

Conversion from ppm to mg/m 3 is 1 ppm = 3.031 mg/m 3

. 

### 1 2 3

4 5 Percentage change compared to control = (treated value – control value) ÷ control v alue × 100.

**Table 1-**  **13**

**. Changes in l**  **iver histopathology**  **in animals following exposure**  **to** tert

**- butanol** 

Reference and s tudy design Results

Acharya et al. ( 1997 ; 1995 )

Wistar rat; 5

– 6 males/treatment Drinking water (0, 0.5%), 0, 575 mg/kg -d 10 weeks ↑ liver glycogen (~ 7 fold)*

↑incidence of centrilobular necrosis, vacuolation of hepatocytes, loss of hepatocyte architecture, peripheral proliferation, and lymphocyte infiltration (incidences and results of statistical tests not reported)
Lyondell Chemical Co. (2004)

Sprague

- Dawley rat; 12/sex/treatment Gavage 0, 64, 160, 400, or 1,000 mg/kg -d

Males: 9 weeks beginning 4 weeks prior to mating Females: 4 weeks prior to mating through PND21 No treatment

- related effects observed.

NTP (1995)

F344/N rat; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, or 40 mg/mL) M: 0, 230, 490, 840, 1,520, 3,610 a mg/kg -d F: 0, 290, 590, 850, 1,560, 3,620 a mg/kg -d 13 weeks H istopathology data for the 13

- week study were not provided
, but the liver was evaluated indicating that no changes in liver histopathology were observed in the 13

- week study.

NTP (1995)

B6C3F 1 mouse; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, 40 mg/mL) M: 0, 350, 640, 1,590, 3,940, 8,210 a mg/kg -d F: 0, 500, 820, 1,660, 6,430, 11,620 a mg/kg -d 13 weeks H istopathology data for the 13

- week study were not provided , but the liver was evaluated indicating that no changes in liver histopathology were observed in the 13

- week study.

NTP NTP (1995)

F344/N rat; 60/sex/treatment (10/sex/treatment evaluated at 15 months) Drinking water ( 0, 1.25, 2.5, 5, 10 mg/mL) M: 0, 90, 200, or 420 a mg/kg -d F: 0, 180, 330, or 650 a mg/kg -d 2 years No treatment

- related effects observed.



### 6

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**Table 1** 

**- 1**  **3**

**. Changes in l**  **iver histopathology**  **in animals following exposure**  **to** tert

- **butanol (**  continued **)**

Reference and s tudy design Results

NTP (1995) B6C3F 1 mouse; 60/sex/treatment Drinking water ( 0, 5, 10, 20 mg/mL) M: 0, 540, 1,040, or 2,070 a mg/kg

- d F: 0, 510, 1,020, or 2,110 mg/kg -d 2 years Males Females

Dose (mg/kg

- d) Incidence of fatty change Dose (mg/kg

- d) Incidence of fatty change

0 12/59 0 11/60

540 5/60 510 8/60

1,040 8/59 1,020 8/60

2,070 29/59* 2,110 6/60

NTP (1997)

F344/N rat; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1

, 643, 3,273 or 6,368 mg/m 3 ) (dynamic whole body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical concentration and method were reported H istopathology data for the 13

- week study were not provided , but the liver was evaluated in control and high

- dose group indicating that no changes in liver histopathology were observed in the 13

- week study.

NTP (1997)

B6C3F 1 mouse; 10/sex/treatment Analytical concentration: 0, 134, 272, 542, 1,080, or 2,101 ppm (0, 406, 824, 1,643, 3,273 or 6,368 mg/m 3 ) (dynamic whole body chamber) 6 hr/d, 5 d/wk 13 weeks Generation method (Sonimist Ultrasonic spray nozzle nebulizer), analytical concentration and method were reported Authors stated that there were no treatment

- related microscopic observations, but data were not provided.



### 1 2 3 4 5 6 7

8

9 10

11

12 a The high

- dose group had an increase in mortality. * Statistically significant p ≤ 0.05 as determined by study authors. Conversions from drinking water concentrations to mg/kg

- d per formed by study authors. Conversion from ppm to mg/m 3 is 1 ppm = 3.031 mg/m 3

.

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**T able**  **1- 14** 

**. Changes in u**  **rinary bladder histopathology**  **in animals following**  

### 1 2 **oral exposure to**  tert

**- butanol** 

Reference and s tudy design Results
NTP (1995)

F344/N rat; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, 40 mg/mL) M: 0, 230, 490, 840, 1,520, 3,610 a mg/kg -d F: 0, 290, 590, 850, 1,560, 3,620 a mg/kg -d 13 weeks Incidence (severity) :

Males Females

Dose (mg/kg

- d) T ransitional epithelial hyperplasia Dose (mg/kg

- d) Transitional epithelial hyperplasia

0 0/10 0 0/10

230 n ot evaluated 290 not evaluated

490 n ot evaluated 590 not evaluated

840 0/10 850 not evaluated

1,520 1/10 (3.0) 1,560 0/10

3,610 7/10* (2.9) 3,620 3/10 (2.0)
Severity: 1 = minimal, 2 = mild, 3 = moderate, 4 = marked
NTP (1995)

B6C3F 1 mouse; 10/sex/treatment Drinking water ( 0, 2.5, 5, 10, 20, 40 mg/mL) M: 0, 350, 640, 1,590, 3,940, 8,210 a mg/kg -d F: 0, 500, 820, 1,660, 6,430, 11,620 a mg/kg -d 13 weeks Incidence (severity):

Males Females

Dose (mg/kg

- d) Transitional epithelial hyperplasia Inflam

mation Dose (mg/kg

- d) Transitional epithelial hyperplasia Inflam

mation

0 0/10 0/10 0 0/10 0/10

350 not evaluated 500 0/1 0/1

640 not evaluated 820 not evaluated

1,590 0/10 0/10 1,660 not evaluated

3,940 6/10* (1.3) 6/10* (1.3) 6,430 0/10 0/10

8,210 10/10* (2.0) 10/10* (2.3) 11,620 3/9 (2.0) 6/9* (1.2)
Severity: 1 = minimal, 2 = mild, 3 = moderate, 4 = marked
NTP (1995)

F344/N rat; 60/sex/treatment (10/sex/treatment evaluated at 15 months) Drinking water ( 0, 1.25, 2.5, 5, or 10 mg/mL) M: 0, 90, 200, 420 a mg/kg -d F: 0, 180, 330, 650 a mg/kg -d 2 years No treatment

- related effects observed



### 3

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**Table 1** 

**- 1**  **4**

**. Evidence pertaining to u**  **rinary bladder effects in**  **animals following oral**  **exposure to**  tert

- **butanol (**  continued **)**

Reference and s tudy design Results
NTP (1995)

B6C3F 1 mouse; 60/sex/treatment Drinking water ( 0, 5, 10, or 20 mg/mL) M: 0, 540, 1,040, 2,070 a mg/kg

- d F: 0, 510, 1,020, 2,110 mg/kg -d 2 years Incidence (severity) :

Males Females

Dose (mg/kg

- d) Transitional epithelial hyperplasia Inflam

mation Dose (mg/kg

- d) Transitional epithelial hyperplasia Inflam

mation

0 1/59 (2.0) 0/59 0 0/59 0/59

540 3/59 (1.7) 3/59 (1.7) 510 0/60 0/60

1,040 1/58 (1.0) 1/58 (1.0) 1,020 0/59 0/59

2,070 17/59* (1.8) 37/59* (2.0) 2,110 3/57 (1.0) 4/57* (2.0)

S everity: 1 = minimal, 2 = mild, 3 = moderate, 4 = marked 

### 1 2 3 a The high

- dose group had an increase in mortality. * Statistically significant p ≤ 0.05 as determined by study authors. Conversions from drinking water c oncentrations to mg/kg

- d performed by study authors.

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### 1

2 3

4 5  = exposures at which the endpoint was reported statistically significant by study authors  = exposures at which the endpoint was reported not statistically significant by study authors

× = exposures at which all animals were dead and unable to be examined for the endpoint

M rat (D) M rat (A) F rat (D) M mouse (D) F mouse (D) M rat (C) F rat (C) Transitional epithelium hyperplasia; M rat (D) Transitional epithelium hyperplasia; F rat (D) Transitional epithelium hyperplasia; M mouse (D) Inflammation; M mouse (D) Inflammation; F mouse (D) Transitional epithelium hyperplasia; F mouse (D) Transitional epithelium hyperplasia; M mouse (D) Inflammation; M mouse (D) Inflammation; F mouse (D) Transitional epithelium hyperplasia; F mouse (D) Increased glycogen; M rat (B) Relative weight; M rat (B) Relative weight; M rat (D) Absolute weight; F rat (D) Relative weight; F rat (D) Absolute weight; M rat (D) Relative weight; M mouse (D) Relative weight; F mouse (D) Absolute weight; M mouse (D) Absolute weight; F mouse (D) Relative weight; M rat (D) Relative weight; F rat (D) Absolute weight; M rat (D) Absolute weight; F rat (D) Fatty tissue; M mouse (D) Fatty tissue; F mouse (D) Relative weight; M rat (C) Absolute weight; M rat (C) Absolute weight; F rat (C) Relative weight; F rat (C)BODY WEIGHT

SubchronicURINARY BLADDERLIVER Subchronic

Chronic

Subchronic

Chronic Reproductive

Reproductive

10 100 1,000 10,000 100,000 

### Dose (mg/kg

- day)

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**Figure 1-8.**  **Exposure**

**- response array of other effects following oral exposure** 
**to** tert

**- butanol** 

**.**Toxicological Review of tert

- Butyl Alcohol



### 1

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### 2 3

4 5 Source : (A) NTP (1997)
**Figure 1-9**  **. E x**  **posure -**  **response array of other effects following inhal ation**  **exposure to** tert

**- butanol** 

**.**Toxicological Review of tert

- Butyl Alcohol



## 1.2. **INTEGRATION AND EVALUATION**  1

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37 **1.2.1. Effects Other Than Cancer**  The strongest evidence following tert

- b utanol exposure is for kidney, with t oxicity observed after oral exposure in two strains of rats and in one str a in of mice and in both sexe s

. In mice
, the only kidney effect observed was an increase in kidney weight (absolute and/or relative) in both

sexes o f mice in the 13

- week study , but no treatment

- related histopathological lesions were reported i n the kidneys of mice at the 13

- week or 2

- year time point s ( NTP, 1995 ) . In male rats,
effects related to the ac c umulation of α 2u - globulin in the kidney have been reported, including precursors to granular casts, linear mineralization, and tubular hyperplasia, but these are not
considered relevant to humans ( Hard et al., 2011 ; Cirve llo et al., 1995 ; NTP, 1995 ; Lindamood et al.,
1992 )

. However, several other effects in the kidney unrelated to α 2u - globulin were observed in female and/or male rats. Absolute and relative kidney weights were increased in both male and
female rats after b oth 13 week s and 15 m onths of treatment ( NTP, 1995 )

. H istopathological examination also indicated kidney toxicity in b oth male and female rat s, with increased incidence of nephropathy after 13 weeks of oral exposure and transitional epithelium hyperplasia observed after 2 years of oral exposure ( NTP, 1995 ) . Additionally, increased inflammation (supp u rative) was noted in females after 2 years oral exposure ( NTP, 1995 ) . EPA identified kidney effects as a human
hazard of t ert

- butanol oral exposure

. Fewer and less severe kidney effects were observed via inhalation than via oral exposure, likely due to the differing levels of internal doses achieved via the different routes. Specifically,
available inhalation studies ( NTP, 1997 ) were conducted at concentrations that are comparable, in
terms of tert

- b utanol blood concentration, to the lower range of doses in oral studies. Moreover, there is convincing toxicokinetic data to indicate th at tert

- butanol is absorbed by both routes, and kidney effects are remote from the site of absorption. EPA identified kidney e ffects as a human
hazard of t ert

- butanol inhalation exposure

. Thyroid follicular cell h yperplasia was observed in the mice after 2 years of exposure via drinking water ( NTP, 1995 ) ; and EPA identified thyroid effects as a potential human hazard of tert

butanol exposure. However, this endpoint most likely reflects early events in the neoplastic

progre s sion of thyroid follicular cell tumors following tert

- butanol exposure (see Section 1.1.2 ) and
was not considered further for dose

- r esponse analysis and derivation of noncancer reference
values. EPA identified suggestive evidence of developmental effects as a potential human hazard of

tert

- b utanol exposure. Exposure to high doses of tert

- butanol during gestation resulted in some effects i n exposed offspring or pups , although the e ffects were not always consistent across exposure routes (oral and inhalation). Dams exhibit effects at the same doses as fet al effects. Neurodevelopmental effects have also been observed; however, the neurodevelopmental studies
h a d limitations in the stu dy design and/or reporting and results were in consistent between studies

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o r across dose. Thus, these effects were not considered further for dose

- response analysis and 

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38 derivation of reference values. EPA concluded that the evidence does not support repr oductive effects, body weight changes, liver effects, and urinary bladder effects as potential human hazards of tert

- b utanol

exposure. Thus, these effects were not considered further for dose

- r esponse analysis and the
derivation of reference values.

**1.2.2. C**  **arci nogenicity** 

Under E PA’s Guidelines for Carcinogen Risk Assessment ( U.S. EPA, 20 05a ) , the database for

tert

- b utanol pr ovides “s uggestive evidence of carcinogenic potential ,” based on a statistically significant i n crease in renal tumors (renal tubule adenomas and carcinomas ) in male F344 rats and

a st atistically significant increase in thyroid follicular cell adenomas in female B6C3F 1 mice , all exposed to tert

- b utanol in drinking water for 2 years ( Cirvello et al., 1995 ; NTP, 1995 )

. There are
no available studies o f ca ncer in humans associated with exposure to tert

- butanol

. In the NTP (1995) rodent bioassay, tert

- butanol

- exposed male rats had a significant increase in renal tumors compared to controls, a result confirmed by a PWG reevaluation ( Hard et al., 2011 )

. Although mechanistic data show that α 2u - globulin

- related processes occur with tert

butanol ex posure, there is insufficient evidence to support a conclusion that α 2u - globulin nephropathy is the sole or primary contributor to renal tumor development. S pecifically,

tert

- b utanol induced tumors at lower doses than those for other precursor effects such as hyperplasia a nd granular casts , with no further increase in tumor i ncidence coinciding with the
induction of additional markers of α 2u - globulin nephropathy. B ased on analysis of available mode of action data, these tumors are not attributed t o α 2u - globulin and are considered relevant in
humans ( U.S. EPA, 1991a )

. tert

- Butanol was negative in a variety of genotoxicity assays in different cell systems including gene m utations, sister chromatid exchanges, micronucleus formation

, and
chromosomal aberrations

. H owever, DNA adducts in male Kunming mice and DNA damage in
human HL

- 6 0 leukemia cells have been observed. Overall, the mode(s) of carcinogenic action for

tert

- b utanol in the kidney and the thyroid are not known , and these tumor data are considered
relevant to humans.
As e mphasized in the Cancer Guidelines ( U.S. EPA, 2005a ) , selection of the cancer descriptor
fol l ows a full evaluation of the available evidence. The carcinogenicity evidence for tert

- butanol could be considered a borderline case between two cancer descriptors —“ s uggestive evidence of carcinogenic potential ” and “ likely to be carcinogenic to humans .” The descriptor of “ suggestive evidence of carcinogenic potential ” is appropriate when a concern for potential carcinogenic effects in humans is raised, b ut the data are judged not sufficient for a stro nger conclusion . Exposure to

tert

- b utanol produced a positive tumor response at more than one site ( kidney and thyroid ) and in more than one species (rat and mouse)

. T hese data appear to correspond closely to one of the examples in the C ancer Guidelines ( U.S. EPA, 2005a ) for the descriptor of “ likely to be carcinogenic
to humans ; ” i.e., “an agent that has tested positive in animal experiments in more than one species,

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s ex, strain, site, or exposure route, with or without evidence of carcinogenicity in humans.” S everal 

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38 aspects of the data support the conclusion that these data ar e not sufficient to characterize tert

butanol a s “ likely to be carcinogenic to humans .”

First
, the renal tumors associated with tert

- butanol exposure in the NTP (1995) rodent bioassay w ere predominantly benign. Based on the PWG reevaluation ( Hard et al., 2011 ) , among the t hree treated groups, only two of the 43 animals with tumors has carcinomas (there were no carcinomas among the 4 control animals with tumors). Additionally, n o kidney tumors were
observed in female rats or in either sex of mice. Furthermore, ETBE

, whic h is rapidly metabolized

to t ert

- butanol, did not induce kidney tumors in the same strain of rats at doses that resulted in
similar internal concentrations of tert

- bu tanol

. Therefore, the level of concern raised by renal
tumors associated with tert

- b utanol exposure is reduced based on the predominance of benign

tumors , a n increase in renal tumors in a single sex/species combination only

, and the lack of
coherence with the metabolically- r elated compound ETBE. The thyroid tumors associated with t ert -b utanol exposure were also predominantly benign.

I n t he NTP (1995) rodent bioassay, o nly female mice had a statistically significant increase in thyroid tumors ; n one of these were carcinomas. In males, decreased survival complicates the interpretation of thyroid tumors because male mice had an in creased incidence of thyroid follicular
cell h yperplasia at all exposure levels , but there was no significant increase in thyroid tumors at any

exposure

. I nterestingly

, one thyroid follicular cell carcinoma occurred in a high

- dose male , but
limited conclusions can be drawn fro m this single observation

. Thyroid tumors were not observed

in ei ther sex of the rat exposed chronically to tert

- butanol

. Addition ally , ETBE did not induce thyroid tumors

, al t hough only rats and not mice were tested

. Therefore, the level of concern rai sed by thyroid tumors associated with tert

- b utanol exposure is reduced based on the predominance of benign tumors and an increase in thyroid tumors in a single sex/species combination only

. Overall

, the cancer descriptor “ suggestive evidence of carcinogeni c potential ” was selected

,

as s ome concern is raised by the positive evidence of predominantly benign renal tumors in male rats and thyroid tumors in female mice.
The Cancer Guidelines ( U.S. EPA, 2005a ) indicate that for tumors occurring at a site other than the initial point of contact, the weight of evidence for carcinogenic po t ential may apply to all routes of exposure that have not been adequately tested at sufficient doses

. A n exception occurs when there is convincing toxicokinetic data that absorption does not occur by other routes

. Information available on the carcinogeni c effects of tert

- butanol via the oral route demonstrates that tumors occur in tissues remote from the site of absorption

. In formation on the carcinogenic
effects of tert

- b utanol via the inhalation and dermal routes in humans or animals is not available

. Based on the observation o f systemic tumors following oral ingestion, and in the absence of information to indicate otherwise, it is assumed that an internal dose will be achieved regardless of the route of exposure

. T herefore, there is “ suggestive evidence of carcinogenic potential ” f rom exposure to tert

- bu tanol by all routes of exposure.

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**1. 2.3.**  **Susceptible Populations and Lifestages for Cancer and Noncancer Outcomes**  

### 1

2 No data were identified to indicate any possible susceptible populations or lifestages.

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### 1 

# 2. DOSE

- RESPONSE ANALYSIS

**2.1. ORAL REFERENCE DOSE FOR EFFECTS OTHER THAN CANCER**  2

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30
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32 The 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

. I t can be derived from a no

- observed

- adverse

- effect level (NOAEL), lowest

- observed

adverse- e ffect level (LOAEL), or the 95% lower bound on the benchmark dose (BMDL), with uncertainty factors (UFs) generally applied to reflect limitations o f the data used.

**2.1.1. I**  **dentification of Studies and Effects for Dose**

**- Response Analysis** 

EPA identified kidney eff ects as a human hazard of tert

- butanol exposure. Studies within
this effe ct category were evaluated using general study quality characteristics (a s discussed in Section 6 of the Preamble) to help inform the selection of studies from which to derive toxicity values. Rationales for selecting the studies and effects to represent each of these hazards are
summarized below. Human studies are preferred o ver animal studies when quantitative measures of exposure are reported and the reported effects are determined to be associated with exposure. However,

t h ere are no available human occupational or epidemiological studies of oral exposure to tert

butanol

. Animal studies were evaluated to determine which studies p rovided : (a) the most relevant routes and durations of exposure; (b) m ultiple exposure levels to provide information about the shape of the dose

- r esponse curve; and (c) power to detect effects a t low exposure levels ( U.S. EPA,
2002 ) . Sufficient data were available to develop a PBPK model in rats for both oral and inhalation exposure in order to perform route

- to

- r oute extrapolation, so rat studies from both routes of exposure were considered for dose

- r esponse analysis. The database for tert

- butanol includes
several st udies and data sets that are potentially suitable for use in deriving reference values. Specifically, effects associated with t ert

- butanol exposure in animals include observations o f organ weight and /o r histological changes in the kidney observed in several chronic and subchronic
studies

.
Kidney To xicity

EPA identified kidney eff ects as a human hazard of tert

- butanol

- induced toxicity based on findings of organ weight c hanges in rat s and mice

, as well as histopathology in rats

. These findings were consistent across multiple chronic, subchronic, and short

- t erm studies following oral and

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inhalation exposure

. Acharya et al. ( 1997 ; 1995 ) used a single exposure group and did not provide 

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38 incidence or severity data, and thus was not considered for dose

- response assessment

. Lyondell Chemical Co. (2004) and NTP (1997) were of subchronic or shorter duration, and so were set aside given the availability of a longer duration study. Therefore, t h e NTP 2

- year drinking water study ( NTP, 1995 ) was identified most suitable for dose

- response assessment considering the study duration, co mprehensive reportin g of outcomes, multiple species tested, and multiple doses tested.
In the NTP (1995) drinking water study, male F344 rats were exposed to approximate doses of 0, 90, 200, or 420 mg/kg

- d ay; female F344 rats were exposed to approximate doses of 0, 180, 330, or 650 mg/kg

- d ay; male B6C3F 1 mice were exposed to approximate doses of 0, 540, 1,04 0 , or 2,070 mg/kg

- day ; and female B6C3F 1 mice were exposed to approximate doses of 0, 510, 1,020, or 2,110 mg/kg

- d ay. Reduced body weights and survival were observed and reflected in
some of the effects

. K idney effects including changes in organ weight and /or histopathology were observed in both sexes in rats and mice. Effects were also observed after 13 weeks, 15 months, and 2 years of treatment ( NTP, 1995 ) . Effects were more consistent and occurred at lower doses in rats as compared to mice, so a s a result, only data in the more sensitive species of rats were used for
dose- r esponse assessment. Endpo ints potentially confounded by the presence of α 2u - globulin nephropathy in male rats, such as linear mineralization and renal tubule hyperplasia, were not used
for dose

- r esponse analysis. Specific endpoints chosen for analysis were absolute and relative kidney weight (observed in males and females), kidney inflammation (observed only in females), and kidney transitional epithelial hyperplasia (observed in males and females). For most endpoints, the data at the longest duration of 2 y ears were selected. However, as discussed in Section 1.1.1, 2

- y ear kidney weight data were not considered because organs were only weighed at
15 months

.

**2.1.2. M**  **ethods of Analysis** No biologically based dose

- r esponse models are available for tert

- butanol . In this situation, EPA evaluates a range of dose

- r esponse models thought to be consistent with underlying biological processes to determine how to best e mpirically model the dose

- response relationship in the range of the observed data. Consistent with this approach, all models available in E PA’s Benchmark Dose Software (BMDS) were evaluated. Consistent with EPA’s Benchmark Dose Technical Guidance

Document ( U.S. EPA, 2012b ) , the benchmark dose (BMD) and the 95% lower confidence limit on the BMD (BMDL) were estimated using a benchmark response (BMR) of 10% change from the control
mean for organ weight data in the absence of information regarding th e level of change that is considered biologically significant . F urthermore, the BMD and BMDL were estimated to facilitate a consistent basis of comparison across endpoints, studies, and assessment s

. A benchmark response (BMR) of 10% extra risk was considered appropriate for the q uantal data on incidence s of kidney
inflammation and kidney transitional epithelial hyperplasia. For each endpoint, the BMDL estimate (95% lower confidence limit on the BMD, as estimated by the profile likelihood method) and Akaike Information Criterion ( AI C ) value were used to select a best

- fit model among models exhibiting

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ad equate fit. If the BMDL estimates were “sufficiently close,” that is, differed by at most 3

- fold, the 

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the best

- fit model for each endpoint can be found in Appendix C of the Supplemental Information. In general, absolute and relative kidney weight data may both be considered appropriate endpoints for analysis ( Bailey et al., 2004 ). However, i n the NTP (1995) 2

- year drinking water study, t here was a noticeable decrease in body weight in exposed animals relative to controls at the 15 month interim sacrifice (see T able 1-1 ). In such a case, relative kidney weights are the preferred, so changes in absolute kidney weights were not analyzed

. Human equivalent doses (HEDs) for oral exposures were derived from th e PODs estimated from the laboratory animal data as described in EPA’s Recommended Use of Body Weight 3/4 as the Default Method in Derivation of the Oral Reference Dose ( U.S. EPA, 2011 ) . In this guidance, EPA
advo c ates a hierarchy of approaches for deriving HEDs from data in laboratory animals, with the preferred approach being physiologically

- b ased toxicokinetic modeling. Other approaches can include using chemical

- sp ecific information in the absence of a complete physiologically

- based
toxicokinetic model. As discussed in Appendix B of the Supplemental Information, several rat physiologically based pharmacokinetic (PBPK) models for t ert

- butanol have been developed and published, but a validated human PBPK model f or te rt

- butanol for extrapolating doses from animals
to humans is not available. In lieu of either chemical

- sp ecific models or data to inform the derivation of human equivalent oral exposures, a body weight scaling to the 3⁄4 power (i.e., BW 3/4 ) approach is ap plied to extrapolate toxicologically equivalent doses of orally administered agents from adult laboratory animals to adult humans for the purpose of deriving an oral RfD. BW 3/4 scaling was not employed for deriving HEDs from studies in which doses were a dministered directly to early postnatal animals because of the absence of information on whether allometric (i.e., body weight) scaling holds when extrapolating doses from neonatal animals to adult humans due to presumed toxicokinetic and/or toxicodynami c d ifferences between lifestages ( U.S. EPA,
2011 ; Hattis et al., 2004 ). Consistent with EPA guidance ( U.S. EPA, 2011 ) , the PODs estimated based on effects in adult
animals are c onverted to HEDs employing a standard dosimetric adjustment factor (DAF) derived
as follows:

D AF = (BW a 1/4 / BW h 1/4 ),

w here

B W a = animal body weight

B W h = human body weight

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Using a standard BW a of 0.25 kg for rats and a BW h of 70 kg for humans ( U.S. EPA, 1988 )

, 

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10 the resulting DAFs for rats is 0.24. Applying this DAF to the POD identified for effects in adult rats yields a POD HED as follows (see Table 2

- 1):

PO D HED = Laboratory animal dose (mg/kg

- day) × DAF

Table 2

- 1 summarizes the sequence of calculations leading to the derivation of a human

equivalent POD for each data set discussed above.

**Table 2**  **-1**

**. Summary of derivations of points of departure** 

Endpoint and

Reference Species/

sex Model a BMR BMD (mg/kg

- d) BMDL (mg/kg

- d ) POD ADJ b (mg/kg

- d) POD HED c (mg/kg

- d)
Kidney

Increased relative kidney weight NTP (1995) Rat/M Exponential (M4) 10% 117 48 48 11.5

Increased relative kidney weight NTP (1995) Rat/F Linear 10% 158 133 133 31.9

Kidney inflammation NTP (1995) Rat/F Log

- probit 10% 254 200 200 48

Kidney transitional epithelial hyperplasia NTP (1995) Rat/M Log

- logistic 10% 30 16 16 3.84

Kidney transitional epithelial hyperplasia NTP (1995) Rat/F Multistage, 3- degree 10% 412 339 339 81.4

a For modeling details, see Appendix C in Supplemental Information

. 

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c HED PODs were calculated using BW 3/4 scaling ( U.S. EPA, 2011 ).

**2.1.3. De**  **rivation of Candidate Values** 

Under EPA’s A Rev i ew of the Reference Dose and Reference Concentration Processes ( U.S. EPA, 2002; Section 4.4.5 ) , also described in the Preamble, five possible areas of uncertainty and variability were considered

. A n explanation follows:

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A n intraspecies uncertainty factor, UF H , of 10 was applied to all PODs to account for 

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38 potential differences in toxicokinetic s and toxicodynamic s in the absence of information on the variability of response in the human population follo wing oral exposure to tert

- butanol

. An interspecies uncertainty factor, UF A , of 3 ( 10 1/2 = 3.16, rou nded to 3) was a pplied to all
PODs b ecause BW 3/4 scaling is used to extrapolate oral doses from laboratory animals to humans. Although BW 3/4 scaling addresses some aspects of cross

- species extrapolation of toxicokinetic and toxicodynamic p rocesses, some residual uncertainty remains. In the absence of chemical

- specific data to quantify this uncertainty, EPA’s BW 3/4 guidance ( U.S. EPA, 2011 ) recommends use of an uncertainty factor of 3

. A su bchronic to chronic uncertainty factor, UF S , of 1 was applied to all PODs since the endpoints examined were all observed f ollowing chronic exposure

. A LO AEL to NOAEL uncertainty factor, UF L , of 1 was applied to all PODs because the current approach is to address this factor as one of the considerations in selecting a BMR for benchmark dose modeling. In this case, BMR s of a 10% change in relative kidney weight, a 10% extra risk of kidney inflammation, and a 10% extra risk of t ransitional cell hyperplasi a were selected under an assumption that they represent min imal biologically significant change s. A d atabase uncertainty factor, UF D , of 1 was applied to all PODs . T he tert

- butanol toxicity
database in cludes a chronic toxicity study in rats and mice ( NTP, 1995 ) , a subchronic toxicity study
in rats and mice ( NTP, 1997 )

, and developmental toxicity studies in rats and mice ( Lyondell Chemical Co., 2004 ; Faulkner et al., 1989 ; Daniel and Evans, 1982 ) . In the developmental studies, no effects were observed at exposure levels below 1000 m g/kg

- day, and effects observed at
≥1000 m g/kg

- day were accompanied by evidence of maternal toxicity. These exposure levels are much higher than the PODs for kidney effects, suggesting developmental toxicity is not a sensitive endpoint. The tert

- butanol database contains a one

- generation reproductive toxicity study in rats ( Lyondell Chemical Co., 2004 ) , though n o multigenerational reproductive study has been performed

. T h ere are no immunotoxicity studies for tert

- butanol. Information provided by studies

on ET BE, which is rapidly metabolized to systemically

- available tert

- butanol , can help in considering the lack of a tert

- b utanol multigenerational reproductive study or an immunotoxicity study

. N o adverse effects were reported in one - and two

- generation reproductive/developmental
studies on ETBE ( Gaoua, 2004a

, b ) , and the database fo r ETBE does not indicate immunotoxicity ( Banton et al., 2011 ; Li et al., 2011 ). Thus, although there are some gaps in the toxicity database for

tert

- b utanol

, the available data on tert

- butanol, informed by the data on ETBE, do not suggest that additional studies would lead to identification of a more sensitive endpoint or a lower POD. Therefore, a database UF D of 1 was applied.
Table 2

- 2 is a continuation of Table 2 -1 and summarizes the application of UFs to each POD
to derive a candidate value for each data set

. T he candidate values presented in the table below are preliminary to the derivation of the organ/system

- s pecific reference values

. These candidate values are considered individually in the selection of a representative oral reference value for a specific

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h azard and subsequent overall RfD for tert

- butanol

. 

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5 4 Figure 2 -1 presents graphically the candidate values, UFs, and PODs, with each bar corresponding t o one data set described in Table 2 -1 and Table 2

- 2

.

**Table 2-**  **2 . E**  **ffects and corresponding derivation of candidate**  **RfDs**

Endpoint and Reference POD HED (mg/kg

- d) POD type UF A UF H UF L UF S UF D Composite UF Candidate value (mg/kg

- d)

Kidney

Increased relative kidney weight; male rat NTP (1995) 12 BMDL 10% 3 10 1 1 1 30 4 × 10 - 1

Increased relative kidney weight; female rat NTP (1995) 32 BMDL 10% 3 10 1 1 1 30 1 × 10 0

Kidney inflammation; female rat NTP (1995) 48 BMDL 10% 3 10 1 1 1 30 2 × 10 0

Kidney transitional epithelial hyperplasia; male rat NTP (1995) 3.8 BMDL 10% 3 10 1 1 1 30 1 × 10 - 1

Kidney transitional epithelial hyperplasia; female rat NTP (1995) 81 BMDL 10% 3 10 1 1 1 30 3 × 10 0



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### 1

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**. Candidate**  **RfD values with corresponding POD and composite UF.** Toxicological Review of tert

- Butyl Alcohol

**2. 1.4.**  **Derivation of Organ/System -S**  **pecific Reference Doses** 

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- 3 distills the candidate values from Table 2

- 2 into a single value for the kidney

. O r gan or system

- speci fic reference values may be useful for subsequent cumulative risk assessments that consider the combined effect of multiple agents acting at a common site.

Kidney Toxicity

For t ert

- butanol , candidate values were for several different effects in both sexes, spanning a range from 1 × 10 -1 to 3 × 10 0 mg/kg -day , for an overall t hirty fold range . To estimate an exposure
level below which kidney toxicity from t ert

- butanol exposure is not expected to occur, the RfD for
increased incidence of transitional epithelial hyperplasia in male rats ( **1 × 10**  **-1 mg/kg** 

**- day**  ) is proposed as the kidney

- sp ecific reference dose for tert

- butanol

. Unlike kidney inflammation, this effect was observed in both sexes, with males appearing to be more sensitive than females.
Addit i onally, it is a more specific and more sensitive indicator of kidney toxicity than the relatively

non

- sp ecific endpoint of kidney weight changes. Confidence in this kidney

- specific RfD is high . The PODs are based on modeled benchmark dose estimates, a nd the candidate values are derived from

a w ell

- conducted study, involving a sufficient number of animals per group , including both sexes, and assessing a w ide range of kidney endpoints

.

**Table 2-**  **3**

**. Organ/system** 

**- specific RfDs and proposed overall RfD**  **for** tert

  ** butanol**

Effect Basis RfD (mg/kg

- day) Exposure description Confidence

Kidney toxicity Increased incidence of transitional epithelial hyperplasia 1 × 10 - 1 Chronic HIGH

Proposed overall RfD Increased incidence of transitional epithelial hyperplasia 1 × 10 -1 Chronic HIGH

**2.1.5. Selection of the Proposed Overall Reference Dose** 

For t ert

- butanol

, only kidney effects were identified as a hazard; thus a single

organ / system

- specific reference dose was derived. Therefore, the kidney

- specific RfD of

**1 × 10**  **-1 mg/kg** 

**- day**  is also proposed as an estimated exposure level below which deleterious
effects from t ert

- butanol exposure are not expected to occur. The overall reference dose is derived to be protective of all types of effects for a given duration of exposure and is intended to protect the population as a whole including potentially susceptible subgroups ( U.S. EPA, 2002 ) . Decisions concerning averaging exposures over time for comparison with the RfD shoul d consider the types of toxicological effects and specific lifestages of concern. Fluctuations in exposure levels that result in elevated exposures during these lifestages

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c ould potentially lead to an appreciable risk, even if average levels over the ful l exposure duration 

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34 were less than or equal to the RfD. In the case of tert

- butanol , no specific lifestages have been identified as a potentially susceptible subgroup.

**2.1.6. Confidence Statement**  A confidence level of high, medium, or low is assigned to the study used to derive the Rf D
, the overall database, and the Rf D it self, 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 high

. Confidence in the principal study ( NTP, 1995 ) is high

. Th is study was well -conducted, complied with FDA GLP regulations, involved a sufficient number of animals per group (including both sexes), and assessed a wide range of tissues and endpoints

. A lt hough there are some gaps in the toxicity databas e for tert

- butanol

, the se areas are informed by the data on ETBE, a parent compound of tert

- b utanol . Therefore, the confidence in the database is high. Reflecting high confidence in the principal study a nd high confidence in the database, confidence in the RfD is high

.

**2.1.7. P**  **revious IRIS Assessment**

An oral assessment for t ert

- butanol was not previously available on IRIS.



## 2.2. **INHALATION REFERENCE CONCENTRATION FOR EFFECTS OTHER THAN CANCER** 

The inhalation reference concen tration ( RfC ) (expressed in units of mg/m 3 ) is defined as an estimate (with uncertainty spanning perhaps an order of magnitude) of a continuous inhalation exposure to the human population (including sensitive subgroups) that is likely to be without an

appr e ciable 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.

**2.2.1. I**  **dentification of Studies and Ef fects for Dose** 

**- Response Analysis** 

EPA identified kidney eff ects as a human hazard of tert

- butanol exposure. Studies within
this effe ct category were evaluated using general study quality characteristics (as discussed in Section 6 of the Preamble) to help i n form the selection of studies from which to derive toxicity values. Rationales for selecting the studies and effects to represent th i s hazard are summarized
below. Human studies are preferred over animal studies when quantitative measures of exposure

a r e reported and the reported effects are determined to be associated with exposure. However,

t h ere are no available human occupational or epidemiological studies of inhalation exposure to

tert

- b utanol

.

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A nimal studies were evaluated to determine which st udy provided : (a) the most relevant 

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35 routes and durations of exposure; (b) multiple exposure levels to provide information about the shape of the dose

- response curve; and (c) power to detect effects at low exposure levels ( U.S. EPA,
2002 ) . Sufficient data were available to develop a PBPK model in rats for both oral and inhalation exposure in order to perform route

- to

- r oute extrapolation, so rat studies from both routes of exposure were considered for dose

- r esponse analysis. The database for tert

- butanol includes a
several st udies and data sets that are potentially suitable for use in deriving reference values. Specifically, effects associated with t ert

- butanol exposure in animals include observations of organ weight and histological changes i n the kidney in several chronic and subchronic studies

.
Kidney To xicity

EPA identified kidney eff ects as a human hazard of tert

- butanol exposure based on findings of organ weight ch anges in rats and mice and histopathology in rats. These findings were consistent across multiple chronic, subchronic, and short

- t erm studies following oral and inhalation exposure. Acharya et al. ( 1997 ; 1995 ) used a single exposure group and did not provide incidence or severity data, so was not considered for dose

- r esponse assessment

. Lyondell Chemical Co. (2004) was of shorter than subc hronic duration, and so was set aside given the availability of a longer duration studies. Given the availability of a chronic study, the subchronic studies of NTP (1995) and NTP (1997) would normally also be set aside for dose

- response analysis. NTP (1997) is the longest duration study via the inhalation route, not requiring route

- to

- r oute extrapolation, so was kept for comparis on purposes. Overall, t he NTP 2

- year drinking water study NTP (1995) was identified a s the study most suitable for dose

- response assessment , given the study duration, comprehensive reporting of outcomes, use of multiple species tested, multiple doses tested , an d availability of a PBPK model for route

- to

- r oute extrapolation

. This study was discussed previously in Section 2.1.1 as part of the derivation of the oral reference dose, so will not be reviewed here again. The NTP (1997) subchronic inhalation study is described in more detail below. NTP (1997) was a well -designed subchronic study that evaluated the effect of tert

- butanol exposure on multiple species at multiple inhalation d oses . Briefly, groups of F344 rats and B6C3F 1 mice (10 pe r sex per species) were exposed to tert

- butanol (>99% pure) at concentrations of 0, 409, 819, 1
, 6 37, 3

, 274 or 6

, 366 mg/m 3 by inhala tion for 6 hours per day, 5 days per week, for 13
weeks ( NTP, 1997 ). Absolute kidney weights were elevated (10

– 11%) in male rats exposed at

≥ 3,274 m g/m 3 ; relative kidney weights were statistically significantly elevated (~9%) in males at

≥ 3,274 m g/m 3 and females at 6,366 mg/m 3 . Male rats exhibited an increase in the severity of chronic nephropathy (characterized as number of foci of regenerative tubules) . T here were few endpoints available for consideration in the subchronic study, but changes in kidney weights were
also observed in the o r al studies, such as the NTP (1995) 2

- year drinking water study

.

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**2. 2.2.**  **Methods of Analysis** 

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37 No biologically based dose

- response models are available for tert

- butanol . In this situation, EPA evaluates a range of dose

- r esponse models thought to be consistent with underlying biological processes to determine how best to e mpirically model the dose -r esponse relationship in the range of the observed data. Consistent with this approach, all models available in E PA’s Benchmark Dose Software (BMDS) were evaluated. Consistent with EPA’s Benchmark Dose Technical Guidance

Document ( U.S. EPA, 2012b ) , the benchmark dose (BMD) and the 95% lower confidence limit on the BMD (BMDL) were estim ated using a benchmark response (BMR) of 10% change from the control
mean for organ weight data in the absence of information regarding what level of change is considered biologically significant, and also to facilitate a consistent basis of comparison ac r oss endpoints, studies, and assessments. A benchmark response (BMR) of 10% extra risk was considered appropriate f or the quantal data on incidence s of kidney inflammation and kidney transitional epithelial hyperplasia. T he estimated BMDLs were used as points of departure (PODs). Further details including the modeling output and graphical results for the best

- f it model for each endpoint can be found in A ppendix C of the Supplemental Information. In general, absolute and relative kidney weight data may b o th be considered appropriate endpoints for analysis ( Bailey et al., 2004 )

. However, i n the NTP (1995) 2

- year drinking water study, t here was a noticeable decrease in body weight in exposed animals relative to controls at the
15 m o nth interim sacrifice (see Table 1-1 ). In such a case, relative kidney weights are preferred, so changes in absolute kidney weights from NTP (1995) were not analyzed

. However, body weights were not impacted in the NTP (1997) subchronic inhalation study. Based on a historical review of
26 studies of control rats f rom 1

- month bioas says , Bailey et al. (2004) concluded that neither absolute kidney weight nor relative k idney:body (or kidney: brain) weight are optim al for evaluating organ weight changes. S ince neither approach is preferred, both were considered to be appropriate for BMD analysis of the NTP (1997) data set

.
PODs from I n halation S tudies Because the RfC is applicable to a continuous lifetime human exposure but derived from animal studies featuring intermittent exposure, EPA guidance ( U.S. EPA, 1994 ) provides mechanisms for: (1) adjusting experimental exposu re concentrations to a value reflecting continuous exposure duration (ADJ) and (2) determining a human equivalent concentration (HEC) from the animal exposure data. The former employs an inverse concentration

- t ime relationship to
derive a health

- p rotectiv e duration adjustment to time

- weight the intermittent exposures used in
the studies. The modeled benchmark concentration from the inhalation study ( NTP, 1997 ) w as adjusted to reflect a continuous exposure by multiplying it by (6 hours per day) ÷ (24 hours per day) and (5 days per week) ÷ (7 days per week) as follows:

BMCL ADJ = BMCL (mg/m 3 ) × (6 

### ÷ 24 ) × (5 

### ÷ 7)

= BMCL (mg/m 3 ) × (0.1786)

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Th e RfC methodology provides a mechanism for deriving a HEC from the duration -adjusted 

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20 POD (BMCL ADJ ) determined from the animal data. The approach takes into account the extra

respiratory nature of the toxicological responses and accommodates species differences by considering blood:air partition coefficients for tert

- butanol in the laboratory animal (r at or mouse) and humans. According to the RfC guidelines ( U.S. EPA, 1994 )

, tert

- butanol is a Category 3 gas
because extra

- r espiratory effects were observed. Kaneko et al. (2000) measured a blood:gas partition coefficient of 531 ± 102 for tert

- butanol in the male Wistar rat, while Borghoff et al. (1996) measured a value of 481 ± 29 in male F344 rats

. A blood:gas partition coefficient of 462 was reported for tert

- bu tanol in humans ( Nihlén et al., 1995 ) . The calculation (H b/g ) A ÷ (H b/g ) H was used to calculate a blood:gas partition coefficient ratio to apply to the delivered concentration. Because
F344 rats were used in t he study, the blood:gas partition coefficient for F344 rats was used. Thus, the calculation was: 481 ÷ 462 = 1.04. Therefore, a r atio of 1.0 4 was used to calculate the HEC

. This
allowed a BMCL HEC to be derived as follows:

BMCL HEC = BMCL ADJ (mg/m 3 ) × (interspecies conversion)

= B MCL ADJ (mg/m 3 ) × ( 481 

### ÷ 462 )

= BMCL ADJ (mg/m 3 ) × ( 1.04 )

Table 2 -4 summarizes the sequence of calculations leading to the derivation of a human

equivalent point of departure for each inhalation d ata set discussed above.

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**T able**  **2-4 . Summary of derivation of PODs**  **following inhalation exposure** 

### 1

Endpoint and Reference Species/
Sex Model a BMR BM C b (mg/ m 3 ) BM C L b (mg/ m 3 ) POD ADJ b (mg/ m 3 ) POD HE C c (mg/ m 3 )

Kidney

Increased relative kidney weight NTP (1997) Male F344

ra ts Linear 10% 6309 4821 861 8 61

Increased absolute kidney weight NTP (1997) Male F344
rats Hill 10% 1931 1705 304 3 04

Increased relative kidney weight NTP (1997) Female F344
rats No model selected 10%

-- -- -- --

Increased absolute kidney weight NTP (1997) Female F344
rats No model selected 10%

-- -- -- --

a For modeling details, see Appendix C in Supplemental Information. 

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25 b BMCs, BMCLs, and PODs were adjusted for continuous daily exposure by multiplying by (hours exposed per day / 24 hrs) × ( days exposed per week / 7 days).

c POD HEC calculated by adjusting the POD ADJ by the DAF (=1.0) for a category 3 gas ( U.S. EPA, 1994 ).
d BMD modeling failed to successfully calculate a BMD value (see Appendix C ).

PODs from oral studies – use of PBPK model for route

- to

- route extrapolation

A PBPK model for tert

- bu tanol in rats has been developed, as described in Appendix B

. Using this model, route

- to

- r oute extrapolation of the oral BMDLs to derive inhalation PODs was performed as follows. First, the internal dose in the rat at each oral BMDL (assuming continuous exposure) was estimated using the PBPK model, to derive an “internal dose BMDL.” Then, the inhalation air concentration (again

, a ssuming continuous exposure ) that led to the same internal dose in the rat was estimated using the PBPK model. The resulting BMCL was then converted to a human equivalent concentration POD using the methodology previously described in “PODs from
inhalation studies ” : BMCL HEC = BMCL ADJ (mg/m 3 ) × (interspecies conversion)

= B MCL ADJ (mg/m 3 ) × ( 481 ÷ 462 )

= BMCL ADJ (mg/m 3 ) × ( 1.04 )

A critical decision in the route

- to

- r oute extrapolation is the selection of the internal dose
metric that establishe s " equivalent” oral and inhalation expo sures. For tert

- butanol

- induced kidney effects, the two options are the concentration of t ert

- butanol in blood and rate of tert

- butanol metabolism. Note that using the kidney concentration of tert

- b utanol will lead to the same route

to

- r oute extrapolation relationship as tert

- butanol in blood, since the distribution from blood to kidney is independent of route. There are no data to s uggest that metabolites of tert

- butanol

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m ediate its renal toxicity. In the absence of evidence that would suggest otherwise, it is assumed 

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6 7 that tert

- butanol itself is the active toxicological agent. Therefore, the concentration of tert

- butanol
in blood was selected as the dose metric.
Table 2 -5 summarizes the sequence of calculations leading to the derivation of a human

equivalent point of departure for each oral d ata set discussed above.

**Table 2-**  **5 . Summary of derivation of**  **inhalation points of departure**  **derived from route** 

**- to** 

**- route extrapolation from oral exposures** 

Endpoint and reference Species/sex BMR BMDL (mg/kg

- d) Internal dose a (mg/L) Equivalent POD HE C b (mg/m 3 )

Kidney

Mean relative kidney weight NTP (1995) Rat/M 10% 48 2.34 79.6

Mean relative kidney weight NTP (1995) Rat/F 10% 133 7.46 231

Kidney inflammation NTP (1995) Rat/F 10% 200 12.6 359

Kidney transitional epithelial hyperplasia NTP (1995) Rat/M 10% 16 0.745 26.1

Kidney transitional epithelial hyperplasia NTP (1995) Rat/F 10% 339 27.9 638



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23 a Average blood concentration of tert

- butanol under continuous oral exposure at the BMDL.
b Continuous inhalation human equivalent concentration that leads to the same average blood concentration of

tert

- but anol as continuous oral exposure at the BMDL.

PODs carried forth to derivation of candidate v alues

For the derivation of candidate values , i t must be considered whether PODs from the inhalation study of NTP (1997) would provide a better basis than the route

- to

- route extrapolated PODs based on the oral study of NTP (1995) . The only endpoint available from NTP (1997) is increased kidney weights. The corresponding PODs from this subchronic inhalation study are
substantiall y higher than those for the same endpoint derived by route

- to

- route extrapolation from the chronic study ( NTP, 1995 ) , consistent with longer duration requiring a lower dose to elicit an effect. Additionally, as discussed in Section 2.1.3, kidney weight is a less

- s pecific endpoint compared to some of the other endpoints available for analysis from the o r al study ( NTP, 1995 )

. Therefore, the PODs derived from PBPK model

- b ased route

- to

- route extrapolat ion are the preferred basis for deriving kidney

- sp ecific candidate RfCs, as they are based on a longer (chronic)
duration and a m ore specific endpoint.

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**2. 2.3.**  **Derivation of Candidate Values** 

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38 Under EPA’s A Review of the Reference Dose and Reference Concentration Processes ( U.S. EPA, 2002; Section 4.4.5 ) , also described in the Preamble, five possible areas of uncertainty and variability were considered. An explanation follows: An intraspecies uncertainty factor, UF H , of 10 was applied to all PODs to account for potential differences in t oxicokinetic s and toxicodynamic s in the absence of information on the variability of response in the human population following inhalation exposure to tert

- bu tanol

. An interspecies u n certainty factor, UF A , of 3 (10 1/2 = 3.16, rou nded to 3) was a pplied to all PODs to account for residual uncertainty in the extrapolation from laboratory animals to humans in the absence of information to characterize toxicodynamic differences between rod e nts and humans after inhalation exposure to tert

- bu tanol . This value is adopted by convention where an adjustment from animal to a human equivalent concentration has been performed as described i n E PA’s Methods for Derivation of Inhalation reference Conc e ntrations and Application of Inhalation Dosimetry ( U.S. EPA, 1994 ).

A su bchronic to chronic uncertainty factor, UF S , of 1 was applied to the POD s derived from
the NTP (1995) study , as the endpoints were observed following chronic exposure

. For the P ODs
derived from t h e subchronic NTP (1997) study, a UF S of 10 was applied to account for extrapolation
from subchronic t o chronic duration.
A LO AEL to NOAEL uncertainty factor, UF L , of 1 was applied to all PODs because the current approach is to address this factor as one of t h e considerations in selecting a BMR for benchmark dose modeling. In this case, BMR s of a 10% change in kidney weight, a 10% extra risk of kidney inflammation, and a 10% extra risk of transitional cell hyperplasia were s elected under an assumption that th ey represent minimal biologically significant change s. A d atabase uncertainty factor, UF D , of 1 was applied to all PODs. T he tert

- butanol toxicity
database in cludes a chronic toxicity study in rats and mice ( NTP, 1995 ) , a subchronic toxicity study
in rats and mice ( NTP, 1997 ) , and developmental toxicity studies in rats and mice ( Lyondell Chemical Co., 2004 ; Faulkner et al., 1989 ; Daniel and Evans, 1982 ) . In the developmental studies, no effects were observed at exposure levels below 1000 m g/kg

- day, and effects observed at
≥1000 m g/kg

- day were accompanied by evidence of maternal toxicity. These exposure levels are much higher tha n the PODs for kidney effects, suggesting developmental toxicity is not a sensitive endpoint. The tert

- b utanol database contains a one

- generation reproductive toxicity study in rats ( Lyo ndell Chemical Co., 2004 ) , though no multigenerational reproductive study has been performed

. There are no immunotoxicity studies for tert

- butanol. Information provided by studies on ETBE, which is rapidly metabolized to systemically

- av ailable tert

- butanol , can help in considering the lack of a tert

- b utanol multigenerational reproductive study or an immunotoxicity study. No adverse effects were reported in one - a nd two

- generation reproductive/developmental
studies on ETBE ( Gaoua, 2004a

, b ) , and the database for ETBE does not indicate immunotoxicity ( Banton et al., 2011 ; Li et al., 2011 ). Thus, although there are some gaps in the toxicity database for

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t ert

- butanol

, the available data on tert

- butanol, informed by the data on ETBE, do not suggest that 

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9 additional studies would lead to identification of a more sensitive endpoint or a lower POD. Therefore, a database UF D of 1 was applied.
Table 2 -6 is a continuation of Table 2 -4 and Table 2 -5

, and summarizes the application of UFs to each POD to derive a candidate value for each data set. The candidate values presented in the table below are preliminary to the derivation of the organ/system

- s pecific reference values.
These candidate v alues are considered individually in the selection of a representative inhalation reference value for a specific hazard and subsequent overall RfC for tert

- bu tanol

.

**Table 2-**  **6 . Effects and corresponding derivation of candidate**  **values**

Endpoint (Sex and species) and Reference POD HE C a (mg/ m 3 ) POD type UF A UF H UF L UF S UF D Composite UF Candidate value (mg/ m 3 )

Kidney
Increased relative kidney weight ; male rat NTP (1997) 861 BMCL 10% 3 10 1 10 1 300 3 × 10 0

Increased absolute kidney weight ; m ale rat NTP (1997) 304 BMCL 10% 3 10 1 10 1 3 00 1 × 10 0

Increased relative kidney weight ; female rat NTP (1997) 1137 NOAEL 3 10 1 10 1 3 00 4 × 10 0

Increased absolute kidney weight ; female rat NTP (1997) 1137 NOAEL 3 10 1 10 1 3 00 4 × 10 0

Increased relative kidney weight ; male rat NTP (1995) 79. 6 BMCL 10% 3 10 1 1 1 3 0 3 × 10 0

Increased relative kidney weight ; female rat NTP (199 5) 231 BMCL 10% 3 10 1 1 1 3 0 8 × 10 0

Kidney inflammation ; female
rat NTP (1995) 3 59 BMCL 10% 3 10 1 1 1 3 0 1 × 10 1

Kidney transitional epithelial hyperplasia ; male rat NTP (1995) 26.1 BMCL 10% 3 10 1 1 1 3 0 9 × 10 - 1

Kidney transitional epithelial hyperplasia ; female rat NTP (1995) 638 BMCL 10% 3 10 1 1 1 3 0 2 × 10 1



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- Butyl Alcohol

**2. 2.4.**  **Derivation of Organ/System -S**  **pecific Reference Concentrations**  

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20 21 Table 2 -7 distills the candidate values from Table 2 -6 into a single value for the kidney

. O r gan or system

- speci fic reference values may be useful for subsequent cumulative risk assessments that consider the combined effect of multiple agents acting at a common site.

Kidney To xicity

For t ert

- butanol , candidate values were for several different effects in both sexes, spanning a range from 9 × 10 -1 to 2 × 10 1 mg/ m 3 , for an overall twenty

- fold range. To estimate an exposure
level below which kidney toxicity from t ert

- butanol exposure is not expected to occur, the RfC for increased incidence of transitional epithelial h yperplasia in male rats ( **9 × 10**  **-1 mg/m**  **3** ) is proposed
as the kidney

- sp ecific reference concentration for tert

- butanol , consistent with the selection of the kidney

- s pecific RfD (see Section 2.1.4). As discussed previously, u nlike kidney inflammation, this

e f

fect was observed in both sexes, with males appearing to be more sensitive than females. Additionally, it is a more specific and more sensitive indicator of kidney toxicity than the relatively

non

- sp ecific endpoint of kidney weight changes. Confidence in this kidney

- specific RfC is medium

. The PODs are based on modeled benchmark dose estimates, an d the candidate values are derived
from a well

- c onducted study, involving a sufficient number of animals per group , including both sexes, assessing a wide range of kidney endpoints , and availability of a PBPK model for route

- to

route extrapolation.

**Table 2-**  **7**

**. Organ/system** 

**- specific Rf**  **C s and proposed overall Rf**  **C for** 
tert

**- butanol** 

Effect Basis Rf C (mg/ m 3 ) Exposure description Confidence

Kidney toxicity Increased incidence of transitional epithelial hyperplasia 9 × 10 - 1 Chronic HIGH

Proposed overall Rf C Increased incidence of transitional epithelial hyperplasia 9 × 10 - 1 Chronic HIGH



### 22

**2.2.5. Selection of the Proposed Overall Reference**  **Concentration** 

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- butanol

, only kidney effects were identified as a hazard; thus, a single

organ / system

- specific reference concentration was derived. Therefore, the kidney

- specific RfC of
**9 × 10**  **-1 mg/m**  **3** is also proposed as an estimated exposure level below which deleterious effects
from t ert

- butanol exposure are not expected to occur.
The overall reference c oncentration 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

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potentially susceptible subgroups ( U.S. EPA, 2002 ) . Decisions concerning averaging exposures over 

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36 time for comparison with the RfC 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 Rf C. In the case of tert

- butanol , no specific lifestages have been identified has a potentially susceptible subgroup.

**2.2.6. Confidence Statement**  A confidence level of high, medium, o r low is assigned to the study used to derive the Rf C
, the overall database, and the Rf C it self, 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 Rf C is high

. Confidence in the principal study ( NTP, 1995 ) is high

. Th is study was well -conducted, compiled with FDA GLP regulations, involved a sufficient number of animals per group (including both sexes), and assessed a wide range of tissues and endpoints

. A lt hough there are some gaps in the toxicity database for tert

- butanol

, the se areas are informed by the data on ETBE, a parent compound of tert

- b utanol . Therefore, the confidence in the database is high. Reflecting high confidence in the principal study and h igh confidence in the database, confidence in the Rf C is high

.

**2.2.7. P**  **revious IRIS Assessment**

An inhalation assessment for tert

- bu tanol was not previously available on IRIS.

**2.2.8. U**  **ncertainties in the Derivation of the Reference Dose**  **and Reference Concentration** 

The f ollowing discussion identifies uncertainties associated with the RfD and RfC for

tert

- b utanol . To derive the RfD, the UF approach ( U.S. EPA, 2000a

, 1994 ) was applied to a POD based on kidney toxicity in rats treated c hronically . To derive the RfC, this same approach was applied

, but a PBPK model was used to extrapolate from oral to inhalation exposure . UFs were applied to the POD to account for extrapolating from an animal bioassay to human exposure, the likely existence of a diverse population of varying susceptibilities, and database deficiencies. These extrapolation s a re carried out with default approaches given the lack of data to inform individual
steps.
The database for tert

- bu tanol contains no human data on adverse health effects from subchronic or chronic exposure. Data on the effects of tert

- bu tanol are derive d from a small
database of studies in rats and mice . T he database for tert

- butanol exposure includes one lifetime bioassay

, s everal reproductive/developmental studies, and several subchronic studies. Although the database is adequate for reference value d er ivation, there is uncertainty
associated with th e lack of a comprehensive multigeneration reproductive toxicity study. Additionally, only subchronic and short

- t erm inhalation studies have been conducted , and no
chronic inhalation studies ar e available

. Developmental studies identified significant increases in

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fet al loss, decreases in fetal body weight, and possible increases in skeletal variations in exposed 

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37 offspring or pups. However, effects were not always consistent across exposure routes, and sign ificant material toxicity was present whenever developmental effects were observed. The toxicokinetic and toxicodynamic differences for tert

- butanol between the animal species in which the POD was derived and humans are unknown. Although sufficient information is available to develop a PBPK model in rats to evaluate difference across routes of exposure, the

tert

- b utanol database lacks an adequate model that would inform potential interspecies differences. Generally, it was found that rats appear more susceptible than mice, and males appear more susceptible than females to tert

- bu tanol toxicity. However, the underlying mechanistic basis of
these ap parent difference s is not understood. Most importantly, it is unknown which animal species a nd/or sexes may be more comparable to humans.



## 2.3. **ORAL SLOPE FACTOR FOR CANCER**
The carcinogenicity assessment provides information on the carcinogenic hazard potential of the substance in question and quantitative estimates of r isk from oral and inhalation exposure
that m ay be derived

. Quantitative risk estimates may be derived from the application of a low

- dose extrapolation procedure

. I f derived, the oral slope factor is a plausible upper bound on the estimate of risk per mg/ k g

- day of oral exposure

.

**2.3.1. A**  **nalysis of Carcinogenicity Data**

As noted in Section 1.2.2

, EPA concluded that there is “ suggestive evidence of carcinogenic potential ” f or tert

- butanol

. The Guidelines for Carcinogen Risk Assessment ( U.S. EPA, 2005a ) state:

When there is suggestive evidence, the Agency generally would not attempt a do se

response assessment, as the nature of the data generally would not support one; however when the evidence includes a well

- c onducted study, quantitative analysis may be useful for some purposes, for example, providing a sense of the magnitude and uncerta i nty of potential risks, ranking potential hazards, or setting research priorities.

The only data available on potential carcinogenicity w as derived from the 2

- year drinking water study in rats and mice by ( NTP, 1995 ) . This stud y was considered suitable for dose

- response analysis. It was conducted in accordance with Food and Drug Administratio n ( FDA) Good Laboratory Practice (GLP) R egulations, and all aspects were subjected to retrospective quality assurance audits. The study included histological examinations for tumors in many different tissues, contained three exposure levels and controls, contained adequate numbers of animals per dose group (~50/sex/group), treated animals for up to 2 years, and included detailed reporting of methods and results. Additionally, the renal tumors were re

- e xamined by a Pathology Working Group ( Hard et al., 2011 )

. Dose

- r elated increasing trends in tumors were noted at the following sites:

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• R en al tubule adenoma s and carcinoma s in male rats ; and 

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35 • Thyroid follicular adenomas in female mice.

These tu mors were statistically significantly increased by pairwise comparison (Fisher
exact test, p ≤ 0.05) and by trend test (Cochran

- Armitage trend test, p ≤ 0.05). Based on analysis of mode of action data, it was concluded that processes other than α 2u - globu lin nephropathy are likely responsible for the male rat renal tumors, so these tumors m ay be suitable for quantitative analysis ( U.S. EPA, 1991a ) . Additionally, a thyroid follicular carcinoma was observed in male mice, so it is possible t hat the thyroid follicular adenomas in female mice could progress to malignant form. Therefore, the thyroid follicular adenomas in female mice m ay also be considered suitable for

quan ti t ative analysis. Considerin g these data along with the uncertainty associated with the suggestive nature of the weight of evidence, EPA concluded that quantitative analyses may be useful for providing a sense of the magnitude of potential carcinogenic risk.

**2.3.2. D**  **ose -**  **Response Analysis**

**— Adjustments and Extrapolations Methods** 

The U.S. EPA G uidelines for Carcinogen Risk Assessment ( U.S. EPA, 2005a ) recommend s that the method used to characterize and quantify cancer risk from a chemical is determined by what is known about the MOA of the carcinogen and the shape of the cancer dose

- r esponse curve . The linear approa ch is recommended if the MOA of carcinogenicity has not been established ( U.S. EPA,
2005a )

. In the case of tert

- butanol , the mode s of carcinogenic action for renal tubule and thyroid
follicular t umors are not fully understood (see Section 1.2.2). Therefore, a linear low

- dose extrapolation approach was used to estimate human carc in ogenic risk associated with tert

- butanol

exposure. The m odeled tert

- butanol PODs were scaled to HEDs according to EPA guidance ( U.S. EPA,
2011

, 2005a ) . In particular, the BMDL was converted to an HED by assuming that doses in animals and humans are toxicologically equivalent when scaled by body weight raised to the 3 / 4 power

. Standard body weights of 0 .025 for mice, 0.25 kg for rats

, and 70 kg for humans was used ( U.S. EPA,
1988 )

. The following formula was used for the conversion of oral BMDL to oral HED for rat endpoints :

S caled HED in mg/kg

- day = (BMDL in mg/kg

- day) × ( animal body weight / 70 ) 1/4 = (BMDL in mg/kg

- day) × 0.24

Details of the modeling and the model selection process can be found in Appendix C of th e Supplemental Information. PODs for estimating low

- d ose risk were identified at doses at the lower end of the observed data, generally corresponding to 10% extra risk. Because initial modeling overestimated the control, due to the non

- m onotonicity of the observed dose

- response, the POD

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was derived after dropping the highest exposure group ( U.S. EPA, 2012b ) . The highest exposure 

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20 group also had increased mortality, which may in part explain the observed non

- monotonicity.

**2.3.3. Derivation of the Oral Slope Factor** 

The PODs estimated for each tumor site are summarized in Table 2 -8 . The lifetime oral cancer slope factor for humans is defined as the slope of the line from the lower 95% bound on the exposure at the POD to the control response (slope factor = 0.1/BMDL 10 ). This slope, a 95% upper
confidence limit

, r epresents a plausible upper bound on the true risk. Using linear extrapolation
from the BMDL 10 , human equivalent oral slope factors were derived for each species /tumor site
combination and are listed in Table 2 -8

. The oral slope factors derived from the NTP (1995) bioassay differ by twenty

- fold, depending on the species and tumor site. T h e most sensitive endpoint of renal tumors was used to derive the oral slope factor b ecause there are no data to support any one result as most relevant for extrapolating to humans. T wo slope factors were derived for this endpoint from the NTP (1995) bioassay, one based on the original reported incidences and the other based on the Hard et al. (2011) reanalysis. The two estimates differ ed by less than 20%, and round ed to the same number at one significant figure . H owever, the Hard et al. (2011) reanalysis is considered p referable, as it is based on a PWG analysis

. T herefore, t he recommended slope factor for providing a sense of the magnitude of potential carcinogenic risk associated with lifetime oral exposure to tert

- b utanol is
**1 × 10**  **-2 per**  **mg/kg**

**- day**  , based on the renal tubule tumor response in male F344 rats

.

**Table 2-**  **8. Summary of**  **the oral slope factor derivations**

Tumor Species/Sex Selected Model BMR BMD ( mg/kg

d ) POD= BMDL ( mg/kg

- d ) BMDL HED a (mg/kg

- d) Slope factor b (mg/kg

- day) -1

Ren al tubule adenoma or carcinoma Male F344
rat ; dose as administered 1° Multistage (high dose dropped) 10% 70 42 10.1 1 × 10 - 2

Ren al tubule adenoma or carcinoma [ Hard et al. (2011) reanalysis ] Male F344
rat ; dose as administered 1° Multistage (high dose dropped) 10% 54 36 8.88 1 x 10 - 2

Thyroid follicular cell adenoma Female B6C3F1

mouse 3° Multistage 10% 2002 1437 201 5 × 10 - 4



### 21 22 23 24 a HED PODs were calculated using BW 3/4 scaling ( U.S. EPA, 2011 ).
b Human equivalent slope factor = 0.1/BMDL 10HED ; see Appendix C of the Supplemental Information for details of

modeling results.

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**2. 3.4.**  **Uncertainties in the Derivation of the Oral Slope Factor** 

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18 19 There is uncertainty when extrapolating data from animals to estimate potential cancer risks to human populations from exposure to tert

- bu tanol (s ee Table 2 -9 )

. Uncertainty in the magnitude of the recommended oral slope factor is reflected to some extent in the range of slope factors; the oral slope factor based on the male rat data was about t w enty

- fold higher than the oral slope factor based on female mouse data ( T able 2 -9 ). These comparisons show that the selection of target organ, animal species, and in terspecies extrapolation can impact the oral cancer risk

estimate. Although the thyroid follicular cell tumors occurred in male and female mice, high mortality in high

- d ose male mice limited the usefulness of the data. Renal tubule tumors occurred

in ma l e rats, but not female rats. Therefore, only the data in male rats and female mice were available for deriving the oral slope factor. There are no other chronic studies to replicate these findings or that examined other animal models. There are no data i n humans to support the tumors observed in animals. Although changing the methods used to derive the oral slope factor could change the results, standard practices were used due to the lack of a mouse or human PBPK model or specific MOA to indicate other m ethods would be preferable. A dditionally, c onsidering the uncertainty associated with the suggestive nature of the weight of evidence, the oral slope factor is recommended only f or providing a sense of the magnitude of potential carcinogenic risk.

**Table 2-**  **9**

**. Summary of uncertaint**  **ies in the derivation of cancer risk values for** 
tert

**- butanol** 

Consideration and Impact on Cancer Risk Value Decision Justification and Discussion
Selection of target organ ↓ oral slope factor, up to twenty

fold, if renal tumors not selected. The kidney was selected as the target organ

. As there are no data to support any one result as most relevant for extrapolating to humans, the most sensitive result for k idney renal tubular ad enomas and carcinomas was used to derive the oral slope factor. However, the overall evidence for carcinogenicity was considered “suggestive.”

Selection of data set Unknown change in oral slope factor, since no other studies are available. NTP (1995) as principal oral (drinking water) study to derive cancer risks for humans. NTP (1995) was a well

- conducted study. It was also the only bioassay available

. Additi onal bioassays might add support to the findings or provide results for different (possibly lower) doses, which may affect the oral slope factor

. Selection of extrapolation approach (Selection of extrapolation approach could change the recommended

cancer risk values.) Oral data used for OSF. No extrapolation methods were used.

Selection of dose metric Alternatives could ↓ or ↑ slope factor Used administered dose converted to HED units. Additional runs using the administered dose without conversion to HED units were also conducted
, result ing in a similar oral slope

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Consideration and Impact on Cancer Risk Value Decision Justification and Discussion
factor. For rats, a PBPK model of internal dose was available, but the POD changed by less than 1.2

- fold when modeling was based on internal doses. For mice, no PBPK mod el was available, so using a PBPK model for determining internal doses could have an unknown e ffect on the estimated OSF value.
Interspecies extrapolation of dosimetry and risk Alternatives could ↓ or ↑ slope factor (e.g., 3.5

- fold ↓ [scaling by body weight] or ↑ 2

- fold [scaling by BW2/3]) The default approach of body weight 3/4 was used. There are no data to suggest an alternative approach. Because the dose metric was not an area under the curve, BW 3/4 scaling was used to calculate equivalent cumulati ve exposures for estimating equivalent human risks. While the true human correspondence is unknown, this overall approach is expected to neither over - nor underestimate human equivalent risks.

Dose

- response modeling Alternatives could ↓ or ↑ slope facto r Used multistage dose

response model to derive a BMD and BMDL. No biologically based models for tert

- butanol were available. The multistage model has biological support and is the model most consistently used in EPA cancer assessments.

Low

- dose extrapolation ↓ cancer risk estimate would be expected with the application of nonlinear low

- dose extrapolation Linear extrapolation of risk in low

- dose region used. Linear low

- dose extrapolation for agents without a known MOA is supported.

St atistical uncertainty at POD ↓ oral slope factor 1. 7- fold if BMD used as the POD rather than BMDL BMDL (preferred approach for calculating plausible upper bound slope factor) Limited size of bioassay results in sampling variability; lower bound is 95% CI on administered exposure at 10% extra risk of renal tumors.
Sensitive subpopulations ↑ oral slope factor to unknown

extent No sensitive populations have been identified. No c hemical

- specific data are available to determine the range of human toxicodynamic variability or sensitivity
, including the susceptibility of children

. B ecause determination of a mutagenic MOA has not been made , an age

- specific adjustment factor is not ap pl ied

. 

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7 A cancer assessment for tert

- butanol was not previously available on IRIS.



## 2.4. **INHALATION UNIT RISK FOR CANCER**
The carcinogenicity assessment provides information on the carcinogenic hazard potential
of the s ubstance in question and quantitative estimates of risk from oral and inhalation exposure may be derived

. Q uantitative risk estimates may be derived from the application of a low

- dose

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e xtrapolation procedure

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37 estimate of risk per μg/m 3 air breathed.

**2.4.1. Analysis of Carcinogenicity Data** 

As noted in Section 1.2.2

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

- butanol

. The Guidelines for Carcinogen Risk Assessment ( U.S. EPA, 2005a ) state:

When t h ere 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

- c onducted study, quantitative analysis may be useful for

s o me purposes. For example, it could provide a sense of the magnitude and uncertainty of potential risks, rank potential hazards, or set research priorities.

The only data available on potential carcinogenicity were from the 2

- y ear drinking water study in r ats and mice by NTP (1995) , discussed previously in Section 2.3.1 . Because a PBPK model
for the rat is available to conduct route

- to

- r oute extrapolation (discussed below), the male rat renal tubule adenoma and carcinoma data are suitable for quantitative analysis to support an inhalation
unit risk. Considering these data and uncertainty associated with the suggestive nature of the weight of evidence, EPA concluded that quantitative analyses may be useful for providing a sense of the magnitude of potential carcinogenic risk.

**2.4.2. D**  **ose Response Analysis –**  **Adjustments and Extrapolation Methods** Details of the modeling and the model selection process can be found in Appendix C of the Supplemental Information. A POD for estimating low

- d ose risk was identified at doses at the lower end of the observed data corresponding to 10% extra risk.

A PBPK model for tert

- bu tanol in rats has been developed, as described in Appendix B. Using this model, route

- to

- r oute extrapolation of the oral BMDL to derive an inhalation POD was performed as follows. First, the internal dose in the rat at the oral BMDL (assuming continuous exposure) was estimated using the PBPK model, to derive an “internal dose BMDL.” Then, the inhalation air concentration (again assuming continuous exposure) that led to the same internal dose in the rat was estimated using the PBPK model, resulting in a route

- to

- r oute extrapolated
BMCL.
A critical decision in the route

- to

- r oute extrapolation is the selection of the internal dose
metric to use that established " equivalent” oral and inhalation exposures. For tert

- butanol -induced kidney effects, the two options are the concentration of tert

- b utanol in blood and rate of tert

butanol m etabolism. Note that using the kidney concentration of tert

- butanol will lead t o the same

route

- to

- r oute extrapolation relationship as tert

- butanol in blood, since the distribution from blood to kidney is independent of route. There are no data that suggest metabolites of tert

- bu tanol mediate its renal toxicity. In the absence of e vidence that would suggest otherwise, it is assumed

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th at tert

- butanol itself is the active toxicological agent. Therefore, the concentration of tert

- butanol 

### 1
2
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17 18 19

20
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29
30
31
32
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35
36
37 in blood was selected as the dose metric to derive the BMCL. The RfC methodology provides a mechan ism for deriving a HEC from the BMCL determined from the animal data. The approach takes into account the extra

- r espiratory nature of the toxicological responses and accommodates species differences by considering blood:air partition
coefficients for tert

- b utanol in the laboratory animal (rat or mouse) and humans. According to the RfC guidelines ( U.S. EPA, 1994 )

, tert

- butanol is a Category 3 gas because extra

- respiratory effects
were observed. Kaneko et al. (2000) measured a blood:gas partition coefficient of 531 ± 102 for tert

butanol i n the male Wistar rat, while Borghoff et al. (1996) measured a value of 481 ± 29 in male
F344 rats

. A bl ood:gas partition coefficient of 462 was reported for tert

- butanol in humans ( Nihlén et al., 1995 ) . The calculation (H b/g ) A ÷ (H b/g ) H was used to calculate a blood:gas par tition coefficient ratio to apply to the delivered concentration. Because F344 rats were used in the study, the blood:gas partition coefficient for F344 rats was used. Thus, the calculation was: 481 ÷ 462 = 1.04. Therefore, a ratio of 1.0 4 was used to calc ulate the HEC

. This allowed a BMCL HEC to be derived as
follows:

BMCL HEC = BMCL ADJ (mg/m 3 ) × (interspecies conversion)

= B MCL ADJ (mg/m 3 ) × ( 481 

### ÷ 462 )

= BMCL ADJ (mg/m 3 ) × ( 1.04 )

The U.S. EPA G uidelines for Carcinogen Risk Assessment ( U.S. EPA, 2005a ) recommend that the method used to characterize and quantify cancer risk from a chemical is determined by what is known about the MOA of the carcinogen and the shape of the cancer dose

- r esponse curve . The linear approach is recommended if the MOA of carcino genicity has not been established ( U.S. EPA,
2005a )

. In the case of tert -b utanol , the mode of carcinogenic action for renal tubule tumors is not fully understood ( see Section 1.2.2). Therefore, a linear low

- dose extrapolation approach was used to estimate human carcinogenic risk associated with tert

- bu tanol exposure.

**2.4.3. I**  **nhalation Unit Risk Derivation** 

The results from route

- to

- r oute extrapolation of the male rat renal tubule tumor data are
summarized in T able 2

- 10 . The lifetime inhalation uni t risk for humans is defined as the slope of the line from the lower 95% bound on the exposure at the POD to the control response ( i nhalation unit risk = 0.1/BMC L 10 ). This slope, a 95% upper confidence limit represents a plausible upper
bound on the true risk. Using linear extrapolation from the BM CL 10 , a human equivalent inhalation unit risk was derived, as listed in T able 2

- 10

. Two inhalation unit risks were derived from the NTP (1995) bioassay : one based on the original reported incidences and one based on the Hard et al. (2011) reanalysis. The two estimates differ by les s than 20%, but the Hard et al. (2011) reanalysis is considered preferable, as it is based

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o n a PWG analysis. The refore, the recommended inhalation unit risk for providing a sense of the 

### 1
2
3
4

5

6 7 8 9 10

11

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16
17
18
19
20
21
22
23
24
25
26
27
28
29 magnitude of potential carcinogenic risk associated with lifetime inhalation exposure to

tert

- butanol is **2 × 10**  **-3 per**  **mg/ m**  **3** , or **2 × 10**  **-6 per**  **μ g/**  **m 3**  , based on the renal tubule tumor response in male F344 rats

.

**Table 2-**  **10**

**. Summary of the inhalation unit risk derivation** 

Tumor Species/Sex BMR BMD L ( mg/kg

- d ) Internal Dose a (mg/L) POD= BM CL HEC c ( mg/ m 3 ) Unit Risk b (mg/ m 3 ) -1

Ren al tubule adenoma or carcinoma Male F344
rat 10% 41.6 2.01 68.7 1 × 10 - 3

Ren al tubule adenoma or carcinoma [ Hard et al. (2011) reanalysis ] Male F344
rat 10% 36.3 1.74 59.8 2 × 10 - 3

a Average blood concentration of tert

- butanol under continuous oral exposure at the BMDL.
b Continuous inhalation human equivalent concentration that leads to the same average blood concentration of

tert

- but anol as continuous oral exposure at the BMDL.

c Human equivalent inhalation unit risk = 0.1/BMC L HEC .

**2.4.4. U**  **ncertainties in the Derivation of the Inhalation Unit Risk** There is uncertainty when extrapolating data from animals to estimate potential cancer risks to human populations from exposure to tert

- bu tanol (s ee Table 2

- 11 )

. Uncertainty in the magnitude of the recommended inhalation unit risk can be inferred to some extent from t he range
of o ral slope factors; the oral slope factor based on the male rat data was about twenty -fold higher than the oral slope factor based on female mouse data ( T able 2 -9 ). These comparisons show that the selection of target organ, animal spec ie s, and interspecies extrapolation can impact the
inhalation unit r isk estimate. Although the thyroid follicular cell tumors occurred in male and female mice, high mortality in high

- d ose male mice limited the usefulness of the data. Additionally,

no PBP K model was available in mice for use in route

- to

- r oute extrapolation, so these data could not be used to estimate an inhalation unit risk. Renal tubule tumors occurred in male rats, but not female rats. Therefore, only the data in male rats were available fo r deriving the inhalation unit risk. There are no other chronic studies to replicate these findings or that examined other animal models. There are no data in humans to support the tumors observed in animals. Although changing the methods used to deri ve the inhalation unit risk could change the results, standard practices were used due to the lack of a mouse or human PBPK model or specific MOA to indicate
other methods wh ich would be preferable. Additionally, c onsidering the uncertainty associated

wi t h the suggestive nature of the weight of evidence, the inhalation unit risk is recommended only for providing a sense of the magnitude of potential carcinogenic risk.

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**T able**  **2- 11** 

**. Summary of uncertaint**  **ies in the derivation of cancer risk values**  

### 1 2 **for** tert

**- butanol** 

Consideration and Impact on Cancer Risk Value Decision Justification and Discussion
Selection of target organ Inhalation unit risk may change by an unknown amount if a PBPK model to extrapolate mouse thyroid tumors to the inhalation route were available. The kidney was selected as the target organ

. No PBPK model to extrapolation mouse thyroid tumors was available. Additionally, the overall evidence for carcinogenicity was considered “suggestive .”

Selection of data set Unknown change in inhalation unit risk, since no other studies are available. NTP (1995) as principal oral (drinking water) study to derive cancer risks for humans. NTP (1995) was a well

- conducted study. It was also the only bioassay available

. Additional bioassays might add support to the findings or provide results for different (possibly lower) doses, which may affect the inhalation unit risk

. Selection of extrapolation approach Different PBPK model could ↓ or ↑ inhalation unit risk. PBPK model

- based extrapolation of oral data used for inhalation unit risk. PBPK model accurately predicted tert

butanol toxicokinetics. Data and model pred ictions were within 2- fold of each other.

Selection of dose metric Alternatives could ↓ or ↑ inhalation unit risk

. Used tert

- butanol concentration in blood as the dose metric for route

to

- route extrapolation, converted to HEC

. In the absence of evidence that would suggest that metabolites of tert

- butanol are responsible for carcinogenicity, it is assumed that tert

- butanol itself is the active toxicological agent. An alternative dose metric of tert

- butanol metabolism would resul t in a 1.2

- fold decrease in the inhalation unit risk.
Interspecies extrapolation of dosimetry and risk Alternatives could ↓ or ↑ inhalation unit risk.

. The default approach for a Category 3 gas was used. There are no data to suggest an alternative approach. While the true human
correspo ndence is unknown, this overall approach is expected to neither over - nor underestimate human equivalent risks.

Dose

- response modeling Alternatives could ↓ or ↑ inhalation unit risk

. Used multistage dose

response model to derive a BMD and BMDL. No biologically based models for tert

- butanol were available. The multistage model has biological support and is the model most consistently used in EPA cancer assessments.

Low

- dose extrapolation ↓ cancer risk estimate wo uld be expected with the application of nonlinear low

- dose extrapolation

. Linear extrapolation of risk in low

- dose region used. Linear low

- dose extrapolation for agents without a known MOA is supported.

Statistical uncertainty at POD ↓ inhalation unit risk 1. 7- fold if the BMD used to derive the inhalation POD rather than BMDL

. BMDL (preferred approach for calculating plausible upper bound) Limited size of bioassay results in sampling variability; lower bound is 95% CI on administered exposure at 10% extra risk of renal tumors.
Sensitive subpopulations No sensitive populations No chemical

- specific data are available to

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Consideration and Impact on Cancer Risk Value Decision Justification and Discussion
↑ inhalation unit risk to unknown
extent

. have been identified. determine the range of human toxicodynamic variability or sensitivity
, including the susceptibility of children

. B ecause determination of a mutagenic MOA has not been made , an age

- specific adjustment factor is no t appl ied

. 

### 1

**2.4.5. Previous IRIS Assessment: Inhalation Unit Risk**  

### 2

3

4

5
6
7
8
9
10
11 A cancer assessment for tert

- butanol was not previously available on IRIS.



## 2.5. **APPLICATION OF AGE**

**- DEPENDENT ADJUSTMENT FACTORS** 

As discussed in the Supplemental Guidance for Assessing Susceptibility from Early

- Life Exposure to Carcinogens ( U.S. EPA, 2005b ) , either default or chemical

- specific age

- dependent adjustment factors (ADAFs) are applied to account for early

- l ife exposure to carcinogens that act through a mu tagenic mode of action. Because chemical

- s pecific life

- stage susceptibility data for cancer are not available, and because the mode of action for tert

- bu tanol carcinogenicity is not
known ( see Section 1.1.4), ADAFs were not applied.

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### 1



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IOM (Institute of Medicine). (2008). Improving the presumptive disability decision

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- 11 8). Lyon, F rance: International Agency for Research on Cancer. http://apps.who.int/bookorders/anglais/detart1.jsp?sesslan=1&codlan=1&codcol=73&cod cch=147 Takahashi, K; Lindamood, C; Maronpot, R. (1993). Retrospective study of possible alpha

- 2 mu

globulin nephropathy and associated cell proliferation in male Fischer 344 ra ts dosed with
t- b utyl alcohol. Environ Health Perspect 101: 281

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- buty l ether and its metabolite to human leukemia cells]. Zhonghua Yufang Yixue Zazhi 31: 334

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- 98 /003). Washington, DC: U.S. Environmental Protection Agency, Risk Assessment Forum. http://www.epa.gov/iris/backgrd.html U.S. EPA (U.S. Environmental Protection Agency). (1986b). Guidelines for the health risk assessment of chemical mixtures. (EPA/630/R

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- 87/008). Cincinnati, OH: U.S. Environmental Protection Agency, National Center for Environmental Assessment

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- globulin: Association with chemically induced renal toxicity and neoplasia in the male rat. (EPA/625/3

- 91/019F). Washington, DC: U.S. Environmental Protection Agency, National Center for Environmen tal Assessment. http://www.ntis.gov/search/product.aspx?ABBR=PB92143668 U.S. EPA (U.S. Envi ronmental Protection Agency). (1991b). Guidelines for developmental toxicity risk assessment. (EPA/600/FR

- 91/001). Washington, DC: U.S. Environmental Protection Agency, Risk Assessment Forum. http://www.epa.gov/raf/publications/guidelines

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- assessment.htm U.S. EPA (U.S. Environmental Protection Agency). (1 994). Methods for derivation of inhalation reference concentrations and application of inhalation dosimetry. (EPA/600/8

- 90/066F). Research Triangle Park, NC: U.S. Environmental Protection Agency, Environmental Criteria and Assessment Office. http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=71993 U.S. EPA (U.S. Environmental Protection Agency). (1996). Guidelines for reproductive toxicity risk assessment. (EPA/630/R

- 9 6/009). Washington, DC: U.S. Environmental Protection Agency, Risk Assessment Forum. http://www.epa.gov/raf/pu blications/pdfs/REPRO51.PDF U.S. EPA (U.S. Environmental Protection Agency). (1998a). Assessment of thyroid follicular cell tumors [EPA Report]. (EPA/630/R

- 97/0 02). Washington, DC. U.S. EPA (U.S. Environmental Protection Agency). (1998b). Guidelines for neurotoxicity risk assessment. (EPA/630/R

- 95/001F). Washington, D C: U.S. Environmental Protection Agency, Risk Assessment Forum. http://www.epa.gov/raf/publications/pdfs/NEUROTOX.PDF U.S. EPA (U.S. Environmental Protection Agency). (2000a). Science policy council handbook: risk characterization. (EPA/100/B

- 00/002). Washington, D.C.: U.S. Environmental Protection Agency, Office of Science Policy. http://www.epa.gov/osa/spc/pdfs/rchandbk.pdf U.S. EPA (U.S. Environmental Protection Agency). (2000b). Supplementary guidance for conducting health risk assessment of chemical mixtures. (EPA/630/R

- 0 0/002). Washington, DC: U.S. Environmental Protection Agency, Risk Assessment Forum. http ://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=20533 U.S. EPA (U.S. Environmental Protection Agency). (2002). A review of the reference dose and reference conc entration processes. (EPA/630/P

- 02/002F). Washington, DC: U.S. Environmental Protection Agency, Risk Assessment Forum. http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=51717 U.S. EPA (U.S. Environmental Protection Agency). (2005a). Guidelines for carcinogen risk assessment. (EPA/630/P

- 03/001F). Washington, DC: U.S. Environmental Protection Agency, Risk Assessment Forum. http://www.epa.gov/cancerguidelines/ U.S. EPA (U.S. Environmental Protection Agency). (2005b). Supplemental guidance for assessing susceptibility from early

- l ife exposure to carcinogens. (EPA/630/R

- 03/003F). Washington, DC: U.S. Environmental Protection Agency, Risk Assessment Forum. http://www.epa.gov/raf/publications/pdfs/childrens_supplement_final.pdf U.S. EPA (U.S. Environmental Protection Agency). (2006a). Approaches for the application of physiologically based pharmacokinetic (PBPK) models and supporting data in risk assessment (Final Report) [EPA Report]. (EPA/600/R

- 05/043F). Washington, DC: U.S. Environmental Protection Agency, National Center for Environmental assessment. http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=157668 

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U. S. EPA (U.S. Environmental Protection Agency). (2006b). A framework for assessing health risk of 

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37 environmental exposures to children. (EPA/600/R

- 05/093F). Washington, DC: U.S. Environmental Protection Agency, National Center for Environmental Assessment

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- 0 9/019F). Research Triangle Park, NC. http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=218686 U.S. EPA (U.S. Environmental Protection Agency). (2011). Recommended use of body weight 3/4 as the default method in derivation of the oral reference dose. (EPA/100/R11/0001). Washington, DC: U.S. Environmental Protection Agency, Risk Assessment Forum. http://www.epa.gov/raf/publications/interspecies

- extrapolation.htm U.S. EPA (U.S. Enviro nmental Protection Agency). (2012a). Advances in inhalation gas dosimetry for derivation of a reference concentration (RFC) and use in risk assessment. (EPA/600/R

12/044). Washington, DC. http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=244650 U.S. EPA (U.S. Environmental Protection Agency). (2012b). Benchmark dose technical guidance. ( EP A/100/R

- 12/001). Washington, DC: Risk Assessment Forum. http://www.epa.gov/raf/publications/pdfs/benchmark_dose_guidance.pdf U.S. EPA (U.S. Environmental Protection Agency). (2012c). Releases: Facility Report. Toxics Release Inventory. Available online at http ://iaspub.epa.gov/triexplorer/tri_release.chemical U.S. EPA (U.S. Environmental Protection Agency). (2014). Chemical Data Reporting 2012, reported in the Che m ical Data Access Tool. Available online at http://www.epa.gov/oppt/cdr/index.html Webb, DR; Kanerva, RL; Hysell, DK; Alden, CL; Lehman

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

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