EPA/635/R-02/006

TOXICOLOGICAL REVIEW

OF

Phenol

(CAS No. 108-95-2)

In Support of Summary Information on the

Integrated Risk Information System (IRIS)

September 2002

U.S. Environmental Protection Agency

Washington D.C.
 DISCLAIMER

Mention of trade names or commercial products does not constitute endorsement or

recommendation for use. Note: This document may undergo revisions in the future. The most

up-to-date version will be made available electronically via the IRIS Home Page at

http://www.epa.gov/iris.
iii

 

CONTENTS - TOXICOLOGICAL REVIEW FOR PHENOL 

(CAS No. 108-95-2)

Foreword . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . vi

Authors, Contributors and Reviewers . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . vii

1. INTRODUCTION ......................................................1

2. CHEMICAL AND PHYSICAL INFORMATION RELEVANT TO

ASSESSMENTS . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 2

3. TOXICOKINETICS RELEVANT TO ASSESSMENTS ......................5

3.1 Absorption . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 5

3.2 Distribution . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 10

3.3 Metabolism . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 12

3.4 Excretion . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 23

4. HAZARD IDENTIFICATION . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 24

4.1 Studies in Humans - Epidemiology, Case Reports, Clinical Controls . . . . . 25

4.1.1 Oral . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 25

4.1.2 Inhalation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 27

4.2 Pre-chronic, Chronic Studies and Cancer Bioassays in Laboratory Animals

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32

4.2.1 Oral . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 32

4.2.2 Inhalation . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 52

4.2.3 Dermal . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 60

4.3 Reproductive/Developmental Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 61

4.4 Other Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 72

4.4.1 Initiation/Promotion Studies, Other Short-term Tumorigenicity

Assays, and Cancer Mechanism Studies . . . . . . . . . . . . . . . . . . . . . . . 72

4.4.2 Genotoxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 75

4.4.3 Neurological Effects . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 83

4.4.4 Immunotoxicity . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 84

4.4.5 Other Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 85

4.5 Synthesis and Evaluation of Major Noncancer Effects and Mode of Action

. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 86

4.6 Weight of Evidence Evaluation and Cancer Characterization - Synthesis of

Human, Laboratory Animal and Other Supporting Evidence, Conclusions

about Human Carcinogenicity, and Likely Mode of Action ............. 92

4.7 Susceptible Populations ........................................... 95

4.7.1 Possible Childhood Susceptibility ............................. 95
iv

 

4.7.2 Possible Gender Differences ................................. 96

5. DOSE RESPONSE ASSESSMENTS ...................................... 96

5.1 Oral Reference Dose (RfD) ....................................... 97

5.1.1 Choice of Principal Study and Critical Effect .................. 97

5.1.2 Method of Analysis - Benchmark dose ....................... 100

5.1.3 RfD Derivation ........................................... 100

5.2 Inhalation Reference Concentration (RfC) ......................... 102

5.2.1 Choice of Principal Study and Critical Effect ................. 102

5.2.2 Methods of Analysis ...................................... 103

5.3 Cancer Assessment ............................................ 103

6. MAJOR CONCLUSIONS IN CHARACTERIZATION OF HAZARD AND DOSERESPONSE ......................................................... 104

6.1 Human Hazard Potential ........................................ 104

6.1.1 Oral Noncancer .......................................... 104

6.1.2 Inhalation Noncancer ..................................... 106

6.1.3 Cancer .................................................. 107

6.2 Dose-response .................................................. 107

7.0 REFERENCES ....................................................... 111

Appendix A. Summary of External Peer Review Comments and Disposition. . . . . . . . . . 124

Appendix B. Benchmark Dose Modeling Results. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 129

Appendix C. Benchmark Dose Modeling Output. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 133
v

 

List of Tables and Figures

Table 1. Physical Properties and Chemical Identity of Phenol . . . . . . . . . . . . . . . . . . . . . . . 4

Table 2. Summary of Oral Toxicity Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 34

Table 3. Total Activity Counts in Rats Provided Phenol in Drinking Water 

(ClinTrials BioResearch, 1998) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 41

Table 4. Individual Data on Dehydration and Week 4 Motor Activity in Rats Provided

Phenol in Drinking Water

(ClinTrials BioResearch, 1998) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 43

Table 5. Effects of Phenol Exposure on Spleen Cellularity

and Selected Blood Parameters in Mice and Rats . . . . . . . . . . . . . . . . . . . . . . . . . . . 47

Table 6. Effects of Phenol Exposure on Spleen Cellularity and Selected Blood Parameters

in Mice and Rats . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 49

Table 7. Summary of Inhalation Toxicity Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 54

Table 8. Selected Results of Two-Generation Drinking Water Study 

(Ryan et al., 2001; IIT Research Institute, 1999) . . . . . . . . . . . . . . . . . . . . . . . . . . . . 63

Table 9. Key Results in Argus Research Laboratories, (1997) Rat Developmental Toxicity

Study . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 67

Table 10. Key Results from Developmental Toxicity Study in Rats

Administered Phenol by Gavage (NTP, 1983a) . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 69

Table 11. Summary of Genotoxicity Studies . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 77

Figure 1. Metabolism of Phenol . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 13

Figure 2. Plot of severity with dose for drinking water or gavage .................... 90
vi

FOREWORD

The purpose of this Toxicological Review is to provide scientific support and rationale

for the hazard identification and dose-response assessment in IRIS pertaining to chronic

exposure to phenol. It is not intended to be a comprehensive treatise on the chemical or

toxicological nature of phenol.

In Section 6, EPA has characterized its overall confidence in the quantitative and

qualitative aspects of hazard and dose-response. Matters considered in this characterization

include knowledge gaps, uncertainties, quality of data, and scientific controversies. This

characterization is presented in an effort to make apparent the limitations of the assessment and

to aid and guide the risk assessor in the ensuing steps of the risk assessment process. 

For other general information about this assessment or other questions relating to IRIS,

the reader is referred to EPA’s IRIS Hotline at 301-345-2870.
vii

AUTHORS, CONTRIBUTORS, AND REVIEWERS

Chemical Manager/Author

Monica A. Barron

Office of Solid Waste

U.S. Environmental Protection Agency

Washington, DC

Contract Authors

Lynne Haber

Toxicology Excellence for Risk Assessment

Cincinnati, OH

Andrew Maier

Toxicology Excellence for Risk Assessment

Cincinnati, OH

Jay Zhao

Toxicology Excellence for Risk Assessment

Cincinnati, OH

Michael Dourson

Toxicology Excellence for Risk Assessment

Cincinnati, OH

Reviewers

This document and summary information on IRIS have received peer review by both

EPA scientists and independent scientists external to EPA. Subsequent to external review and

incorporation of comments, this assessment underwent an Agency-wide review process whereby

the IRIS Program Manager achieved a consensus approval among the Office of Research and

Development; Office of Air and Radiation; Office of Prevention, Pesticides, and Toxic

Substances; Office of Solid Waste and Emergency Response (OSWER) of Water; Office of

Policy, Economics, and Innovation; Office of Children’s Health Protection; Office of

Environmental Information; and the Regional Offices.

Internal EPA Reviewers

Dorothy Canter

OSWER
viii

Nancy Chiu

Office of Water, Health and Ecological Criteria Division (HECD)

Annie Jarabek 

National Center for Environmental Assessment 

Edward Ohanian

Office of Water/HECD

Diana Wong

Office of Water/HECD

Robert MacPhail (reviewed portions)

National Health and Environmental Effects Research Laboratory

Selene Chou

Agency for Toxic Substances Disease Registry

External Peer Reviewers

Rolf Hartung

University of Michigan

Michele Medinsky

ToxCon

Anthony Scialli

Department of Obstetrics and Gynecology

Georgetown University Medical Center

Summaries of the external peer reviewers’ comments and the disposition of their

recommendations are presented in Appendix A.
1

1. INTRODUCTION

This document presents background and justification for the hazard and dose-response

assessment summaries in U.S. Environmental Protection Agency’s (EPA’s) Integrated Risk

Information System (IRIS). IRIS summaries may include an oral reference dose (RfD), an

inhalation reference concentration (RfC), and a carcinogenicity assessment.

The RfD and RfC provide quantitative information for noncancer dose-response

assessments. The RfD is based on the assumption that thresholds exist for certain toxic effects

such as cellular necrosis but may not exist for other toxic effects such as some carcinogenic

responses. It is expressed in units of milligrams per kilogram per day (mg/kg-day). In general,

the RfD is 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 noncancer effects during a lifetime. The inhalation RfC is

analogous to the oral RfD, but it provides a continuous inhalation exposure estimate. The

inhalation RfC considers toxic effects for the respiratory system (portal of entry) and effects

peripheral to the respiratory system (extrarespiratory or systemic effects). It is generally

expressed in units of milligrams per cubic meter (mg/m3 ).

The carcinogenicity assessment provides information on the carcinogenic hazard

potential of the substance in question and quantitative estimates of risk from oral exposure and

inhalation exposure. The information includes a weight-of-evidence judgment of the likelihood

that the agent is a human carcinogen and the conditions under which the carcinogenic effects

may be expressed. Quantitative risk estimates are presented in three ways. The slope factor is

the result of application of a low-dose extrapolation procedure and is presented as the risk per

mg/kg/day. The unit risk is the quantitative estimate in terms of either risk per μg/L drinking

water or risk per μg/m3

 air breathed. Another form in which risk is presented is a drinking water

or air concentration that provide cancer risks of 1 in 10,000 1 in 100,000 or 1 in 1,000,000. 

Development of these hazard identification and dose-response assessments for phenol has

followed the general guidelines for risk assessment as set forth by the National Research Council

(1983). EPA guidelines that were used in the development of this assessment may include the

following: Guidelines for Carcinogen Risk Assessment (U.S. EPA,1986a), Guidelines for the

Health Risk Assessment of Chemical Mixtures (U.S. EPA, 1986b), Guidelines for Mutagenicity

Risk Assessment (U.S. EPA, 1986c), Guidelines for Developmental Toxicity Risk Assessment
2

(U.S. EPA, 1991), Proposed Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1996a),

Guidelines for Reproductive Toxicity Risk Assessment (U.S. EPA, 1996b), Guidelines for

Neurotoxicity Risk Assessment (U.S. EPA, 1998a); Recommendations for and Documentation of

Biological Values for Use in Risk Assessment (U.S. EPA, 1988); (proposed) Interim Policy for

Particle Size and Limit Concentration Issues in Inhalation Toxicity (U.S. EPA, 1994a); Methods

for Derivation of Inhalation Reference Concentrations and Application of Inhalation Dosimetry

(U.S. EPA, 1994b); Peer Review and Peer Involvement at the U.S. Environmental Protection

Agency (U.S. EPA, 1994c); Use of the Benchmark Dose Approach in Health Risk Assessment

(U.S. EPA, 1995); Draft Revised Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999);

Science Policy Council Handbook: Peer Review (U.S. EPA, 1998b, 2000a); Science Policy

Council Handbook: Risk Characterization (U.S. EPA, 2000b).

The literature search strategy employed for this compound was based on the CASRN and

at least one common name. At a minimum, the following databases were searched: RTECS,

HSDB, TSCATS, CCRIS, GENETOX, EMIC, EMICBACK, DART, ETICBACK, TOXLINE,

CANCERLINE, MEDLINE, and MEDLINE backfiles. Any pertinent scientific information

submitted by the public to the IRIS Submission Desk was also considered in the development of

this document. The literature search was conducted in June 1999; selected key articles published

after that date are also included.

2. CHEMICAL AND PHYSICAL INFORMATION RELEVANT TO ASSESSMENTS

Phenol is a monosubstituted aromatic hydrocarbon. In its pure state, it exists as a

colorless or white solid. This pure compound is mixed with water and commercially sold as a

liquid product. Phenol gives off a sweet, acrid smell detectable to most people at 40 ppb in air

and at about 1–8 ppm in water (ATSDR, 1998). It evaporates more slowly than water and is

moderately soluble in water. Phenol is also combustible. 

Phenol is produced through both natural and anthropogenic processes. It is naturally

occurring in some foods, in human and animal wastes, and in decomposing organic material, and it

is produced endogenously in the gut from the metabolism of aromatic amino acids. Phenol has been

isolated from coal tar, but it is now synthetically manufactured. Currently, the largest use of phenol

is as an intermediate in the production of phenolic resins, which are used in the plywood, adhesive,
3

construction, automotive, and appliance industries. Phenol is also used in the production of

synthetic fibers such as nylon and for epoxy resin precursors such as bisphenol-A. Phenol is toxic to

bacteria and fungi, and it is used as a slimicide and disinfectant. Because of its anesthetic effects,

phenol is used in medicines such as ointments, ear and nose drops, cold sore lotions, throat lozenges

and sprays (such as those sold under the Cepastat® and Chloraseptic® labels), and antiseptic lotions.

The greatest potential source of exposure to phenol is in the occupational setting, where

phenol is used in manufacturing processes. People are also exposed via consumer products, such as

medicines and lotions, and some foods and tobacco smoke. Phenol has been found in drinking

water.

The physical and chemical properties of phenol are shown in Table 1.
4

OH

Table 1. Physical Properties and Chemical Identity of Phenol

CAS Registry Number 108-95-2 Lide, 1993

Synonym(s) Benzenol, hydroxybenzene,

monophenol, oxybenzene,

phenyl alcohol, phenyl

hydrate, phenyl hydroxide

ATSDR, 1998

Registered trade name(s) Carbolic acid, phenic acid,

phenic alcohol

ATSDR, 1998

Melting point, N

C 43 Lide, 1993

Boiling point, N

C 181.8 Lide, 1993

Vapor pressure, at 25 NC 0.3513 HDSB, 1996

Density, at 20 NC relative to

the density of H2O at 4 N

C

1.0576 Lide, 1993

Flashpoint (open cup) 85 N

C ATSDR, 1998

Water solubility, g/L at 25 NC 87 Lide, 1993

Log KOW 1.46 HDSB, 1996

Odor threshold 0.047 ppm (0.18 mg/m3) -

 100% response

0.006 ppm (0.02 mg/m3

) -

 sensitive

U.S. EPA, 1986d

Molecular weight 94.12 Calculated

Conversion factors 1 ppm (v/v) = mg/m3 x 0.260

1 mg/m3

 = ppm (v/v) x 3.85

Calculated

Empirical formula C6H6O Lide, 1993

Chemical structure
5

3. TOXICOKINETICS RELEVANT TO ASSESSMENTS

Phenol is readily absorbed by the inhalation, oral, and dermal routes. The portal-of-entry

metabolism for the inhalation and oral routes appears to be extensive and involves sulfate and

glucuronide conjugation and, to a lesser extent, oxidation. The primary oxidative metabolites

include hydroquinone and catechol, which are also substrates for conjugation. Secondary products

of hydroquinone or catechol metabolism, including benzoquinone and trihydroxybenzene, can also

be formed. Once absorbed, phenol is widely distributed in the body, although the levels in the lung,

liver, and kidney are often reported as being higher than in other tissues (on a per-gram-tissue basis). 

Elimination from the body is rapid, primarily as sulfate and glucuronide conjugates in the urine,

regardless of the route of administration. Phenol does not appear to accumulate significantly in the

body.

3.1. ABSORPTION

Extensive absorption following inhalation exposure has been demonstrated in both human

and laboratory animal studies. Piotrowski (1971) studied lung and skin absorption of phenol in

volunteers (seven male and one female) exposed to air concentrations of 6 to 20 mg/m3

 for 8 hours. 

The study subjects were staff of the research institute in Poland, and all had undergone a previous

medical examination. In the lung absorption studies, the subjects inhaled phenol through a face

mask, eliminating the potential for most dermal absorption. These subjects retained 60–88% of the

inhaled phenol, and the percent retained did not vary with exposure concentration. The absorption

rate leveled off after approximately 3 hours of exposure, indicating that absorption had reached

steady state. In the skin absorption studies, subjects wore underwear and denim coveralls or were

unclothed for different trials of the experiment; in each case they were supplied with fresh air from

outside the chamber for breathing. The absorption coefficient did not appear to vary greatly with

exposure for 6 hours to concentrations in air ranging from about 5 to 25 mg/m3, and clothing did not

appear alter the absorption rate. The mean absorption coefficient was 0.35 m3

/hr, indicating that the

amount of phenol present in 0.35 m3 of air was absorbed through the skin per hour. These data show

that dermal absorption can contribute significantly to the systemic dose of phenol following

exposure to phenol in air. However, the quantitative data from the dermal exposure study are limited

for the development of an RfC because of the short duration of the exposure and the absence of a

direct determination of whether the absorption rate had reached steady state.
1

This study has not been peer-reviewed, but it was well-conducted according to EPA guidelines

for a pharmacokinetics study (with minor deviations).

6

Other studies of workers exposed to phenol provide evidence for significant absorption via

the inhalation route; however, the contribution of dermal absorption from direct contact with liquid

phenol or from phenol in air was not assessed in these studies. Ohtsuji and Ikeda (1972) studied the

urinary free and conjugated phenol levels in Bakelite® factory workers. The total and conjugated

phenol levels tended to increase with increasing air concentration, but free phenol levels were not

affected. This suggests that at the concentrations studied (up to 12.5 mg/m3

), phenol conjugation

was not rate limiting. The investigators did not specifically evaluate the levels of oxidative phenol

metabolites, so no conclusion can be made regarding whether Phase I metabolism was rate limiting. 

On the basis of mass balance analysis, the authors concluded that phenol is efficiently absorbed from

the lung, because the phenol dose (air concentration * air volume/hr) was similar to the total mass

excreted in the urine.

Other occupational studies provide qualitative evidence for lung absorption, reporting

increasing urinary excretion of phenol metabolites with increasing workplace air concentrations. In

many cases, the data are not adequate to estimate the rate and degree of absorption through this

route, and potential contributions of dermal absorption are often inadequately described (Ogata et

al., 1986; ACGIH, 1991).

Absorption through the lung has also been evaluated in laboratory animals report following

inhalation exposure or intratracheal administration. In an unpublished study, Dow Chemical Co.

(1994) studied the kinetics of 14C-phenol in Fischer 344 (F344) rats following inhalation exposure

to 25 ppm (96 mg/m3

) for 6 hours (nose only).1 Radioactivity in the blood was at steady-state levels

at the first measured time point (120 minutes after beginning the 6-hour exposure), indicating rapid

absorption kinetics. Hughes and Hall (1995) evaluated the disposition of phenol following

intratracheal and intravenous (i.v.) administration of 63.5 nmol of 14C-phenol to female Fisher 344

rats. The recovery of radioactivity in tissues and excreta for both routes was approximately 90% of

the administered dose within 72 hours. Because the amount of radioactivity recovered was nearly

equal for the intratracheal and the i.v. dose routes (and fecal excretion is minimal), the authors

concluded that absorption was near 100%.

Hogg et al. (1981) administered 14C-phenol intratracheally in isolated perfused rat lungs

from MRC hooded rats. At the end of the experiment (perfusions were approximately 85 minutes),
7

approximately 92% of the administered radioactivity was in the perfusate, 6% was recovered in the

lung lavage, and approximately 3% was associated with lung tissue. The high recovery in the

perfusate indicated that phenol is nearly completely absorbed across the airways.

Human evidence for oral absorption indicates rapid and complete absorption. In a study of

three human volunteers, Capel et al. (1972) found that 85–98% of a 0.01 mg/kg oral dose of 14Cphenol was excreted in the urine—primarily as phenylsulfate and phenylglucuronide—within 24

hours. In addition, case reports of oral poisoning provide qualitative evidence for gastrointestinal

absorption of phenol, but the ingested and absorbed dose were not estimated in these reports, and in

some cases both oral and dermal exposure were involved (Tanaka et al., 1998). 

Numerous laboratory animal studies have found that orally administered phenol is readily

absorbed. In most cases, absorption rates were not calculated, but the rapid recovery of high

percentages of administered doses in the urine—with only minimal recovery in

feces—demonstrated nearly complete absorption. In various studies in the rat (Kao et al., 1979;

Edwards et al., 1986; Kenyon et al., 1995), the percentage of the administered dose recovered in

urine ranged from 65 to 96.5% over a wide range of doses. Varying degrees of absorption have

been observed in a variety of other species. In a study of 18 animal species orally administered

single doses ranging from 20 to 50 mg/kg 14C-phenol, Capel et al. (1972) found that the percentage

of radiolabel recovered in the urine within 24 hours varied from 31% of the administered dose in

squirrel monkeys to 95% in Wistar rats. It is not clear, however, if these differences represent

differences in the degree of absorption or in the rate of elimination. 

Hughes and Hall (1995) found that in female F344 rats administered 63.5 nmol of 14Cphenol by oral gavage, total recovery of radioactivity (in tissues and excreta) was approximately

90% of the administered dose within 72 hours. Because fecal excretion is approximately 1–3%,

and the recovered dose was nearly equal for the oral and the i.v. dose routes, the authors concluded

that the absorption was near 100%. The difference between the 90% recovery and 100% total

absorption was attributed to losses that were consistent across doses.

Rapid absorption of orally administered phenol has been observed in a number of studies. 

Dow Chemical Co. (1994) studied the kinetics of 14C-phenol in F344 rats following oral dosing by

gavage or in drinking water. Total recovery of the administered radioactivity in the urine, feces,

tissues and carcass, and exhaled air was approximately 95%, regardless of the dosing protocol. In

the high-dose gavage animals (150 mg/kg), peak levels of radioactivity in blood were detected 15
8

minutes post-administration, indicating rapid uptake kinetics. Humphrey et al. (1980) found that

free phenol levels in the plasma of rats given an oral dose of 300 mg/kg radiolabeled phenol

reached a maximum of 26 μg/mL at the first measured time, about 10 minutes after dosing, and

declined rapidly to background by 60 minutes. They observed similar results in beagle dogs given

a 40 mg/kg dose, with rapid peak levels of 7.8 μg/mL and complete removal of free phenol by 1

hour.

More quantitative kinetics data are available from in situ perfusion studies. Humphrey et al.

(1980) administered 14C- phenol (1 mg/mL) to the gut lumen of rats by means of a duodenal

cannula. The remaining radioactivity was measured at 3-minute intervals over 30 minutes in

perfusate collected by an ileal cannula. The results from the intestinal perfusion studies indicated

that removal of 14C-phenol obeys first-order kinetics, with a lumenal T1/2 of 5.5 minutes and a rate

constant for absorption of 0.127 min-1. These authors also measured the plasma concentrations of

phenol in the portal vein and posterior vena cava of dogs following intraduodenal dosing with

either 40 or 160 mg/kg phenol. At either dose, the concentration was already maximal in the portal

vein plasma within 3 minutes after dosing (the first measurement taken) and had decreased to

nondetectable levels within 1 hour at the low dose and to 33% at the high dose. These data show

that in both species phenol is rapidly absorbed from the gut. 

Powell et al. (1974) added 14C-phenol to the mucosal medium of isolated rat gut

preparations and measured the level of radioactivity in the mucosal and serosal medium over 2

hours. They found that 78% of the administered radiolabel had been transferred to the serosal

medium over this period. Kao et al. (1979) administered 14C-phenol (12.5 or 25 mg/kg) to rats

intraduodenally. Recovery of the radioactivity was rapid, with more than 70% recovered in the

urine within 2 hours. 

The dermal route of exposure is an important one. Both absorption of phenol liquid directly

in contact with skin and dermal absorption from exposure to phenol vapor are of concern. 

Significant dermal absorption can result from phenol in air, so that phenol in air results in both

dermal and inhalation exposure (Piotrowski, 1971). On the basis of an analysis of the Piotrowski

(1971) data, ATSDR (1998) concluded that in air concentrations ranging from 5 to 25 mg/m3

, the

amount of phenol absorbed through the skin will be about half of that absorbed through the lungs. 

The conclusion was reached by estimating the amount of phenol absorbed through the lung as the

product of the human ventilation rate of 0.8 m3

/hour and the steady-state lung retention fraction of

0.7 reported by Piotrowski (1971). The resulting lung absorption coefficient of 0.6 m3/hr is nearly

twice the skin absorption coefficient of 0.35 m3/hr. This analysis is limited, however, because it is
9

not clear that the exposure duration was long enough for steady state to be reached in the dermal

absorption studies. Absorption via the dermal route may be lower at steady state due to the

potential for a back-pressure from phenol levels in blood.

A number of case reports and in vitro studies have led to quantitative estimates of phenol

absorption through the skin. Baranowska-Dutkiewicza (1981) applied a reservoir of 2.5, 5, or 10

g/L phenol solution on a small area of the forearm of 12 male volunteers. The absorption rate was

dependent on the concentration and ranged from 0.08 mg/cm2

/hr at the low concentration to 0.301

mg/cm2/hr at the high concentration. At the low concentration, the total amount of phenol

absorbed—but not the absorption rate—increased with increased time; approximately 13% of the

applied dose was absorbed over a 30-minute period. In an in vitro study, 20% of applied doses

ranging from 1.3 to 2.7 μg/cm2 were absorbed from unoccluded human skin patches within 72

hours; addition of a Teflon cap resulted in 47% absorption over this same period (Hotchkiss et al.,

1992).

Bentur et al. (1998) reported an accidental dermal poisoning case in which a solution of

90% phenol was spilled on the left foot (3% of body surface). The exposure site remained

occluded, and no attempt at decontamination was made until the onset of symptoms, which began

within 4.5 hours. Following admission to the hospital shortly afterwards, peak serum phenol levels

of 21.6 μg/mL were measured. This study is presented here for completeness, but quantitative

exposure data from studies at lower phenol concentrations are more relevant to environmental

exposures.

The ability of phenol to be absorbed through the skin has also been evaluated in laboratory

animals. Hughes and Hall (1995) administered 63.5 nmol of labeled phenol to an occluded dermal

patch (2.5 cm2

) of female F344 rats. Maximal recovery of the radioactivity was approximately

70%. The site of dermal application was washed 72 hours post-treatment and yielded 14% of the

recovered dose; 1.6% of the recovered dose was present in the skin at this site. Thus,

approximately 15% of the dose was not absorbed within 72 hours. In an in vitro study, Hotchkiss et

al. (1992) found that phenol absorption by rat skin is similar to that of human skin: approximately

20–50% in 72 hours, depending on the conditions.

Taken together, the human and laboratory animal data demonstrate that phenol is readily

absorbed following exposure by all dose routes. The recovery of greater than 90% of the

administered phenol dose as urinary metabolites provides direct evidence that the administered dose

was nearly completely absorbed. The route of administration appears to play a limited role, with

skin absorption reported as less extensive than absorption from the lung or gut. In most studies,
10

absorption rate constants have not been calculated; however, the measurement of peak blood phenol

concentrations within minutes of dosing indicates that absorption is rapid

3.2. DISTRIBUTION

 Studies in humans and laboratory animals indicate that phenol is widely distributed

throughout the body regardless of exposure route. Because phenol is rapidly excreted, studies on

tissue distribution typically evaluate only a small fraction of the absorbed dose.

Several fatal poisoning case studies evaluated phenol concentrations in multiple tissues

(Tanaka et al., 1998). Generally, phenol is widely distributed. Higher tissue concentrations relative

to blood have been reported for some organs, particularly for the liver and kidneys, although this

finding has not been reported consistently across all studies.

Morrison et al. (1991) reported on the kinetics of phenol injected intramuscularly in a motor

point block procedure in pediatric patients. Administered doses ranged from 6.7 to 70 mg/kg, and

the blood phenol concentration was measured at 5, 15, 30, 60, and 120 minutes after the last

injection. Phenol reached peak levels 5 to 15 minutes after administration and rapidly declined to 3

to 34% of peak levels within 120 minutes. Peak phenol concentration (μg/mL) in blood as a

function of administered dose (x, in mg/kg) was determined (y = 0.483x - 3.244; r = 0.873). 

Pretreatment levels of blood phenol ranged from 0.3 to 0.8 μg/mL and post-treatment levels ranged

from 2.5 to 36 μg/mL. 

The laboratory animal data provide additional evidence for elevated tissue concentrations in

the lung, liver, and kidney, although the magnitude of the tissue differences varies from study to

study. Liao and Oehme (1981) evaluated the tissue distribution of 207 mg/kg 14C-phenol orally

administered to male Sprague-Dawley rats. Total radioactivity in tissues declined rapidly from a

maximum of 28.4% of the administered dose at 0.5 hours to 16.6% at 1 hour and 0.3% at 16 hours. 

Tissue concentrations of radioactivity measured at time points between 0.5 and 16 hours were

significantly greater than in plasma for the liver, spleen, kidney, and adrenal gland; tissue

concentrations in lungs and thyroid were also marginally elevated. The liver had the greatest

amount of radioactivity, accounting for 29–56% of the total radioactivity recovered from tissues at

the various time points. The study authors attributed the high levels in the liver to both an elevated

tissue concentration and the large relative organ size. Because the study measured total

radioactivity without further identification of the radiolabeled compounds, it is not known whether

the observed radioactivity represented phenol or its metabolites. 
11

Dow Chemical Co. (1994) conducted a study of 14C-phenol administered to F344 rats by

oral gavage at 1.5, 15, or 150 mg/kg in drinking water at 5000 ppm or via nose-only inhalation at

25 ppm for 6 hours. Tissue levels of radioactivity were measured in the kidneys, liver, lung,

muscle, skin, spleen, testes, ovaries, and carcass 24 hours after exposure by the various routes. The

only sites with a statistically significant increase in radioactivity levels were the kidney and liver

(levels 5- to 10-fold higher than in other tissues). This finding was consistent across dosing

regimens. 

Hughes and Hall (1995) evaluated the disposition of radiolabeled phenol administered

dermally, by oral gavage, intravenously, or intratracheally to female F344 rats. When the rats were

sacrificed 72 hours after administration by any of these four routes, tissue concentrations

represented only 1–5% of the recovered dose. No tissue appeared to have higher concentrations of

radiolabel following oral dosing, but the lung concentrations were markedly higher following

intratracheal administration. There was no substantive difference across tissues following dermal

dosing, although untreated skin had a slightly higher level. Marginal elevations in the liver and

kidneys were observed following i.v. dosing. The authors concluded that phenol is distributed

widely in tissues, with some accumulation in the large organs (lung, liver, and kidney, based on

within-route comparisons to the levels in blood). 

Powell et al. (1974) treated juvenile rats (50 g) with less than 1 mg/kg 14C-phenol orally or

intraperitoneally. Whole-body radiograms indicated that the liver was not a site for accumulation

of the phenol; rather, it was widely distributed. It is not clear whether the difference between the

findings of this study and others is due to the differences in the sensitivity of the analysis or to

differences in dose levels. Thus, the data from animals studies at doses ranging from 1.5 to 207

mg/kg, which included several doses higher and lower than the chronic No Observed Adverse

Effect Level (NOAEL) of 60 mg/kg (NTP 1983a; Argus Research Laboratories, 1997) (see Chapter

5), showed that phenol is rapidly distributed to a wide range of tissues.

No direct studies of the placental transfer of phenol were identified. However, Ghantous

and Danielsson (1986) evaluated the placental transfer of benzene, of which phenol is a primary

metabolite. B6 mice were exposed for 10 minutes to benzene (at a target concentration of 2000 

ppm) in air on gestation day (GD) 11, 14, or 17. The investigators conducted whole-body

radiography analysis and determined of tissue concentrations. Radioactivity was distributed to the

fetuses, but it was not specifically identified as phenol. The concentration of volatile and

nonvolatile radioactivity in the fetuses was, however, lower than that in maternal tissues.
12

The human and laboratory animal data indicate that phenol is widely distributed in the body. 

Although the human data are inconclusive, the laboratory animal data consistently indicate that

highly perfused organs such as the liver, kidney, and lung have higher tissue concentrations in

comparison to the blood concentration. 

3.3. METABOLISM

Metabolic pathways for phenol are shown in Figure 1. Phenol is directly conjugated with

sulfate or glucuronic acid. Phenol that is not directly conjugated can also be a substrate for

oxidation reactions. The cytochrome P450 2E1 isozyme (CYP2E1) catalyzes the addition of one

oxygen atom to a variety of low-molecular-weight substrates such as benzene and chloroform, and

it is thought to be the primary P450 isozyme for phenol oxidation, although a minor role by other

cytochrome P450 enzymes cannot be discounted. The oxidation products of phenol generated by

CYP2E1 activity appear to be primarily hydroquinone and catechols, which can themselves

undergo further oxidation by CYP2E1 to trihydroxybenzene or by peroxidation to benzoquinone. 

Alternatively, the hydroquinone or catechol metabolites can undergo conjugation reactions. In

addition to P450-mediated oxidation, some studies have suggested that peroxidative metabolism of

phenol can also take place, producing biphenols and diphenoquinones.

Direct sulfate and glucuronic acid conjugations are detoxifying mechanisms that represent

the bulk of phenol metabolism, as evidenced by the metabolic profiles observed in both humans
13

OH O O HO

OH

OHOH

OH

OH

OH

O

O

OH

Biphenol

Biphenol polymers

Diphenoquinone*

peroxidase

PST

Glucuronide conjugate Sulfate conjugate

CYP2E1 CYP2E1

PST

UGT

Sulfate and glucuronide conjugation

Catechol Hydroquinone

Sulfate and glucuronide conjugates

PST

UGT

Benzoquinone

Glutathione c

Trihydroxybenzene

PST

UGT

Sulfate and glucuronideconjugates

*

*

peroxidase

*

Phenol

UGT

CYP2E1 CYP2E1

PST=

UGT=

Figure 1. Metabolism of Phenol.

UDP-DEP

GLUCURO

PHENOL 

*INDICATES METABOLIT

peroxidase

OH
14

and laboratory animals. In humans and in most other species tested, sulfation predominates at

the lower doses. Capel et al. (1972) studied the urinary metabolites following oral

administration of 0.01 mg/kg 14C-phenol to three male volunteers. In these men, 85–98% of the

dose was excreted in 24 hours—69–90% as phenyl sulfate, 4–23% as phenyl glucuronide, and

trace amounts as hydroquinone conjugates. This high degree of conjugation indicates that, at

low doses, ingested phenol is nearly completely conjugated.

Laboratory animal studies have clearly shown that as the dose increases, the role of

glucuronidation increases, until at sufficiently high doses it becomes the predominant reaction. 

The formation of oxidative metabolites (primarily by CYP2E1) also increases with increasing

dose. These dose-dependent changes have been best characterized in rat studies, which show

that at low doses sulfation predominates, with the glucuronidation beginning to predominate at

approximately 133 μmol/kg (12.5 mg/kg) (Kao et al., 1979; Powell et al., 1974; Hogg et al.,

1981; Koster et al., 1981; Edwards et al., 1986; Meerman et al., 1987; Dow Chemical Co., 1994),

a dose that is below the chronic NOAEL of 60 mg/kg for phenol toxicity (NTP 1983a; Argus

Research Laboratories, 1997), as discussed later in detail. There is considerable interspecies

variation, however, in the contribution of the sulfation and glucuronidation pathways (Capel et

al., 1972; Mehta et al., 1978). 

The formation of oxidative metabolites increases at high doses. Dow Chemical Co.

(1994) reported that hydroquinone conjugates represented 3, 7.8, and 17.0% of the eluted

radioactivity in an analysis of urinary metabolites following single oral doses of 1.5, 15, and 150

mg/kg 14C-phenol, respectively. Legathe et al. (1994) administered an intraperitoneal (i.p.) dose

of 75 mg/kg to B6 mice and reported urinary metabolites as 34.5% phenyl sulfate, 28.5% phenyl

glucuronide, and 32.4% hydroquinone glucuronide, indicating substantial contribution of

oxidative metabolism at this high dose. 

The formation of oxidative metabolites is thought to result primarily from reactions

catalyzed by CYP2E1. Koop et al. (1989), using hepatic microsomes prepared from male New

Zealand white rabbits, showed that CYP2E1 was the most active of six P450 isoforms tested. 

Treatment of the lysates with an antibody to CYP2E1 inhibited hydroquinone formation by 68% and

89% in acetone-induced and uninduced microsomes, respectively. Snyder et al. (1993) studied

phenol metabolism in vitro in rat hepatic microsomal preparations. Addition of phenol to the

CYP2E1 microsome preparation yielded hydroquinone and, to a lesser degree, catechol metabolites. 

Incubation of 14C-phenol and 3

H-glutathione in the CYP2E1 microsome preparation yielded an

additional metabolite that cochromatographed with the compound formed from the reaction of
15

benzoquinone with glutathione. The formation of the glutathione adduct was not dependent on

addition of glutathione-S-transferase. Lunte and Kissinger (1983) also reported the formation of

glutathione conjugates in microsomal (prepared from liver of male Swiss mice) metabolism of

phenol to hydroquinone. In addition to benzoquinone, hydroquinone and catechol can also be

oxidized to trihydroxybenzene (Sawahata and Neal, 1983).

Other in vitro studies using hepatic microsomes from rats treated with various P450 inducers

or inhibitors have also provided evidence for the importance of CYP2E1 in phenol metabolism

(Sawahata and Neal, 1983; Gilmour et al., 1986; Chapman et al., 1994; Kenyon et al., 1998). 

CYP2E1 appears to predominate phenol oxidation. Contributions by other P450 enzymes cannot be

excluded, however, because only 68% of the induced hydroquinone formation was blocked by antiCYP2E1 antibody, and several inducers of other P450 enzymes (such as phenobarbital and arochlor)

enhanced phenol metabolism in these studies. 

An alternative oxidative pathway involving peroxidation has been described for phenol. 

Several investigators have used in vitro cell preparations with high peroxidase activity, such as

peritoneal macrophages or neutrophil preparations (Eastmond et al., 1986; Post et al., 1986; 

Eastmond et al., 1987; Kalf et al., 1990), purified peroxidase enzymes (Smart and Zannoni, 1984;

Subrahmanyam and O’Brien, 1985), or cell lines that have high myeloperoxidase activity

(Kolachana et al., 1993), to show that phenol can be metabolized in these reactions. Metabolites

resulting from these reactions include 4,4'-biphenol and diphenoquinone. Although the peroxidation

of phenol has been demonstrated in vitro, no direct in vivo evidence for these peroxidative reactions

was identified.

The shift from sulfation to glucuronidation at increasing doses has been postulated to result

from depletion of sulfate pools (Kim et al., 1995). Alternatively, it has been suggested that the

difference between the Km values for sulfate and glucuronide conjugation drives the conjugation

shift (Weitering et al., 1979). The effects of differing metabolizing enzyme activity across the zones

of the liver has also been suggested as an explanation for the metabolic profiles of phenol

(Medinsky et al., 1995). The functional units of the liver include lobules with blood supply

provided from the perimeter (periportal region) of the lobule though the portal vein and the hepatic

artery. The blood flows from the periphery of the lobule toward the terminal hepatic vein (also

called the central vein) at the center of the lobule through a series of differing metabolic regions or

zones. Both sulfotransferases and glucuronosyltransferases are present in periportal zone 1, with the

sulfotransferases predominating. Glucuronosyltransferases are present in zone 2, while both
16

glucuronosyltransferases and monooxygenases such as CYP2E1 are present in pericentral zone 3. 

According to the model, phenol entering the hepatic circulation would be metabolized first

in the periportal region, where sulfation predominates. Because the blood flows from the periportal

region to the pericentral region and then to the central vein and general circulation, little

unconjugated phenol is available for glucuronide conjugation or oxidation when it reaches the

pericentral regions of the liver. This model is consistent with the shift from sulfation to

glucuronidation at increasing doses. As the dose increases, more of the phenol reaches the

pericentral region unconjugated, and thus is available for glucuronidation. The model also explains

the increase in oxidative metabolites at high doses that exceed the conjugating capacity of zones 1

and 2 (Kenyon et al., 1995).

The model is also consistent with the observation that oral dosage with benzene results in

greater production of hydroquinone than that seen after oral dosage with phenol, even though

benzene is metabolized to hydroquinone via phenol. Benzene enters the liver in the periportal

region, is oxidized to phenol and other metabolites in the pericentral region, and then leaves the

liver via the hepatic vein. Because benzene must be oxidized before it is conjugated, more

unconjugated phenol would be released into the blood following benzene exposure than following

phenol exposure (Medinsky et al., 1995).

Direct evidence for this model was presented by Ballinger et al. (1995), who studied phenol

and metabolite transport kinetics in isolated perfused liver from rats, and by Hoffmann et al. (1999),

who conducted similar experiments in mice. The effects of enzyme distribution in the zones of the

liver were studied by contrasting phenol metabolite profiles resulting from antero- and retrograde

perfusions. It is noteworthy that the importance of the location of enzyme activities within the liver

would only be significant at oral phenol doses that were not conjugated at the portal of entry, and

thus were available for transport to the liver via the hepatic portal vein.

There is some evidence that the capacity for phenol conjugation varies with the portal of

entry. Cassidy and Houston (1984) conducted an elegant series of experiments in which they

injected rats intra-arterially, intravenously, or intraduodenally with phenol and then measured the

systemic availability of phenol. This allowed them to evaluate the first-pass metabolism by different

organ systems at doses ranging from 0.4 to 15 mg/kg. They were able to use this approach to

determine metabolism by the liver and gut. However, results on metabolism by the respiratory tract

from this study should be treated with caution, because environmental exposure results in exposure

of the epithelial respiratory tract (i.e., the portion exposed to the outside), whereas this study
17

involved exposure of the endothelial respiratory tract (i.e., the portion exposed to the inside of the

body). Thus, any differences between the metabolic capacity of the endothelial and epithelial cells

would not be taken into account by the study design.

In this series of experiments, phenol that was systemically available had not been conjugated

or metabolized. The doses at which phenol became systemically available thus reflected the doses at

which the relevant metabolic enzyme systems became saturated. Metabolism became nearly

saturated between 4.5 and 15 mg/kg for the endothelial lung, and between 0.4 and 1.5 mg/kg in the

liver, and it was not saturated at the high dose for the gut. The affinity of metabolic pathways also

varied among the organ systems. The liver and gut, which removed 88% and 86% (respectively) of

phenol at the 0.4 mg/kg dose, demonstrated high affinity in comparism with the endothelial lung,

which removed 58% of the phenol at this same dose. Taken together, these data indicate that the gut

is a high-affinity and high-capacity site of metabolism, and the liver has high affinity but its capacity

is readily exceeded.

The data also suggest that the lung provides substantial metabolizing capacity but has lower

affinity than the gut and liver. Clear conclusions regarding the metabolic capacity of the lung

following exposure by the inhalation route are not possible because of the potential for differences

between the metabolic capacity of the epithelial and endothelial cells of the lung. However, one

would expect the potential for metabolism of inhaled phenol to be similar to that seen in this study,

because systemically absorbed phenol must pass through the endothelial cell layer. 

The area under the blood concentration curve (AUC) for 14C-phenol was route- and dosedependent, reflecting the effects of portal-of-entry metabolism. In contrast, the AUC for phenol

metabolites did not differ by dosing route, indicating that phenol is extensively metabolized and the

effect of portal-of-entry metabolism is to reduce the amount of parent compound available for

metabolism by other organ systems. 

Studies using isolated perfused rat liver were also conducted and correlated well with the in

vivo data. The percent of phenol removal from blood by first-pass metabolism declined from 73% at

a blood concentration of 2.8 μg/mL to 26% at 26 μg/mL, indicating extensive saturation at the

higher dose level.

Dow Chemical Co. (1994) studied the differential metabolism kinetics of phenol by differing

exposure routes. This study evaluated the kinetics of 14C-phenol in F344 rats following dosing

regimens that included single or eight daily oral gavage doses of 1.5, 15, or 150 mg/kg 5000 ppm in
18

drinking water for 1 or 8 days or 25 ppm via inhalation for 6 hours (nose-only) for 1 or 8 days. The

authors estimated the doses resulting from the drinking water and inhalation exposures. For

drinking water administration (males only), doses were estimated by measurement of daily water

consumption. The administered dose was 291 mg/kg for the single-day protocol and 405 mg/kg for

the last day of the 8-day treatment, based on the water intake and the weight of each animal; thus,

the drinking water doses were higher than the oral gavage doses.

The doses used in the drinking water study and the high dose in the gavage study were all

higher than the chronic NOAEL of 60 mg/kg (NTP 1983a; Argus Research Laboratories, 1997), as

discussed in detail in Section 5. In contrast, the inhalation doses were estimated as 11.5 and 17.8

mg/kg for males and females, respectively, following a single exposure period, and the dose was

21.4 mg/kg (males only) on the last day of the 8-day exposure protocol. Thus, the inhalation route

more closely resembled the middle gavage dose level, and the absorbed dose via inhalation was

lower than the chronic NOAEL (NTP 1983a; Argus Research Laboratories, 1997).

Metabolic profiles revealed ratios of 0.61 for glucuronide/sulfate conjugates in urine at the

two lower gavage doses and were similar following inhalation (0.24–0.39). The ratio at the high

gavage dose was 1.16, and it was similar following drinking water exposure (1.43 and 1.87 for the

single and 8 day exposures). The observed formation of oxidative products, as shown by urine

levels of hydroquinone glucuronide, was also dependent on total dose. The formation of oxidative

metabolites following inhalation paralleled the low-dose gavage data, whereas the drinking water

levels paralleled the high-dose gavage levels. The pattern of phenol metabolism correlated with the

magnitude of the absorbed dose and did not appear to be dependent on the route of administration. 

Metabolism of phenol appears extensive in the lung, liver, and gastrointestinal tract; however

limited data are available for other organs. Metabolism appears to be extensive in the kidney

(Tremaine et al., 1984). No data were identified that addressed portal-of-entry metabolism for the

skin.

One consequence of the portal-of-entry metabolism of phenol is that phenol serum levels are

not necessarily linear with dose or exposure levels. At low doses, almost all of the absorbed phenol

is conjugated and excreted, without entering the bloodstream. At higher doses, free phenol and its

metabolites appear in the blood and increase with dose. This nonlinearity of blood phenol levels

with dose is illustrated by the data of Dow Chemical Co. (1994). Peak phenol blood concentrations

in rats following an oral bolus dose of 150 mg/kg were 2320-fold higher than the peak blood

concentrations following an oral bolus dose of 1.5 mg/kg.
19

The role of peak levels may be significant for induction of at least some aspects of systemic

toxicity. Dow Chemical Co. (1994) sheds some light on the relationship between metabolism and

toxicity. The high-dose gavage group in this study developed a cluster of behaviors that the authors

termed “phenol twitching behavior (PTW)” that included tremors, sudden jerks, hyper-reactivity to

stimuli, and excessive blinking. PTW began almost immediately after dosing and disappeared by 37

minutes post-dosing. Blood phenol levels also peaked almost immediately after dosing, and PTW

was not apparent at blood phenol concentrations below approximately 3 μg/mL. PTW was not

observed at the lower gavage doses or following inhalation exposure; peak blood phenol levels in

these groups were well below 1 μg/mL. 

Interestingly, PTW was also not observed in the drinking water exposure groups, even

though the total dose in these groups was higher than the high gavage dose, and the drinking water

doses had a similar metabolic profile to the high gavage dose. Unfortunately, blood phenol levels

were not sampled in the drinking water groups, so the peak blood phenol level is not available. 

However, given the rapid clearance of phenol from the blood, it is likely that the peak blood level

was much lower in the drinking water group than in the high-dose gavage group. This suggests that

PTW is more closely related to peak phenol blood levels than to a measure of total dose, such as

AUC. Because phenol metabolite levels paralleled those of phenol, these data cannot be used to

distinguish between phenol and its metabolites being the toxic agent. These data do not identify the

appropriate dose metric (e.g., peak concentration vs AUC) for other toxic endpoints.

One indication that the oxidative metabolites are important determinates of toxicity is based

on experiments by Chapman et al. (1994). They studied the dysmorphogenic and embryotoxic

effects of benzene and its metabolites to the whole rat conceptus in vitro. Phenol at 1.6 mM elicited

only minor effects, but inclusion of S9 microsomal fractions greatly increased the potency of phenol,

with significant effects observed at doses as low as 0.01 mM. Metabolite analysis indicated that

hydroquinone and catechol were the primary metabolites. When evaluated singly, hydroquinone,

catechol, and benzoquinone induced similar embryotoxicity, producing 100% lethality at 0.1 mM. 

The addition of phenol and hydroquinone together induced a more-than-additive embryotoxicity,

which the authors suggested as evidence for a peroxidative mechanism for phenol bioactivation,

based on the potential for electron cycling between phenol and hydroquinone.

Intraspecies variability has also been studied. Campbell et al. (1987) isolated human liver

sulfotransferases, the enzymes responsible for the conjugation of phenol with sulfate, and 

analyzed their apparent activities toward p-nitrophenol (as a model compound for simple phenols). 
20

The average phenol sulfotransferase (PST) activity measured in liver samples of 20 patients (13

male, 7 female) was 35.8±10.6 standard error of the mean (SEM) units/mg protein. No correlation

between enzyme activity and patient age or gender was found, although the power to detect any such

correlation was not noted. Seaton et al. (1995) studied the kinetics of phenol sulfation and

hydroquinone conjugation, both of which varied over a range of approximately three-fold in a

sample of liver fractions from 10 humans. Using lysates from a single human liver, saturation of

phenol sulfation was apparent above 800 μM; the observed kinetics were consistent with two

contributing enzymes, PST1 and PST2. The expression of two distinct PST enzymes has also been

demonstrated in human nasal epithelium (Beckmann et al., 1995). 

Kawamoto et al. (1996) studied the effect of various lifestyle factors and of genetic

polymorphisms in five metabolizing enzymes, including aldehyde dehydrogenase (ALDH2), Nacetyl transferase (NAT2), cytochrome P450 1A1 (CYP1A1), CYP2E1, and glutathione-Stransferase mu (GSTM1) on urinary levels of phenol in a cohort of men who were not

occupationally exposured. Step-wise multiple regression analysis was performed to identify

important determinates of urinary phenol levels. On the basis of this analysis, there was no

relationship between polymorphisms (including for CYP2E1) and background urinary phenol levels. 

In the total sample (n = 351), the geometric mean urinary phenol level was 7.64 mg/L and the

geometric standard deviation was 2.9. No data are available, however, on how genetic

polymorphisms affect the levels of metabolites produced from exogenously dosed phenol.

The changes in enzyme activity or expression of genes that encode enzymes important for

phenol metabolism with age have been studied. The status of CYP2E1 in fetuses remains unclear,

with conflicting results reported. Most of the existing studies indicate that this enzyme is expressed

in human adults but not in human fetuses, even when measured using sensitive assays (reviewed in

Hakkola et al., 1998). However, at least two studies (Carpenter et al., 1996; Vieira et al., 1996)

indicate that CYP2E1 is expressed at least to some degree in fetal liver. Vieira et al. found that

CYP2E1 protein could not be detected immunochemically in fetal human liver, and there was only

minimal evidence of CYP2E1 mRNA or CYP2E1 activity in fetal liver microsomes. (The difference

in assay results may be due to differences in sensitivity or to cross-reaction of CYP1A1 activity.) 

However, the authors found, that CYP2E1 protein levels rise rapidly in the first few hours after

birth, with a slow increase in protein levels and in CYP2E1 mRNA levels during childhood. 

Results of animal studies of developmental CYP2E1 regulation are consistent with the

human data in providing uniform evidence of the rapid induction of this gene soon after birth (Song
21

et al., 1986; Umeno et al., 1988; Schenkman et al., 1989; Ueno and Gonzalez, 1990). Thus, overall,

the data show that if CYP2E1 activity exists in human fetuses, levels are much lower than those in

adults. Regardless of fetal CYP2E1 expression, the enzyme is rapidly induced upon birth. For this

reason, children would be expected to be capable of phenol metabolism, although the amount of

CYP2E1 may be less than that present in adults.

Age-dependent changes in phase II conjugation have also been evaluated. In an evaluation

of how PST activity varies with age in rats, Iwasaki et al. (1993) studied ß-naphthol metabolism by

PST in fetal rat liver, in the liver of 2-, 9-, 17-, and 25-day-old neonates, and in adult rats. Activity

was analyzed in the livers of both sexes. The fetal liver had little conjugating ability, but this

activity developed rapidly after birth. However, activity was substantially lower in neonates of all

ages evaluated when compared with adult levels. Heaton and Renwick (1991) administered i.p.

doses of 25 mg/kg 14C-phenol to rats varying in age from 3 to 16 weeks and measured metabolites in

urine collected in 24 hours. The percentage of the administered dose recovered in the urine in 24

hours ranged from 61–90% in males and 63–99% in females, with increasing recovery with age. 

Importantly, the formation of hydroquinone conjugates was greater in the younger animals. 

In males, 38% of the administered dose was recovered as hydroquinone conjugates in the 3-weekold animals; 8.2% of the urinary metabolites was recovered in 16-week-old rats. In females, 17.8%

of the administered dose was recovered as hydroquinone conjugates in 4-week-old rats and 10.5%

was recovered in 15-week-old rats. Taken together, the evidence indicates that both sulfate

conjugation and P450 metabolism are lower early in life and increase as adulthood is reached. 

However, even in the face of limited P450 activity, significant formation of oxidation products can

occur because of limited sulfation capacity. The oxidative products become substrates for

glucuronidation, and this does not appear to be limited in the young. 

Phenol metabolism may also be gender dependent, although the data are less substantial than

those for differences due to age. Iwasaki et al. (1986) reported that PST activity was similar in both

sexes up to 3 weeks of age and was higher in males than in females in 7-week-old rats. Activity in

2-year-old rats of both sexes was similar and fell between the levels for males and females at 7

weeks to 1 year. Kenyon et al. (1995) administered 14C-phenol to B6 mice of both sexes and

observed that, males excreted a greater proportion of HQ-glucuronide than did females at all doses;

the difference was roughly twofold at a dose of 40 μmol/kg. These results are consistent with the

greater degree of hydroquinone conjugates excreted in the urine of male versus female rats reported

by Heaton and Renwick (1991). Sex-based differences in metabolism have also been reported in
22

rats (Meerman et al., 1987), with slightly lower total recovered radioactivity in the urine of females

versus males (i.e., more rapid metabolism in males). However, the magnitude of this difference

(91.2% vs 87.3%) was limited.

Interspecies differences in phenol metabolism have also been evaluated. Seaton et al. (1995)

found that the rates of both phenol sulfation and hydroquinone conjugation in mouse and rat liver

were comparable to those of human liver preparations. Schlosser et al. (1993) reported that mouse

liver microsomes metabolized approximately twice as much phenol as did rat liver microsomes,

although the relative proportions of metabolites were roughly similar. 

Phenol is formed endogenously in the gut by bacterial metabolism of aromatic amino acids in

protein. The amount formed is related to the amount of protein ingested, but the amount in humans

typically varies from 1 to 10 mg/day, corresponding to approximately 0.014 to 0.14 mg/kg-day

(Bone et al., 1976; Lawrie and Renwick, 1987; Renwick et al., 1988).

A physiologically based pharmacokinetic (PBPK) model for the distribution of benzene and

metabolites was developed by Bois et al. (1991). The model was developed to predict phenol and

metabolite distributions to fat, well-perfused tissue, poorly-perfused tissue, bone marrow, liver, lung,

and gut using Monte Carlo simulations of 64 parameters. The model was not validated using

empirical data. The Bois et al. model consistently predicted that phenol administration would

produce higher levels of phenol and hydroquinone in the blood than seen following benzene

administration.

The first phase in the development of a model of the in vitro kinetics of phenol and benzene

biotransformation by liver microsomes was described by Schlosser et al. (1993) and enhanced by

Medinsky et al. (1995). The model described the following reaction sequences: benzene > phenol >

catechol > trihydroxybenzene and phenol > hydroquinone > trihydroxybenzene. All reaction steps

were assumed to be catalyzed by cytochrome P450 2E1, and benzene, phenol, catechol, and

hydroquinone were all assumed to compete through reversible binding for the same reaction site on

cytochrome P450. Parameters were identified that were successful at predicting the concentration

with time of all five chemicals in incubations with rat or mouse liver microsomes (Schlosser et al.,

1993). The observation of a lag time in the production of hydroquinone from benzene—in

comparison to the rate of production of hydroquinone from phenol—supported the assumption that

all of the substrates compete for the same enzyme reaction site. 

Medinsky et al. (1995) extended the data into a conceptual model of the differences between

phenol and benzene metabolism. The goals of the conceptual model included explaining the
23

observed differences between the carcinogenicity and genotoxicity of phenol and benzene and

explaining why urinary hydroquinone levels are higher after benzene dosing than after phenol

dosing. The latter observation would appear to be inconsistent with the prediction of the Bois et al.

(1991) model that blood hydroquinone levels are higher following phenol dosing than following

benzene dosing. As described earlier in this section, differences between benzene and phenol

toxicity were attributed to zonal differences in the distribution of hepatic metabolic enzymes.

In summary, phenol is an endogenous metabolite that undergoes further metabolism

efficiently. At low doses the bulk of the phenol appears to be conjugated with sulfate or glucuronide

at the portal of entry. As the dose increases, the sulfation pathway becomes saturated, and the

relative contribution of glucuronidation and oxidation reactions increases. Saturation of first-pass

metabolism may be important for producing peak levels of phenol that correlate with acute systemic

toxicity. In addition, saturation of conjugation, which leads to increases in oxidative metabolism,

may also be an important determinate of toxicity. The data on intraspecies variability are limited,

but they do not indicate great variation in metabolic capacity in humans. In rodents, males and

younger animals appear to rely more heavily than females and adult animals on oxidative

metabolism, respectively, but the differences are no more than twofold. The metabolism of phenol

in humans and rodents appears to be similar, although some evidence suggests that mice metabolize

phenol more rapidly than do humans or rats.

3.4. EXCRETION

The existing human and laboratory animal studies consistently report that phenol is rapidly

excreted, with little tendency for accumulation. Elimination is primarily in the urine in both humans

and laboratory animals, with only a minor contribution of elimination in the bile. Ohtsuji and Ikeda

(1972) studied the urinary free and conjugated phenol levels in Bakelite® factory workers. Workers

were exposed to phenol vapor by inhalation on a daily basis. The workers were also possibly

exposed by the dermal route, but the contribution of this route to the total exposure was not directly

measured. Analysis of urinary phenol levels at different times during the work shift and across work

shifts indicated that in workers exposed to 7.8 to 9.6 mg/m3

, the urinary levels increased

significantly from the beginning of the work shift to the end of the work shift, but they did not tend

to accumulate across the work shifts. A slight increase in the morning sample on the sixth

consecutive work day was observed, but after two days off, pre-shift samples were no longer

elevated.
24

Rapid clearance from the blood in humans has also been observed. Bentur et al. (1998)

presented a case report from a dermal poisoning in which a solution of 90% phenol was spilled on

the left foot (3% of body surface). Even at this high exposure level, clearance from the blood was

rapid, with blood levels decreasing from 21.6 to 2.8 μg/mL in the first 12 hours. The authors

estimated that the half-life elimination was 13.86 hours, but they did not include the initial rapid

decline in serum concentration that was apparent over the first 12 hours post-admission. Both

absorption and clearance would be expected to be more rapid at lower exposure levels, as high

exposure levels can lead to protein denaturation and saturation kinetics. 

Laboratory animal studies have consistently found that phenol is rapidly excreted. Clearance

of phenol from the blood is rapid. Rats given an oral dose of 300 mg/kg, a level comparable to the

rodent median lethal dose (LD50)of 340 mg/kg (Deichmann and Witherup, 1944), had maximum

blood concentrations of 26 μg/mL at the first measured time point (about 10 minutes), and blood

levels declined rapidly to background by 60 minutes (Humphrey et al., 1980). Similar results were

observed by the same authors in dogs given a 40 mg/kg dose, with rapid peak levels (7.8 μg/mL)

and complete removal of free phenol by 1 hour. A half-life of 12 minutes in blood was reported for

rats administered 150 mg/kg by gavage (Dow Chemical Co., 1994). Legathe et al. (1994) reported

biphasic elimination kinetics from the blood, with a terminal half-life of 22 minutes. Similarly,

Cassidy and Houston (1984) reported biphasic kinetics with a half-life of approximately 5 minutes

following intra-arterial administration. 

The elimination kinetics in multiple tissues was studied by Liao and Oehme (1981). Total

radioactivity in tissues was maximal within 30 minutes of dosing, representing 28.4% of the

administered dose. Tissue levels accounted for 16.6% of the administered dose at 2 hours and 0.3%

at 16 hours. Although maximum levels varied considerably across tissues, the rate of elimination

did not appear to differ with tissue type. Numerous laboratory animal studies indicate that urinary

elimination of sulfate and glucuronide conjugates accounts for most of the excretion, ranging from

70 to 90% of the administered dose within 24 hours, whereas excretion in feces represents only a

small fraction of the administered dose, approximately 1–3% (Edwards et al., 1986; Meerman et al.,

1987; Dow Chemical Co., 1994; Hughes and Hall, 1995).

4. HAZARD IDENTIFICATION
25

4.1. STUDIES IN HUMANS—EPIDEMIOLOGY, CASE REPORTS, CLINICAL

 CONTROLS

The epidemiology data on phenol are limited. Epidemiology studies have typically included

confounding exposures and have not adequately adjust for smoking. Kauppinen et al. (1986)

reported a significant increase in respiratory cancer in phenol-exposed workers, but this observation

appears to be due to confounding exposures, because there was no dose-response, and the effect

decreased after accounting for latency. No effect on cancer mortality was observed in workers

exposed to phenol in the rubber industry (Wilcosky et al., 1984) or in workers exposed to

formaldehyde and phenol (Dosemeci et al., 1991). An occupational study (Shamy et al., 1994) and

case studies (e.g., Merliss, 1972) have reported liver effects following exposure to phenol. Immune

effects were also been reported in an occupational study of workers exposed to phenol as part of a

mixture of solvents (Baj et al., 1994). Studies of populations whose drinking water was

contaminated with phenol found elevated incidences of diarrhea, nausea, mouth sores, and dark

urine (Jarvis et al., 1985; Baker et al., 1978).

4.1.1. Oral

Estimated lethal oral doses of phenol in adults vary widely, from 1 g (14 mg/kg, assuming an

adult body weight of 70 kg) to as much as 65 g (930 mg/kg) (Deichmann and Klepinger, 1981). In

another report (Bruce et al., 1987), the minimum lethal oral dose in adults was estimated as 140

mg/kg.

Jarvis et al. (1985) reported on illness associated with consuming water contaminated with

phenol from a spill into the river that served as the drinking water source. A retrospective mail

survey was sent to a total of 594 English households: 250 that were highly exposed, 94 that were

exposed at low concentrations (from a reservoir that diluted the contaminated river water), and 250

that were unexposed (selected from a telephone book). On the basis of data from the water

authority, the estimated phenol concentrations in drinking water in the low-exposure area (0.05

mol/L, equivalent to 4.7 μg/L) was roughly half that in the high-exposure area (0.11 μmol/L,

equivalent to 10 μg/L) for the first 24 hours. The next day, the phenol concentration for both groups

was 0.05 μmol/L, and the concentration was < 0.01 μmol/L (< 0.9 μg/L) by the third day after the

contamination incident. 

Chlorination of the water resulted in production of chlorophenols. The chlorophenol

concentration followed a similar pattern, but the chlorophenol concentrations, which ranged from
26

0.43 to 0.2 μmol/L at the first measurement (84.7–39.4 μg/L, assuming all chlorophenols were in

the form of trichlorophenols), were higher than those for phenol. There were no data on phenol

concentrations in the unexposed area, but an unspecified local press report implicated a possibility

of phenol contamination. Due to the similarity of the two exposed areas in the measured

concentrations of phenol and chlorophenol, these two areas were combined in the data analysis. 

The percentage of responding households was similar in all of the groups and ranged

from 69 to 77%, resulting in 172 households (448 people) in the unexposed area and 254

households (754 people) in the exposed area being evaluated. The two groups had similar

distributions in sex, age, and usual water usage. Compared with the unexposed individuals,

those in the exposed area had significantly higher incidences of gastrointestinal illnesses, such

as diarrhea, nausea, vomiting, and abdominal pain. Other symptoms, such as headache, rash,

and malaise, were also observed at a significantly elevated incidence in the exposed group. The

day of onset of symptoms corresponded with the period of elevated phenol concentrations in the

contaminated drinking water. The associations were stronger among those who reported that

they drank the water than among those in the exposed area who reported not drinking the water. 

(Others may have consumed the water in cooking.) In another analysis, gastrointestinal

symptoms did not significantly correlate with whether the water tasted bad. However, because

of the confounding exposure of chlorophenol in the water, the usefulness of the result for risk

assessment is limited.

Baker et al. (1978) reported on phenol poisoning in humans due to an accidental

contamination of their drinking water on July 16, 1974. A train derailment resulted in a

spillage of 37,900 L of pure phenol onto the ground, and the spillage caused contamination of

drinking water in wells of nearby houses. Two wells near the spill were tested initially on

July 23 and were found to have phenol concentrations of 0.21 and 3.2 mg/L. Further testing

in late July and August of the six wells nearest the spill found peak concentrations between 15

and 126 mg/L. Within approximately 2 months after the spill, “most families” began to obtain

water from other sources (from neighbors or bottled water). Phenol concentrations in well

water as high as 1130 mg/L were reported over the next 6 months, with the higher levels

observed after flushing of the spill site (unspecified number of wells tested). The authors

investigated the health effects in three groups of people. Group 1 (n = 39) consisted of all

families living 120–310 m from the spill site and having at least one water test greater than

0.1 mg phenol/L (at least once between July and February). Group 2 (neighborhood control)

(n = 61) consisted of all families adjacent to Group 1 (210–670 meters from the spill) whose

wells had phenol concentrations of between 0.1 and 0.0001 mg/L. Group 3 (distant control)
27

(n = 58) lived at least 1.9 km from the spill and had no phenol in its wells. 

Group 1 reported significantly more diarrhea, mouth sores, burning mouth, and dark urine

than the combined control groups. About 44% of the individuals in Group 1 reported at least two

of these responses in the 7 months following the accident and were considered “affected

individuals”; only 8% and 3% of Group 2 or 3 subjects, respectively, had the same responses. 

Responses in Group 1 were primarily restricted to the first 2 months of exposure, before the use of

bottled water began. Responses in the other two groups tended to occur throughout the 8-month

period (July through February). Other than the four reported symptoms, no abnormal observations

in physical examinations and serum biochemical evaluations were evident in Group 1 as compared

with controls when monitoring was done in February. 

On the basis of water testing data and water preference histories, the authors estimated that

the daily oral dose of phenol for the 17 affected individuals in Group 1 was between 10 and 240

mg. However, this range may overestimate the amount of phenol ingested because phenol’s

unpleasant odor might have discouraged ingestion of water with concentrations above 0.1 mg/L. 

In contrast, this range does not include phenol that may have been absorbed during skin contact

with contaminated water. It was also not clear whether the subjects continued to shower with the

contaminated water after switching their drinking water source. Based on a default adult body

weight of 70 kg, this daily oral dose corresponds to 0.14 to 3.4 mg phenol/kg-day. Thus, there is a

considerable range in the estimated phenol dose associated with symptoms. In addition, because

“most” (but apparently not all) families switched to other water sources within the first 2 months

of exposure, the exposure duration for the affected individuals is not known. Therefore, it is

difficult to use these data for quantitative analysis, although they might be useful for placing

bounding estimates on the risk values for systemic effects estimated from laboratory animal

studies.

4.1.2. Inhalation

Kauppinen et al. (1986) reported a case-control study on respiratory cancers and chemical

exposures in the wood industry. A cohort of 3805 Finnish men who worked in the particle board,

plywood, sawmill, or formaldehyde glue industries for at least 1 year between 1944 and 1965 was

followed until 1981. From the cohort, 60 cases of respiratory malignant tumors were identified. 

The tissue locations of these tumors included tongue (1), pharynx (1), larynx or epiglottis (4), and

lung or trachea (54). No cases with tumor in the mouth, nose, or sinuses were identified. Among

the 60 cases, 2 were rejected due to a false preliminary diagnosis of cancer and 1 was rejected as
28

chronic lymphocytic leukemia. The final size of the group of cases was thus 57. The control 

group contained three subjects for each case, selected from the cohort and matched by birth year, for

a total size of 171.

The job exposure was estimated from the industrial hygiene data of the plant, general

hygiene data on exposures, and information on ventilation, work procedures, and other relevant

factors at the plants. However, the authors, gave no information on direct phenol measurements;

thus, the quality of the estimated exposure levels could not be evaluated. The work histories of the

subjects were assessed primarily from plant registers and supplemented with personal interviews. 

Individual phenol exposures were determined qualitatively as “yes” or “no” and as a function of

exposure time. 

Smoking histories were determined by a mail survey that resulted in smoking information on

39 of 57 cases (68%) and on 130 of 171 controls (76%). Because there were few nonsmokers and

information on the amount smoked was not as complete as on years of smoking, the subjects were

compared only as light or heavy smokers, based on whether their years of smoking exceeded 35. 

Phenol exposure resulted in a statistically significant odds ratio (OR) of 3.98 or 4.94 for

respiratory tumors with or without the adjustment for smoking years, respectively. When the

duration of phenol exposure was considered, both exposures < 5 years and > 5 years resulted in a

statistically significant OR < of 5.86 or 4.03, respectively (i.e., no duration response). When a

provision for a 10-year latency was introduced (excluding exposure during the 10 years immediately

preceding the diagnosis of cases), phenol exposure resulted in a nonsignificant OR of 2.86 adjusted

for smoking years but a significant or of 3.98 without smoking adjustment. Of the 39 cases for

which smoking information was available, 12 had been exposed to phenol (9 to phenol in wood

dust), and 7 had been exposed to phenol with a 10-year latency (4 to phenol in wood dust). Because

the OR did not increase with duration of phenol exposure and the provision for the 10-year latency

period resulted in lower values of ORs, a confounding factor may have been responsible for the

observed statistically significant ORs.

One of the confounding factors could have been concurrent exposure to multiple pesticides,

which was in the same study shown to increase the OR for respiratory tumors. An exclusion of

workers exposed to both phenol and pesticides resulted in a change of the OR from a significant 4.9

to a nonsignificant 2.6. Thus, a confounding effect due to exposures to pesticides was very possible. 

Considering the location of the tumors, formaldehyde exposure was also a likely confounder. 

Generally similar results were observed in this study for workers exposed to phenol in wood
29

dust. Exposure to phenol in wood dust resulted in a statistically significant OR with or without

adjustment for smoking. For the workers exposed to phenol but not wood dust, provision for a

latency period eliminated the observed statistically significant OR. Among the workers exposed to

phenol in wood dust, however, the OR did increase with exposure duration and was statistically

significant in those exposed > 5 years (OR of 4.77) but not in those exposed for < 5 years (OR of

3.84). On the basis of these results, the phenol-exposed workers had an elevated risk of respiratory

cancer, but phenol itself did not appear to be the causative agent; rather, it appears that there was a

confounding exposure. 

Wilcosky et al. (1984) reported a case-control study of cancer mortality and solvent

exposures in the rubber industry. From a cohort of 6678 active and retired male rubber workers of a

large plant in Ohio, 183 decedents from stomach cancer, prostate cancer, lymphosarcoma and

reticulum cell sarcoma, lymphatic leukemia, and respiratory cancer were selected as cases. As a

control, 20% of an age-stratified random sample of the cohort (calculated as 1336 subjects) was

selected. Including phenol, a total of 25 solvents were authorized to be used in the plant. The

exposure to any particular solvent was determined from the records of annual authorization for use

of these solvents in each work area. Only workers who had cumulative exposures of more than 1

year were considered exposed.

On the basis of the analysis of the age-adjusted exposure ORs, no association was seen

between phenol exposure and mortality from stomach cancer, prostate cancer, lymphosarcoma and

reticulum cell sarcoma, lymphatic leukemia, or respiratory cancer. However, this study had several

major limitations. One limitation was that the estimation of whether workers were exposed to a

solvent was based solely on authorization and not on actual usage, which would tend to lead to an

overestimation of exposure. In addition, the analysis was based solely on a qualitative evaluation of

whether a given solvent was used; no estimates of exposure were made, and so no exposureresponse assessment was conducted. Although smoking can confound evaluation of cancer risk, this

factor was not investigated. Finally, it was common for workers to be simultaneously exposed to

multiple solvents; therefore, solvents other than phenol may have affected the study outcome. In

this study, phenol exposure was not associated with a risk of several cancers, but this lack of an

association cannot be considered definitive because of the study limitations mentioned above.

In an occupational epidemiology study, Dosemeci et al. (1991) evaluated mortality among

14,861 white male workers in five companies that used formaldehyde and phenol. Unfortunately,

the phenol exposure was confounded by co-exposure to other compounds, such as formaldehyde,
30

asbestos, urea, melamine, hexamethylenediamine, wood dust, plasticizers, carbon black, ammonia,

and antioxidants. On the basis of phenol concentrations obtained from historical monitoring and

industrial hygiene surveys, the investigators assigned each job/department/year combination to

groups with no, low, medium, or high phenol exposure and then calculated cumulative exposure. 

Compared with the entire U.S. population, the entire cohort, had no significant increases in

standardized mortality ratios (SMRs) for all causes of death or any diseases. The phenol-exposed

workers as a group had slightly elevated SMRs for cancers of the esophagus (1.6), rectum (1.4),

kidney (1.3), and Hodgkin’s disease (1.7); however, none of these increases were statistically

significant when compared with those in general population. In addition, an analysis of mortality by

level of cumulative exposure showed that none of these increases had dose-response relationships

with exposure to phenol. The only significant observations were decreases of SMRs for infective

and parasitic diseases and for accidents in the entire cohort and exposed workers. These

observations were attributed to the healthy worker effect. This study provided no evidence of

phenol-induced morbidity, mortality, or carcinogenicity.

Baj et al. (1994) reported an epidemiology study of 22 Polish office workers (18 females and

4 males) exposed to Ksylamit® vapor for 6 months and 27 age- and sex-matched healthy volunteers

from the same town. The exact composition of the Ksylamit® vapor was not reported. The study

authors stated that Ksylamit® consists of “a mixture of chlorinated benzenes, pentachlorophenol, -

chloronaphthalene, chloroparaffin and kerosene.” The only exposure information reported was that

at the end of 6-month exposure period the concentrations of formaldehyde and phenol in the

workplace atmosphere were 0.8 mg/m3

 and 1.3 mg/m3, respectively. The study authors did not

address how exposure to formaldehyde or phenol resulted from the reported product constituents. In

addition, it cannot be determined from the presented information whether the analytical methods

used would differentiate between phenol and pentachlorophenol (ATSDR, 1998).

The exposed workers reported chronic symptoms such as headache, cough and sore throats,

burning eyes, and fatigue, but morbidity during the 6-month exposure period was comparable to that

of the controls. Although all evaluated hematological parameters were normal in the exposed

workers as a group, some statistically significant changes were observed in a subset of eight workers

who had elevated urinary phenol levels 3 days after the last day of exposure (mean of 18.2 mg/L,

compared with 12.1 mg/L in the exposed workers and 7.9 mg/L for the general population). 

Compared with the matched controls, there was a small, but statistically significant decrease among
31

in erythrocyte counts and a statistically significant increase in eosinophil and monocyte counts. 

Levels of CD3, CD4, and CD8 lymphocytes were also decreased in the exposed group, but there was

no effect on the CD4/CD8 ratio, and the effect was not stronger in the apparently more

highlyexposed subset. Decreases in lymphocyte proliferation induced by phytohemagglutinin and

alloantigens were also observed in exposed workers, whereas reactivity to concanavalin A (Con A)

was unchanged.

These results suggest that exposure to Ksylamit® could affect the immune and hematological

systems. However, the poor characterization of the chemical exposure, including uncertainties

regarding the source of the phenol as well as the marginal dose-response for phenol in urine, mean

that conclusions regarding the contribution of phenol to the observed effects are limited. 

Shamy et al. (1994) reported a cross-sectional investigation of phenol-induced biochemical

changes in workers at an oil refining plant in Egypt. The study included 20 workers who were

exposed to a time=weighted average concentration of 5.4 ppm phenol and 30 office workers who

had no exposure to organic solvents. The phenol-exposed workers worked in the aromatic

extraction of distillates; other potential exposures were not described. The mean concentration of

phenol in spot urine samples was 68.6 and 11.5 mg/g creatinine in the exposed and control groups,

respectively. The average duration of exposure was 13.15 years. At the end of the shift of the last

working day of the week, blood samples were collected for hematological and serum biochemistry

evaluations.

Small but statistically significant increases (approximately 55% and 80%, respectively) were

observed in serum glutamic oxaloacetic transaminase (SGOT) and serum glutamic pyruvic

transaminase (SGPT). There were also small but statistically significant increases in hemoglobin,

hematocrit, mean corpuscular hemoglobin, and mean corpuscular volume, although there was no

effect on red blood cell count. This finding is in contrast with those of laboratory animal studies, in

which decreases in erythrocytes and hematocrit have been reported. Other small, but statistically

significant changes included increased basophils and neutrophils, decreased monocytes, and

increased clotting time. A nonsignificant increase in prothrombin time and decrease in platelets was

also observed.

Overall, these data suggest subclinical effects on the liver and hematopoietic system, based

on the small changes in SGOT and SGPT; the observed increases in the hematology endpoints are

not adverse. Although the authors described the workers as exposed to phenol alone and compared

them with other workers exposed to mixed solvents, it appears that the phenol-exposed workers may
32

have also been exposed to other organic compounds that can cause hepatic or hematologic effects,

and the observed effects cannot be clearly attributed to phenol exposure. 

Merliss (1972) reported a case of phenol marasmus. A laboratory technician had been

frequently exposed to phenol through vapor or skin contact for 13.5 years. He developed lessened

appetite, weight loss, muscle pain in his legs and arms, and dark color in his urine. Serum

biochemical evaluations indicated liver damage, with SGOT and SGPT at much higher than normal

levels. The patient’s symptoms improved after the exposure ended.

4.2. PRE-CHRONIC, CHRONIC STUDIES AND CANCER BIOASSAYS IN

LABORATORY ANIMALS

4.2.1. Oral

Acute lethality of oral phenol has been evaluated in numerous animal studies. Oral LD50

values in rodents range from 300 mg/kg to 650 mg/kg (Deichmann and Witherup, 1944; Von

Oettingen and Sharpless, 1946; Flickinger, 1976; Berman et al., 1995). The acute toxicity of phenol

when administered by gavage appears to be at least partly dependent on phenol concentration or

total administered volume (Deichmann and Witherup, 1944; NTP, 1983a), and it may be more

severe in young animals (Deichmann and Witherup, 1944). In addition to lethality, acute oral dosing

has been reported to induce adverse renal (tubular necrosis, protein casts, papillary hemorrhage),

hematological (reduction in poly/normochromatic erythrocyte ratio), respiratory (dyspnea and rales),

neurological (muscle twitching, decreased motor activity, ataxia, tremors, convulsions, coma), and

reproductive and developmental effects (Berman et al., 1995; Narotsky and Kavlock, 1995; Liao and

Oehme, 1981; Moser et. al. 1995; NTP, 1983a,b).

As shown in Table 2, there is an extensive database of oral studies relevant to the RfD. 

Chronic drinking water studies have been conducted in rats and mice, but the only noncancer

endpoints evaluated were body weight and histopathology (NCI, 1980). Hematology and serum

biochemical evaluations were included in a recent two-generation drinking water study conducted in

rats (Ryan et al., 2001; available in unpublished form as IIT Research Institute, 1999). A specialized

subchronic neurotoxicity study was conducted with rats exposed to phenol in drinking water

(ClinTrials BioResearch, 1998).

These drinking water studies consistently found effects only at exposure levels where water

consumption was also decreased, sometimes by as much as 80%. The decreased water consumption

was presumably due to poor palatability of the drinking water. Effects seen in these studies included
33

tremors, decreased final body weight as compared with the controls (possibly as a result of

dehydration), decreased motor activity, and kidney inflammation. Decreased pup survival was also

observed in the two-generation study. The only drinking water study in which effects were seen in

the absence of decreased water consumption was a 28-day study with mice by Hsieh et al. (1992). 

In that study, decreased hematocrit and decreased immune endpoints were observed at doses at least

an order of magnitude lower than the NOAELs in the other drinking water studies. Although the

Hsieh et al. study is useful for hazard identification, its findings need to be confirmed before they

can be used in dose-response assessment. The differing findings of this study and the abovementioned two-generation study in rats in which no immunological effects were observed suggest

marked interspecies differences between rats and mice for this endpoint. 

Toxicity in gavage studies with phenol is typically much higher than that in the drinking

water studies. NOAELs for systemic effects were 5–10 - fold lower in gavage studies (Berman et

al., 1995; Moser et al., 1995; Dow Chemical Co., 1945) than those seen in the drinking water

studies. Effects observed included tremor and liver and kidney histopathology. As described in

greater detail in Section 4.5, this difference between gavage and drinking water exposure is 
34

Table 2. Summary of Oral Toxicity Studies

Strain,

Species,

Sex

Reference Dose Route/

Volume

Duration Response NOAEL

mg/kg-day

LOAEL

mg/kg-day

Comments

Systemic Toxicity

F344 rat,

50/sex/

group

NCI, 1980 0, 2500, 5000

ppm; 

0, 260, 585 (M),

0, 280, 630 (F)

mg/kg-day

Drinking

water

103 weeks Kidney inflammation

and decreased body

weight (compared to

controls) in both sexes, 

decreased water

consumption at high

dose

260 585 Study authors stated there

were no noncancer effects,

but independent evaluation

for this assessment found

significant increase in

kidney inflammation.

B6C3F1

mouse, 

50/sex/

group

NCI, 1980 0, 2500, 5000

ppm;

0, 450, or 660

mg/kg-day

Drinking

water

103 weeks Decreased body

weight, decreased

water consumption 

450 660 Effect apparently

secondary to decreased

water consumption.

F344 rat,

10/sex/

group

NCI, 1980 0, 100, 300, 1000,

3000, 10,000

ppm;

0, 16, 48, 160,

480, 800 (M),

0, 17, 51, 170,

510, 1140 (F)

mg/kg-day

Drinking

water

13 weeks Decreased body

weight, decreased

water consumption.

480 800 Range-finding for

bioassay; effect apparently

secondary to decreased

water consumption.

B6C3F1

mouse,

10/group

NCI, 1980 0, 100, 300, 1000,

3000, 10,000

ppm;

0, 25, 75, 250,

450, 500 (M),

0, 26, 78, 260,

468, 520 (F)

mg/kg-day

Drinking

water

13 weeks Decreased body

weight, decreased

water consumption.

450 500 Range-finding for

bioassay; effect apparently

secondary to decreased

water consumption.
Strain,

Species,

Sex

Reference Dose Route/

Volume

Duration Response NOAEL

mg/kg-day

LOAEL

mg/kg-day

Comments

35

SpragueDawley

rat, 15/sex/

group

ClinTrials

BioResearch,

1998

0, 200, 1000,

5000 ppm;

0, 18, 83.1, 308

(M),

0, 24.6, 107, 360

(F) mg/kg-day

Drinking

water

13 weeks Decreased motor

activity in females,

decreased body weight

in males and females,

decreased water

consumption

107 360 Specialized neurotoxicity

study; decreased body

weight apparently

secondary to decreased

water consumption;

unpublished GLP study

BMDL = 219 mg/kg-day 

CD-1

mouse, 

5 M/group

Hsieh et al.,

1992

0, 4.7, 19.5, 95.2

ppm;

0, 1.8, 6.2 or 33.6

mg/kg-day

Drinking

water

28 days Decreased immune

endpoints (particularly

in plaque-forming cell

assay and ELISA),

decreased red blood

cells and hematocrit

6.2 33.6 Study limited by small

sample size; confirmation

in a study conducted

according to modern

methods would be useful. 

Rat

10/group,

strain &

sex NS

Dow

Chemical

Co., 1945

0, 50, 100

duration adjusted: 

0, 35.7, 71.4

mg/kg-day

Gavage/

volume

NS

6 months,

5 days/wk

Liver and kidney

histopathology;

mortality in 4/10 at

low and high doses

None 35.7 Unpublished study, small

group size, incomplete

reporting; authors raised

questions about the

mortality, although it is

unclear what the issue was.

F344 rat, 

8 F/group

Berman et al.,

1995; Moser

et al., 1995

0, 4, 12, 40, 120

mg/kg-day

Gavage/

1 mL/kg

14 days Tubular degeneration,

tremor, increased

rearing post-exposure 

12 40 Small group size, small

dosage volume

Reproductive and Developmental Toxicity

SpragueDawley rat

30/sex/

group P1; 

20/sex/

group F1

Ryan et al.,

2001; IIT

Research

Institute,

1999

0, 200, 1000,

5000 ppm;

0, 14.7, 70.9, 301

(P1 M),

0, 20, 93, 320.5

(P1 F)

mg/kg-day

Drinking

water

2 generationsDecreased parental

and pup body weight,

decreased pup

survival, decreased

water consumption

70.9 301 Study also included

evaluation of hematology,

serum biochemistry, and

developmental landmarks.

Effects may be secondary

to decreased water

consumption. Decreased

uterine weight at all doses,

but not considered adverse
Strain,

Species,

Sex

Reference Dose Route/

Volume

Duration Response NOAEL

mg/kg-day

LOAEL

mg/kg-day

Comments

36

SpragueDawley

Rat

25 F/group

Argus

Research

Laboratories,

1997

0, 60,120 or 360

mg/kg-day

Gavage

10

mL/kg

GD 6-15 Decreased maternal

weight gain;

Decreased fetal body

weight and delayed

ossification

60

(maternal)

120 (dev)

120

(maternal)

360 (dev)

Doses were divided into 3

administrations/day

One dam died at 360

mg/kg-day

Unpublished GLP study

BMDL = 93 mg/kg-day

CD Rats

20-22

F/group

NTP, 1983a 0, 30, 60 or 120

mg/kg-day

Gavage

5 mL/kg

GD 6-15 Decreased fetal body

weight

120

(maternal)

60 (dev)

None

(maternal)

120 (dev)

None

Developmental LOAEL is

equivocal

BMDL not calculated,

because data on fetal

weight by sex was not

available

CD Rat

5-10

F/group

NTP, 1983a 0-250 

mg/dg-day

Gavage

1-7.5

mL/kg

GD 6-15 Toxicity (tremors,

liver and lung

pathology, death)

markedly higher in

smaller dosing

volume. 

N/A N/A Range-finding studies.

CD-1

Mouse,

31-36

F/group

NTP, 1983b 0, 70, 140 or 280

mg/kg-day

Gavage

10

mL/kg

GD 6-15 4/36 dams died,

tremors, reduced

maternal body weight

(10%); reduced fetal

body weight

140

(maternal)

140 (dev)

280

(maternal

FEL)

280 (dev)

None

F344 Rat

15-20

F/group

Narotsky and

Kavlock,

1995

0, 40, 53.3

mg/kg-day

Gavage

1 mL/kg

GD 6-19 Maternal rales and

dyspnea, marginal

decreases in maternal

body weight 

N/A N/A Screening study

NS = Not Stated; dev = developmental; ELISA = enzyme-linked immunosorbent assay 
37

consistent with toxicokinetic data suggesting that toxicity is correlated with peak blood concentrations

rather than with total daily intake.

Several developmental toxicity studies are available (Argus Research Laboratories, 1997;

NTP, 1983a; NTP, Narotsky and Kavlock, 1995). All of these studies were conducted via the gavage

route, although the Argus Research Laboratories study used large dosing volumes and a divideddosing protocol, apparently to reduce the effect of peak blood levels. The developmental toxicity

studies found that the primary fetal effect is decreased body weight, which occurs at doses similar to

those that cause decreased maternal body weight gain. The National Toxicology Program (NTP,

1983a) also found that toxicity increased markedly if the same phenol dose was administered in a

lower dosing volume. Because the observed signs of toxicity included tremors and liver and kidney

pathology and were not targeted to the portal of entry, the effect of dosing volume is not related to the

concentration of a direct-contact toxicant. 

The National Cancer Institute (NCI, 1980) conducted a carcinogenicity bioassay in which

F344 rats (50/sex/group) received analytical-grade phenol (approximately 98.5% pure) in drinking

water at concentrations of 0, 2500, or 5000 ppm for 103 weeks and were sacrificed 1–2 weeks later. 

Using the reference water intake of 0.13 and 0.14 L/kg-day for chronic exposure of male and female

F344 rats, respectively (U.S. EPA, 1988), the doses can be estimated as 0, 260, and 585 mg/kg-day for

male rats and 0, 280, and 630 mg/kg-day for female rats. The doses shown here were adjusted to

account for the reported water consumption of 80% and 90% of control at the low and high doses,

respectively. The animals were observed daily for clinical signs and examined weekly for palpable

masses. Body weights and food consumption were recorded every 2 weeks for the first 12 weeks and

then monthly thereafter; water consumption was recorded weekly. 

At the end of study, the animals were sacrificed, and complete gross and histopathological

examinations were performed. Organs and tissues examined included the bone marrow, spleen,

cervical and mesenteric lymph nodes, heart, liver, kidney, thyroid, reproductive organs, brain, and

other major tissues. No evaluation of other noncancer endpoints, such as hematological effects or

serum biochemistry, was conducted. 

The survival rate at study termination was comparable among all three groups of males

(approximately 50%) and females (approximately 75%). Dose-related decreases in body weight

compared with the controls were observed in male and female rats, with a decrease of approximately

15% in high-dose males and approximately 10% in high-dose females. Water consumption was

reduced by approximately 10% at the high dose. The authors stated that the non-neoplastic lesions

were similar to those occurring naturally in aged F344 rats. However, an analysis conducted for this

assessment found statistically significant increases (using a chi square test) in chronic kidney
38 

inflammation in high-dose males and females; there were no significant changes at the low dose. No

other differences in the incidence of non-neoplastic lesions between the controls and exposed rats

were observed. 

On the basis of increased kidney inflammation and decreased body weight as compared with

controls at the high dose of 5000 ppm (585 mg/kg-day for males and 630 mg/kg-day for females), the

NOAEL in this study can be considered to be the low dose, 260 mg/kg-day in males and 280 mg/kgday for females, resulting in an overall study NOAEL of 260 mg/kg-day. These effects also indicate

that the maximum tolerated dose (MTD) was reached.

In the NCI (1980) rat study, there were no dose-related trends in cancer incidence in male or

female rats, but the authors reported several tumors for which statistically significant increases were

seen in low-dose males only, based on pairwise comparisons. These increases were seen in the

incidences of pheochromocytomas of the adrenal medulla (13/50, 22/50, and 9/50 in the control, low-

, and high-dose groups, respectively) and “leukemias or lymphomas” (18/50, 31/50, and 25/50). The

historical control incidences of pheochromocytomas in the bioassay program was 9% (data for the

test laboratory were not reported), and the historical control incidence of leukemias or lymphomas in

the test laboratory was 26%. The study authors stated that the leukemias were “of the type usually

seen in untreated F344 rats.” There were no significant increases in tumor incidence in any tissue in

female rats. Because there was no clear dose-response in males and the tumors were not observed in

female rats, an association between the tumors and phenol exposure cannot be established. NCI

concluded that phenol was “not carcinogenic in male or female F344 rats.” However, the report

noted uncertainties regarding the possible increase in leukemia in male rats, and the NCI reviewers

recommended that phenol be considered for a retest. 

In a parallel study, NCI (1980) administered phenol at 0, 2500, or 5000 ppm in drinking water

to B6C3F1 mice (50/sex/group) for 103 weeks and sacrificed the mice 1–2 weeks later. For B6C3F1

mice, the reference water intake was 0.24 L/kg-day for both sexes. The study found that water

consumption was decreased to 75% and 50–60% of the control levels at the low and high doses,

respectively. The resulting doses (adjusting for decreased water intake) were 0, 450, and 660 mg/kgday for both sexes. Dose-related decreases in body weight as compared with the controls were

attributed to the decrease in water consumption. No other clinical signs of toxicity were observed,

and mortality rates (approximately 10% in males and 20% in females) were comparable between

experimental and control groups. 

Histopathological examination and statistical analyses revealed no phenol-related signs of

toxicity or carcinogenicity; lesions in all systems observed in the dosed groups were comparable to

those in the controls. NCI concluded that, under the conditions of the assay, phenol was not
39

carcinogenic in male or female B6C3F1 mice. On the basis of the decreased body weight as compared

with controls observed at 5000 ppm, the low dose of 2500 ppm (450 mg/kg-day) can be considered the

study NOAEL. The observed effect, however, is likely secondary to the decreased water consumption

due to poor palatability.

In light of the marked decrease in water consumption, higher doses of phenol in drinking water

probably could not have been tested. If the authors had attempted to overcome the palatability issue by

administering the high dose by gavage rather than in drinking water, high toxicity would have been

expected in light of the higher toxicity of phenol administered by gavage (see Section 4.5 and Table 2). 

These considerations suggest that an MTD was also reached in mice, although the conclusion is less

clear than for rats.

In the range-finding test for the carcinogenicity bioassay (NCI, 1980), F344 rats and B6C3F1

mice (10/sex/group) were given drinking water containing 0, 100, 300, 1000, 3000, or 10,000 ppm

phenol (at a purity of 98.47%) for 13 weeks. Using the reference water intake of 0.16 and 0.17 L/kgday for subchronic exposure of male and female F344 rats, respectively (U.S. EPA, 1988), the doses

can be estimated as 0, 16, 48, 160, 480, and 800 mg/kg-day for male rats and 0, 17, 51, 170, 510, and

1140 mg/kg-day for female rats. The high doses shown here were adjusted to account for the

decreased water consumption described below. For B6C3F1 mice, the reference water intake was 0.25

L/kg-day for males and 0.26 L/kg-day for females. The corresponding doses (adjusting for decreased

water intake at the high dose) were 0, 25, 75, 250, 450, and 500 mg/kg-day for males and 0, 26, 78,

260, 468, and 520 mg/kg-day for female mice.

Body weights, appearance, behavior, and food and water consumption were recorded weekly. 

After 13 weeks, all animals were sacrificed and tissues were subjected to histopathological

examinations. All of the rats and mice survived the phenol treatment. The only significant observation

was the decreased final body weights (compared to controls) in rats of both sexes (11–14%) and in

male mice (12%) that received 10,000 ppm. Because drinking water consumption in these groups was

decreased to 50–70% (rats) and 20–60% (mice) of the control value, the decreased body weight was

likely due to the low water consumption. No histopathological changes attributable to phenol

treatment were observed. 

This study suggests that the second-highest dose (480 mg/kg-day for male rats, 510 mg/kg-day

for female rats, 450 mg/kg-day for male mice, and 470 mg/kg-day for female mice) was a NOAEL,

based on the decreased final body weight (compared to controls) at 10,000 ppm, which was secondary

to decreased water consumption due to poor palatability at the high dose.

In an unpublished 13-week neurotoxicity study conducted according to good laboratory

practices (GLP) guidelines (ClinTrials BioResearch, 1998), groups of 15 male and 15 female Sprague-
2

This study has not been peer-reviewed, but it was conducted (with minor deviations) according

to EPA guidelines for a neurotoxicity screening battery, it is well documented, and it contributes useful

information to the hazard identification and dose-response portions of the assessment. The study was

designed to comply with the U.S. EPA Enforceable Consent Agreement for Phenol (Docket No. OPPTS42150).

3

Results of the palatability study were not provided in the IITRI study report (IIT Research

Institute, 1999; Ryan et al., 2001), which reports the results of the two-generation reproduction study. 

The reproduction study was conducted during the same time period as the neurotoxicity study and

reported similar problems of markedly decreased drinking water consumption at the high dose of 5000

ppm (see Section 4.3).

40

Dawley rats received phenol via drinking water at concentrations of 0, 200, 1000, or 5000 ppm (at a

purity of 100%) for 13 weeks, followed by a 4-week recovery period.2

 The authors calculated that the

average doses were 0, 18.1, 83.1, and 308.2 mg/kg-day for males and 0, 24.6, 107.0, and 359.8 mg/kgday for females. These concentrations were chosen on the basis of preliminary palatability studies

conducted at a different laboratory (IITRI project No. L08657).3

During the exposure period, clinical signs and water intake were recorded daily, and body

weight and food consumption were recorded weekly. In addition, a functional observational battery 

(FOB) and a motor activity test were conducted pre-study and once each during weeks 4, 8, 13, and 17. 

At the end of the exposure and the end of the recovery period, five rats/sex in the control and 5000

ppm groups underwent neuropathological evaluations (including a thorough evaluation of the brain and

several nerves). The rest of the rats in the ClinTrials BioResearch (1988) study were sacrificed at the

end of the 4-week recovery and were subjected to gross necropsy.

One high-dose female was euthanized on day 14 due to poor condition. Clinical signs observed

in this female prior to sacrifice included dehydration, hunched posture, tremors, reduced activity, and

cold to touch. Among the rest of the high-dose animals, the primary clinical sign was dehydration,

which was accompanied by reduced activity and tremors in one female and by a thin appearance in

additional animals. Dehydration was also observed in mid-dose rats (2/15 in each sex). Dehydration

was assessed qualitatively and independently of drinking water consumption by grabbing the scruff on

the back of the animal’s neck; a delay in returning to the normal position was considered dehydration. 

Dehydration was associated with marked decreases in water consumption at the high dose and

smaller decreases at the mid dose. Decreases in water consumption were more pronounced in females

than in males and were most evident during the first week of dosing. Water consumption was

decreased to approximately 90% of the control level in mid-dose males and females, to approximately

60% of control levels in high-dose males, and to approximately 55% (40% during the first week) of

control levels in high-dose females. Water consumption rebounded to levels higher than those of
41

controls during the recovery period. The decreased water consumption was likely due to the poor

palatability of phenol at high concentrations rather than being a manifestation of an overt toxicological

effect. In addition, the high-dose group had decreased body weights when compared to controls (8%

for males and 12% for females) and decreased food intake (approximately 10% for males and 10–20%

for females). 

The only toxicologically significant neurological effect was decreased motor activity in

females. There was a statistically significant reduction in total group mean motor activity counts at

week 4 in the 5000 ppm group. The authors reported that the rate of linear change of motor activity

with time was also significantly decreased at weeks 8 and 13 in the 1000 ppm and 5000 ppm groups,

although supporting data were not provided. Motor activity in females at week 4 exhibited a dose

dependency at the first five (of six) analysis intervals, although the total counts for the low- and middose groups were not significantly different from control. High-dose females also had markedly lower

total activity counts than did controls and lower counts in the first four intervals, at week 4, although

there was no statistically significant difference in mean total counts (Table 3). By contrast, the highdose males had markedly lower group mean activity counts pre-study but activity comparable to or

higher than the controls at weeks 4, 8, and 13.

 The authors attributed the decreased activity to dehydration, noting that the control group mean

total activity increased by >20% at week 4 as compared with pre-study levels, whereas activity of

dehydrated females in the 5000 ppm group at week 4 was decreased by 17%, and activity of females in

this group that were not dehydrated increased by 2%. To address whether the decreased activity could

be attributed to dehydration, this assessment evaluated the data in greater detail. Table 4 presents the

individual animal data for week 4 total motor activity counts and compares them with the individual

animal dehydration data. If the individual clinical data reported an animal as dehydrated, the days of

that notation are shown. With the exception of animal 4502 (which died) and animal 4507 (which had

severe dehydration) dehydration was noted as slight or moderate. For clarity of presentation, the

individual animal data are shown for the control and high-dose groups, but only the average data are

shown for the low- and mid-dose groups. 

Table 3. Total Activity Counts in Rats Provided Phenol in Drinking Water 

(ClinTrials BioResearch, 1998)

Dose

Group

Prestudy

(Mean ±SD)

Week 4

(Mean ±SD)

Week 8 

(Mean ±SD)

Week 13

(Mean ±SD)

Females
42

Control 384±116 468±118 436±75 309±77

200 ppm 386±89 451±149 440±99 338±66

1000 ppm 384±103 394±78 436±104¶ 343±124¶ 

5000 ppm 372±142 337±127** 363±111¶¶ 366±145¶¶

Males

Control 354±109 339±89 320±90 260±68

200 ppm 340±107 346±132 323±88 256±78

1000 ppm 335±126 356±137 359±105 274±103

5000 ppm 277±59 321±95 352±91 275±116

**Significantly different from control, p<0.01 (T-test)

¶Linear constructed variable significantly different from control, p<0.05 (T-test)

¶¶Linear constructed variable significantly different from control, p<0.01 (T-test)
43

Table 4. Individual Data on Dehydration and Week 4 Motor Activity in Rats Provided Phenol in Drinking Water

(ClinTrials BioResearch, 1998) 

Group 1

(Control)

Group 2

(200 ppm)

Group 3 

(1000 ppm)

Group 4

(5000 ppm)

Animal # Total

Counts

Days

Dehydrated 

Animal

#

Total

Counts

Days

Dehydrated

Animal 

#

Total

Counts

Days

Dehydrated

Animal

#

Total

Counts

Days

Dehydrated

1501 383 No 2501 397 No 3501 322 No 4601 501 14, 21

1502 321 No 2502 529 No 3502 402 18, 21 4502 No data - sacrificed 

day 14 due to poor

condition

1502 621 No 2503 427 No 3603 270 No 4503 227 7, 14, 21, 28

1504 437 No 2504 558 No 3504 370 No 4504 258 14, 21, 28,

35, 42, 70

1505 630 No 2505 245 No 3505 572 21, 28, 35,

42, 49, 56,

70

4505 396 No

1506 365 No 2506 537 No 3506 452 No 4506 277 70

1507 591 No 2507 470 No 3507 461 No 4507 399 7-9, 11, 12,

13-15, 17,

20, 21, 70,

77

1508 318 No 2508 284 No 3508 342 No 4508 271 No
Group 1

(Control)

Group 2

(200 ppm)

Group 3 

(1000 ppm)

Group 4

(5000 ppm)

Animal # Total

Counts

Days

Dehydrated 

Animal

#

Total

Counts

Days

Dehydrated

Animal 

#

Total

Counts

Days

Dehydrated

Animal

#

Total

Counts

Days

Dehydrated

44

1509 479 No 2509 527 No 3509 462 No 4509 387 No

1510 469 No 2510 823 No 3510 452 No 4510 450 No

1511 309 No 2511 561 No 3511 390 No 4511 439 7

1512 574 No 2512 424 No 3512 383 No 4512 130 No

1513 566 No 2513 302 No 3513 320 No 4513 242 49, 56

1514 381 No 2514 289 No 3514 403 No 4514 180 7, 14, 21, 28

1515 578 No 2515 386 No 3515 311 No 4515 556 No

Overall

Average

468 -- 451 -- 394 -- 337 --

Average - 

Dehydrated

animals

N/A -- N/A -- 487

(n=2)

-- 315 --

Average - 

nondehydrated

animals

468 -- 451 -- 380 -- 365 --

N/A = Not applicable
45

Only two animals in the mid-dose group were reported as dehydrated on any day, and neither

of these animals had decreased motor activity. As shown, the average activity was lower in the

dehydrated high-dose females than in those not reported as dehydrated, but an association of

decreased activity with dehydration was not consistently supported on an individual-animal basis. 

(For the purpose of calculating averages, animals were considered dehydrated if they dehydrated at

any point in the study. This is a limitation to the analysis, because some were reported as dehydrated

only prior to week 4 and others were reported as dehydrated only after week 4. In addition, basing

the analysis on the clinical sign of dehydration may not appropriately reflect whether the animals

were dehydrated, because no objective measure of dehydration was used and because decreased

water consumption in this group occurred throughout the study.) 

As shown in Table 4, animal 4601 was reported as dehydrated on days 14 and 21, but it had

one of the highest total activity counts. Conversely, animal 4512 had the lowest activity count, but it

was never reported as being dehydrated. Furthermore, the mean activity of the dehydrated high-dose

females was 67% of concurrent controls, compared to 78% of concurrent controls for the

nondehydrated high-dose females. These data indicate that the difference between the control and

high-dose animals was greater than the difference between the dehydrated and nondehydrated

animals at the high dose. 

Overall, the data indicate that there was not a tight linkage between dehydration and

decreased motor activity in the high-dose females. The data for high-dose males also did not indicate

a clear correlation between low activity and dehydration. The clinical signs for one high-dose male

(4003) for week 2 included severe dehydration and decreased activity, but no effect (i.e., no

dehydration or decreased activity) was seen when the animal underwent the objective activity

analysis in week 4. The finding of dehydration in males without the accompanying decrease in

activity further supports the conclusions that only severe (not mild or moderate) dehydration results

in decreased motor activity levels and that the decrease observed in females was phenol related. 

Conversely, the absence of other findings in the FOB and the presence of a statistically significant

effect on motor activity only at 4 weeks and not at later time points argue against a neurotoxic effect

of phenol.

As an additional investigation of whether decreased motor activity was related to

dehydration, the very limited literature on water deprivation and motor activity was reviewed. 

Campbell and Cicala (1962) evaluated the effects of terminal water and food deprivation (i.e.,

deprivation until death from dehydration or starvation) on motor activity of male and female Wistar

rats. Motor activity was measured using a stabilimeter, which is similar to the figure-8 mazes used in
46

the ClinTrials BioResearch (1998) study in that ambulation (as opposed to simply movement) is

measured. The study found that water deprivation alone did not result in decreased motor activity

until approximately days 5–7 (depending on age), at which time activity continuously declined until

death. By contrast, food deprivation resulted in an initial increase in activity followed by decreasing

activity until death.

Only the pooled data for males and females were reported. These results are not directly

comparable to the results of the ClinTrials study because the latter involved long-term, lower-level

dehydration; however, they do support the conclusion that the decreased motor activity in high-dose

females was due at least partially to phenol exposure. The most appropriate way to address this issue

would be to conduct the neurotoxicity study with a water-restricted control group. Overall, based on

the decreased motor activity, the study NOAEL in females was 1000 ppm phenol (107 mg/kg-day)

and the Lowest Observed Adverse Effect Level(LOAEL) was 5000 ppm (360 mg/kg-day ). No

LOAEL was identified in males; the high dose of 308 mg/kg-day was a NOAEL. A 95% lower

confidence limit on the benchmark dose (BMD) of 219 mg/kg-day was calculated for decreased

motor activity in week 4 in this study (see Appendix B).

Hsieh et al. (1992) investigated the effects of phenol exposure on hematological, immune,

and neurochemical endpoints in a study of 6-week-old male CD-1 mice administered actual

concentrations of 0, 4.7, 19.5, or 95.2 ppm in drinking water for 28 days. On the basis of measured

concentrations and water intake, the authors reported that the corresponding daily doses were 0, 1.8,

6.2, and 33.6 mg/kg-day.

The mice were housed in groups of five per cage. Drinking water was prepared and changed

every 3 days. Drinking water was provided in glass water bottles with stainless sipper tubes

containing ball bearings to minimize evaporation; the bottles were shaken frequently during

treatment. Food and water consumption were monitored continuously, and the animals were

weighed weekly. After 28 days, the mice were sacrificed by decapitation, gross pathological

examinations were performed, and the liver, spleen, thymus, and kidney were weighed. Blood was

taken at sacrifice for analysis. Splenocytes were prepared for analysis of mitogen-stimulated

lymphocyte proliferation, mixed lymphocyte response, and cell-mediated cytolytic response.

Data were reported for five animals per group for each assay. During the 28-day exposure,

no mortality and no overt clinical signs occurred in exposed mice. Phenol treatment had no effect on

food or water consumption or on body weight gain. Exposed mice had no gross lesions in the liver,

kidney, spleen, thymus, lung, heart, or brain, and there were no effects on organ weights for the 

liver, kidney, spleen, and thymus. As shown in Table 5, a decreased antibody response to sheep red
47

blood cells was observed, as indicated by both the plaque-forming cell (PFC) assay (expressed as

PFC/million spleen cells and PFC/spleen) and the antibody titer using an enzyme-linked

Table 5. Effects of Phenol Exposure on Spleen Cellularity

and Selected Blood Parameters in Mice and Rats

Concentration

(mg/L)

Dose (mg/kgday)PFC/106

 splenic

cells

PFC/total spleen Antibody titer b

Hsieh et al. (1992) - 4-week study in CD-1 mice

0 0 1,123 ± 99a 265,947 ± 53,099 0.446 ± 0.039

4.7 1.8 896 ± 60 214,678 ± 17,500 0.392 ± 0.068

19.5 6.3 795 ± 49c 207,659 ± 18,886 0.325 ± 0.019c

95.2 33.6 616 ± 83c 130,185 ± 18,202c 0.263 ± 0.037c

IIT Research Institute (1999); Ryan et al. (2001) - 2-generation study in Sprague-Dawley rats, effects in P1

generation

0 0 1343±890d 5.54x105±3.70x105 Not assayed

200 15 1668±788 6.42x105

±3.40x105 Not assayed

1000 71 1781±1151 9.01x105±7.16x105 Not assayed

5000 301 1880±865 9.81x105

±5.02x105 Not assayed

Positive control

(n=5)

cyclophosphamide0±0 0±0 Not assayed

a

 Values are given as mean ± S.E. (n=5). bArbitrary as change in absorbance at 490 nm, using 1:2000

diluted serum. c Significant (P<0.05) difference from the control value. d Values are given as mean

±SD (n=8-9)
48

immunosorbent assay (ELISA). Two of these measures were statistically significantly decreased at

the mid dose, and PFC/spleen was significantly decreased only at the high dose. Decreases in the

absolute splenocyte lymphoproliferative responses to mitogens and the mixed lymphocyte response

(the proliferative ability of splenic lymphocytes in response to alloantigens) were also observed at

the high dose; there was no effect on the cytolytic response to tumor cells at any dose. 

Although these assays were conducted according to the methods of the day, the latter two

assays do not conform to modern protocols, and there is little biological significance in the results of

the mitogen response assay. In particular, the approach used cannot distinguish between an effect on

the lymphocyte’s ability to initiate a proliferative response and confounding due to contamination by

nucleated red blood cell precursors.

A statistically significant, dose-related decrease in erythrocyte counts was observed at all

doses (Table 6). The hematocrit was decreased only at the high dose. A decreased erythrocyte count

in the absence of an effect on hematocrit may have been due to macrocytosis (enlarged erythrocytes),

but insufficient data were provided to evaluate this possibility. In the absence of such data, the

decreased erythrocyte counts are insufficient to form the basis for identification of a LOAEL. The

erythrocyte counts in all dosed groups were markedly lower than the historical control values

provided by the animal distributor (Charles River Laboratories, 1986), although the hematocrit

concentration in all groups was above the historical control mean. There was no effect on total or

differential leukocyte counts. Interestingly, total white blood cells for all groups, including the

controls, were below the historical control data provided by the distributor.

Hsieh et al. (1992) also observed dose-related decreases in the concentration of several

neurotransmitters and their metabolites in the brain, including levels of norepinephrine, indoleamine

serotonin, and dopamine and their metabolites. In the absence of a clear correlation with clinical

effects, the toxicological significance of these neurobiochemical findings is unclear.

Thus, this study found dose-related, statistically significant decreases in red blood cells at all

doses, but the significance of this finding is uncertain, because decreased hematocrit was observed

only at the high dose. Statistically significant decreases in antibody response were observed at the

mid dose, and these decreases reached 40% (a value often used by immunotoxicologists as a rule of

thumb for clinically relevant decreases) at the high dose. Identification of a NOAEL in this study is

somewhat problematic, because immunotoxicity risk assessment guidelines have not been developed. 

The determination of what degree of decrease is adverse is also problematic, because the

clinical relevance of a decrement in immune function will depend on the magnitude and type of
49

Table 6. Effects of Phenol Exposure on Spleen Cellularity and Selected Blood Parameters in Mice and Rats

Concentration

(mg/L)

Dose

(mg/kg-day)

Spleen

cellularity

(x 10-7)

WBCa

(x 10-3)

RBCa

(x 10-6)

Hematocrit

%

Differential counts as % of WBCs

Lymphocyte Neutrophil Monocyte

Hsieh et al. (1992) - 4-week study in CD-1 mice

0 0 8.59 ± 0.34b 6.06 ± 0.17 7.17 ± 0.56 48.00 ± 0.52 74.20 ± 1.83 17.00 ± 1.00 4.60 ± 0.51

4.7 1.8 7.94 ± 0.20 5.82 ± 0.60 4.90 ± 0.54c 49.10 ± 0.68 71.80 ± 2.06 19.40 ± 0.75 4.80 ± 1.02

19.5 6.3 7.31 ± 0.40 5.05 ± 0.53 4.64 ± 0.76c 48.20 ± 1.24 69.20 ± 3.25 21.80 ± 2.40 4.60 ± 0.81

95.2 33.6 7.26 ± 0.55 5.68 ± 0.69 3.23 ± 0.68c 44.10 ± 0.81c 73.60 ± 2.32 17.00 ± 1.55 6.20 ± 1.16

Historical control valued Not available 9.0 (8.9-9.1) 7.6 (7.2-8.0) 42 (36-48) 70 (52-86) 25 (10-42) 4 (0-8)

IIT Research Institute (1999); Ryan et al. (2001) - 2-generation study in Sprague-Dawley rats, effects in P1 generation

0 0 43.2±13.5e 13.1±2.01 9.22±0.37 46.5±1.44 N/A N/A N/A

200 15 38.3±8.78f 13.8±1.98 9.08±0.62 46.2±3.65 N/A N/A N/A

1000 71 48.1±11.0 14.5±2.42 9.03±0.34 46.4±1.56 N/A N/A N/A

5000 301 52.7±13.4 14.9±2.93 8.81±0.44 45.1±1.75 N/A N/A N/A

a

Cells/mm3 . bValues are given as mean ± S.E. (n=5); N/A = not assayed. cSignificant (P < 0.05) difference from the control value. dMean (Range ±2S.D) for mice 6-

8 weeks of age, based on 20 studies, from Charles River Laboratories (1986). e

Mean ±SD (n=9-10 for hematology endpoints and n=8-9 for spleen cellularity). 

f

Standard deviation reported in Ryan et al. (2001) has a typographical error; correct standard deviation obtained from IIT Research Institute (1999)
4

Although this unpublished study is not a primary reference for this assessment, it is presented

here because it contributes some useful information to the overall hazard identification phase of the

phenol assessment.

50

immune challenge, with a sufficiently large challenge resulting in illness even for unimpaired

individuals. In a paper on the use of immunotoxicity data for risk assessment, Selgrade (1999)

recommended that any statistically significant and consistent change be considered a risk for the

purposes of hazard identification, but the degree of change considered adverse for the purposes of

dose-response assessment was not addressed. 

On the basis of magnitude of the decreases in antibody response observed in three

related—assays, supported by decreased hematocrit and red blood cells—the high dose (33.6 mg/kgday) can be considered the study LOAEL, and the mid dose (6.2 mg/kg-day) can be considered the

study NOAEL. There is, however, considerable uncertainty regarding the reliability of these values

because of issues of study interpretation and because the study used only five animals per group as

opposed to the recommended eight per group (U.S. EPA, 1998c). 

The results of BMD modeling conducted for this study are presented in Appendix B for

completeness. However, it is unclear what the appropriate benchmark response (BMR) would be for

an in vivo/in vitro immunotoxicity study, and so the modeling results are considered very

preliminary. In particular, it is unclear whether the default of one standard deviation is appropriate

as the BMR for this study design in light of the small sample size (n = 5) but the relatively tight data.

In contrast to the minimal effects observed in these drinking water studies, gavage dosing

with phenol produces severe toxicity, including liver and kidney pathology, and death at doses that

cause only minimal effects when delivered in drinking water. 

Dow Chemical Co. (1945) administered 0, 50, or 100 mg/kg phenol by gavage 5 days/wk to

10 rats per group (sex and strain not reported) for 6 months (0, 35.7, or 71.4 mg/kg-day after

adjusting for intermittent dosing). The dosing volume was not reported. Mortality occurred in 1/10,

4/10, and 4/10 rats in the control, low-, and high-dose groups. The authors raised questions about

whether the mortality was treatment-related, but it is not clear whether they questioned whether the

deaths were due to phenol or to gavage accidents. Other observed effects were slight cloudy swelling

of the liver and of the tubular epithelium at the high dose and slight tubular degeneration at the low

dose. This unpublished study4

 is limited by the incomplete reporting of methods and results, but the

low dose of 35.7 mg/kg-day appears to be a LOAEL.
51

In a series of toxicological screening tests, the systemic, neurological, and developmental

effects of phenol in F344 rats following acute and short-term oral exposure were examined

(Narotsky and Kavlock, 1995; Berman et al., 1995; and Moser et al., 1995; MacPhail et al., 1995). 

In these tests, systemic and neurological effects were examined on the same animals following

exposure by gavage to a single dose of phenol or to 14 consecutive daily doses. Developmental

toxicity was also examined in pregnant rats that received phenol by gavage on GDs 6–19. The

dosing volume was 1 mL/kg (Moser et al., 1995; Narotsky and Kavlock, 1995).

In the acute toxicity study of this series, groups of eight female rats were given a single

gavage dose (1 mL/kg volume) of phenol at 0, 12, 40, 120, or 224 mg/kg in water (Berman et al.,

1995; Moser et al., 1995). An FOB evaluating autonomic and neuromuscular functions, activity,

excitability, and sensorimotor and physiological measures was conducted prior to the exposure and

at approximately 4 and 24 hours after exposure. Immediately after the 24-hour FOB, the animals

were sacrificed, blood samples were collected for serum chemistry analyses, and the liver, kidneys,

spleen, thymus, and adrenals were weighed and subjected to histopathological examinations. Two

rats (25%) died within 4 hours of exposure to 224 mg/kg, and one rat died 24 hours after exposure to

120 mg/kg phenol. The only treatment-related effects observed were confined to these two dose

groups, and they included tremor, decreased motor activity, and kidney pathology (necrosis, protein

casts, and papillary hemorrhage). Hepatocyte necrosis was also observed at 40 and 120 mg/kg but

not at 224 mg/kg. No other effects were reported at the lower doses, although the primary data were

not provided.

In the short-term study, groups of eight female rats were given daily gavage doses of phenol

in water at 0, 4, 12, 40, or 120 mg/kg-day for 14 consecutive days (Berman et al., 1995; Moser et al.,

1995). As in the acute study, the FOB was conducted prior to exposure as well as on days 4 and 9

(before the daily dose) and approximately 24 hours after the last dose. After the last FOB, blood

samples were collected for serum chemistry analyses, and internal organs were removed, weighed,

and subjected to histopathological examinations. All rats administered the high dose died during the

study, but deaths occurred over the entire dosing period. Tremor was also seen in the high-dose

(120 mg/kg) group immediately after the first administration but not after subsequent treatment. 

Vacuolar degeneration of the liver, kidney necrosis and protein casts, and “necrosis or atrophy of

spleen or thymus” were reported at 40 mg/kg-day. The increased incidences were not large enough

to be statistically significant; the statistical power of the study was also low, with only 8 rats per

group.

Additional information on this study is available from a preliminary abstract (Schlicht et al.,

1992) and from a recent WHO (1994) review. According to these sources, the renal pathology
52

consisted of 3/8 rats with renal vascular stasis, 2/8 rats with tubular degeneration in the papillar

region, and 1/8 rats with protein casts in the tubules. WHO (1994) states that, according to a

personal communication from one of the study authors, the pathology report attributed the renal

findings to decreased vascular perfusion. 

The study also found slight but not statistically significant decreases in motor activity at 40

mg/kg-day. The only statistically significant effect in this group was increased rearing in the postexposure measurement. The only effect at 12 mg/kg-day was “necrosis or atrophy in the spleen or

thymus” in 1/8 rats. On the basis of the liver, kidney, and thymus/spleen pathology findings, which

are rarely observed in control animals in 2-week studies, and the decreased motor activity, the

second dose (40 mg/kg-day) was the study LOAEL, and the mid dose of 12 mg/kg-day was the study

NOAEL.

4.2.2. Inhalation

The laboratory animal inhalation data for phenol are very limited, with only one 2-week

toxicity study being conducted using modern methodology and documentation (Hoffman et al.,

2001; additional details available in the unpublished version, Huntingdon, 1998). Although a

subchronic study conducted with multiple laboratory animal species is available (Sandage, 1961),

this latter unpublished study tested only one concentration and was insufficiently documented for

definitive risk assessment purposes. Other short-term (Dalin and Kristoffersson, 1974) or

subchronic (Deichmann et al., 1944) inhalation toxicity studies are limited by short duration,

inadequate documentation, or lack of a modern exposure protocol. Nonetheless, the data are

consistent that the respiratory tract, kidney, and nervous system are targets of inhalation exposures. 

In conducting dosimetric conversions from animal studies to human exposure scenarios, U.S.

EPA (1994b) classifies gases according to their water solubility and reactivity. Category 1 gases are

highly water-soluble and/or rapidly reactive and do not penetrate the blood. Category 3 gases are

water insoluble, and uptake from the lungs is limited by perfusion. Category 2 gases are

intermediate between these two groups. They are moderately water-soluble and rapidly-reversibly

reactive or moderately-to-slowly irreversibly metabolized in the respiratory tissue. On the basis of

phenol’s chemical/physical properties (see Table 1) of moderate water solubility and moderate

reactivity (based on the evidence of irritation and corrosivity seen following direct contact), it can be

considered a Category 2 gas. This conclusion is supported by the finding of both respiratory effects

(from direct contact) and systemic (extrarespiratory) effects (from absorbed phenol) following

inhalation exposure, as described below.
53

Because the equations for the regional gas dosimetry ratio (RGDR) for Category 2 gases are

currently undergoing EPA reevaluation (eqs. 4-29 through 4-44, pages 4-52 through 4-57 of U.S.

EPA, 1994b), dosimetric adjustments for extrarespiratory effects were made using the Category 3

equations (eq. 4-48, page 4-60 of U.S. EPA, 1994b), in which the RGDR is based on the blood:air

partition coefficient for the chemical in the experimental animal species and in humans. No data on

the blood:air partition coefficient for phenol in laboratory animals or in humans were located. 

Therefore, the default value of 1 for the ratio of the laboratory animal-to-human partition coefficient

was used, and the human equivalent concentration (HEC) for systemic effects was the same as the

duration-adjusted concentration.

HECs for respiratory tract effects were calculated using the equations of U.S. EPA (1994b)

for a Category 1 gas (eqs. 4-17 through 4-28, pages 4-47 through 4-51). When the EPA reanalysis is

complete, revised dosimetric conversions may be calculated. The inhalation toxicity data for phenol

are summarized in Table 7.

The acute toxicity studies support the findings of the short-term and subchronic studies that

the respiratory tract and nervous system are targets of inhaled phenol. For example, tremors were

seen in rats and guinea pigs exposed to 187 or 540 ppm (720 or 2080 mg/m3

) phenol for 30 minutes

in a whole-body inhalation chamber (UBTL, 1991). By contrast, no tremors were observed in rats

exposed via nose-only inhalation at 25 ppm (96 mg/m3

) for 6 hours (Dow Chemical Co., 1994). 

Phenol also caused sensory irritation in mice, as evidenced by decreased respiratory rate (De

Ceaurriz et al., 1981). The concentration associated with a 50% decrease in rate (RD50) was

estimated to be 166 ppm (639 mg/m3). No acute lethality studies were identified for phenol

following exposure by the inhalation route.

In a 2-week inhalation study conducted according to GLP guidelines (Hoffman et al., 2001;

full study report available as Huntingdon, 1998), groups of 20 F344 rats per sex were exposed noseonly to actual concentrations of 0, 0.52, 4.9, or 25 ppm phenol (0, 2.0, 18.9, or 96.2 mg/m3

) 6

hrs/day, 5 days/wk for 2 weeks. The duration-adjusted concentrations were 0, 0.36, 3.4, and 17

mg/m3, respectively. The animals were observed twice daily for mortality and abnormal clinical

signs. Animal body weights and food consumption were recorded twice pre-test, weekly thereafter,

and just prior to sacrifice. At the end of 2 weeks of exposure, 10 rats of each sex in each group were

sacrificed. The rest of the rats were sacrificed after 2 weeks of recovery. Blood samples were

collected just prior to sacrifice for hematological (including differential leukocyte count) and
54

Table 7. Summary of Inhalation Toxicity Studies

Strain

Species,

Sex

Reference Exposure

mg/m3

Duration DurationAdjusted

mg/m3

Response NOAEL/

LOAEL

mg/m3

NOAEL/

LOAEL (HEC) 

mg/m3

Comments

F344 Rat

(20/sex)

Huntingdon

1998

0, 2.0, 18.9,

96.2

Nose-only

6 hr/d

5 d/week 

2 weeks

0, 0.36, 3.4, 17 Red nasal discharge, but no

histopathology lesions.

17/None 2.5 or 17

/None

Well-conducted study, but

authors did not note the

clinical signs

Rat, 7

exposed,

11-12

controls

(Strain &

sex NS)

Dalin and

Kristoffersson

1974

0, 100 15 days

continuous

0, 100 Nervous system effects,

increased serum liver

enzymes

None/100 None/100 Exposure measurement not

done according to modern

methods, no histopathology

exam

Guinea pig,

12

(Strain &

sex NS)

Deichmann et

al. 1944

100-200 7 hr/day

5 d/week

6 weeks

31 (based on

midpoint of

range)

FEL – 5/12 dead None None/

31 is FEL

Minimal documentation, 

outdated exposure methods,

no controls 

Rabbit

6 exposed

(Strain and

sex NS)

Deichmann et

al. 1944

100-200 7 hr/day

5 d/week

13 weeks 

31 (based on

midpoint of

range)

Pneumonia, heart

inflammation, liver necrosis,

kidney tubular degeneration

None None/

31

Minimal documentation, 

outdated exposure methods,

no controls

Rat, 12

(Strain and

sex NS)

Deichmann et

al. 1944

100-200 7 hr/day

5 d/week

74 days

31 (based on

midpoint of

range)

No effect, no evidence of

histopathology

31/None 31/None Minimal documentation,

sensitivity of assay unclear,

outdated exposure methods,

no controls

Rhesus

monkey

10 M/

group

Sandage 1961 0, 18.2 90 days,

continuous

0, 18.2 Liver and kidney pathology

(Not further described)

None/18.2 None/18.2 Pathology reported to be

minimal, but limited by

minimal description

Unpublished study
Strain

Species,

Sex

Reference Exposure

mg/m3

Duration DurationAdjusted

mg/m3

Response NOAEL/

LOAEL

mg/m3

NOAEL/

LOAEL (HEC) 

mg/m3

Comments

55

SpragueDawley rat

50 M/

group

Sandage 1961 0, 18.2 90 days,

continuous

0, 18.2 Liver and kidney pathology

(Not further described)

None/18.2 None/18.2 Pathology reported to be

minimal, but limited by

minimal description

Unpublished study

Albino

mouse 100

M/group

Sandage 1961 0, 18.2 90 days,

continuous

0, 18.2 Lung pathology

(Not further described)

None/18.2 Not

determinable*

Pathology reported to be

minimal, but limited by

minimal description

Unpublished study

NS = Not stated

* HEC cannot be determined because the region of the respiratory tract affected is not clear.
56

biochemical examinations. Gross pathological evaluations were conducted on all of the animals,

and organ weights were determined. Histopathological examinations were conducted on the liver,

kidney, and respiratory tract tissues (including three sections of the lungs with mainstem bronchi, the

pharynx, and three sections of the nasal turbinates) of the control and high-exposure groups; the

spleen of mid-concentration females was also analyzed. 

During the exposure, one male rat in the low-concentration group accidentally died from

trauma caused by turning itself within the nose-only restraint tube. All of the other rats survived

until sacrifice at the end of the 2-week exposure or 2-week recovery periods. During exposure and

recovery, there were no treatment-related changes in weekly physical examinations, body weight,

weight gain, or food consumption. The authors reported no effect on clinical signs. However, there

was a concentration- and duration-related increase in the incidence of a red nasal discharge in the

males. The incidence was 0/20, 0/20, 3/20, and 4/20 at 0, 0.52, 4.9, and 25 ppm, respectively, in the

first week and 0/20, 0/20, 7/20, and 10/20 in the second week of exposure. This detailed information

on nasal discharge was presented only in the unpublished report (Huntingdon, 1998); the published

version reported scattered observations of chromodacryorrhea and nasal discharge. 

In an analysis done for this assessment, the incidence at the mid and high concentrations was

statistically significant, using the Fisher exact test. In females, nasal discharge was seen in 1/20 at

the low concentration and 3/20 at the mid concentration in the second week, but no discharge was

reported in high-concentration females. Prior to exposure, a nasal discharge was observed in a

single control male and a single high-concentration female. Considering the exposure chamber

design, it does not appear that the discharge was an artifact of the rats’ noses being in contact with

phenol condensate on the chamber walls. Instead, in the absence of nasal histopathology, it is likely

that the discharge reflected a nonspecific response to stress in the rats. A tear-like nasal discharge in

rats can be a generalized response to stress from a variety of causes. Porphyrins in the discharge

lead to a red color. In light of the dose-related response in males, it appears that the stress in this

study was related to an effect of phenol, either as an irritant or a toxicant.

Hematological and biochemical examinations showed slight but statistically significant

increases in prothrombin time at the low concentration only and in albumin concentration in highexposure females; these changes were not considered to be biologically significant. No other

significant changes in hematology or biochemistry were observed. The only statistically significant

changes in organ weights were decreases in liver-to-body, spleen-to-body and spleen-to-brain weight

ratios in mid-concentration (18.9 mg/m3

) females. Because the changes in organ weights did not

occur at the highest phenol exposure concentration (96.2 mg/m3), and the same responses occurred in
57

female rats but not in male rats, these responses were considered by the authors to be not

toxicologically significant.

Gross pathological and microscopic examinations of these organs did not exhibit any

differences from those of the controls. Although there were a number of histopathology findings in

the respiratory tract (e.g., inflammatory cells in the nasolacrimal ducts, alveolar macrophages, and

eosinophilic and basophilic material), these findings occurred at similar incidences in the control and

exposed groups. The lesions reported were also those typically seen in control animals. The only

lesion of concern was minimal to slight lung hemorrhage, which was reported in 4/10 control males

and 6/10 high-exposure males at the terminal sacrifice. However, there was no clear concentrationrelated increase in incidence or severity, this lesion was not found in the females, and this lesion was

not seen in exposed animals post-recovery or in control animals.

Thus, it appears that the only effect of concern in this study was the red nasal discharge,

which was observed in males but not in females, and this effect was probably due to a nonspecific

response to stress. In addition, no supporting histopathology was observed in a thorough

examination. On the basis of these considerations, the highest concentration (96.2 mg/m3

) in this

study was NOAEL. The HEC cannot be definitively determined in the absence of an affected

endpoint. However, assuming that the respiratory tract would be affected first, as shown in other

studies, a conservative NOAEL (HEC) based on a nasal effect would be 2.5 mg/m3

. Assuming that

the nasal discharge reflects a nonadverse systemic stress response, the NOAEL(HEC) would be 17

mg/m3

. No LOAEL was identified.

Deichmann et al. (1944) conducted subchronic inhalation studies of phenol toxicity in rabbits,

rats, and guinea pigs. Twelve guinea pigs, 6 rabbits, and 12 rats (strain and sex not reported) were

exposed (whole body) in a single exposure chamber to phenol vapor at “a concentration ranging from

0.1 to 0.2 mg/L (100-200 mg/m3)” for 7 hrs/day, 5 days/wk for 6 weeks, 13 weeks, or approximately

11 weeks, respectively. The actual exposure concentration apparently could not be controlled more

precisely. Using the midpoint of 150 mg/m3 as the exposure concentration, the duration-adjusted

concentration was 31 mg/m3.

Among the three tested species, the guinea pig was the most sensitive and the rat was the least

sensitive to phenol exposure. Deaths were observed in 5/12 guinea pigs during the 6-week exposure

period. Other signs of toxicity in the guinea pigs included decreased activity during the first week

and respiratory difficulties and paralysis of hind quarters after 4 weeks of exposure. 

Histopathological evaluations revealed lesions of the lungs (pneumonia and bronchitis), heart

(inflammation, fibrosis, and necrosis), liver (fatty changes and necrosis), and kidneys (tubular
5

Although this unpublished study is not a primary reference for this assessment, it is presented

here because it contributes some useful information to the overall hazard identification phase of the

phenol assessment.

58

degeneration and edema). At the end of exposure, the surviving guinea pigs had a concentration of

1.0 mg free phenol/100 mL blood and 0.4 mg conjugated phenol/100 mL blood. (Details on the

analytical procedures used to measure phenol in blood were not reported.). 

No deaths or clinical signs of toxicity were observed in the rabbits exposed for 13 weeks, but

lobular pneumonia and fibrosis was observed in these animals. Histopathology lesions in the heart,

kidney, and liver were similar to, but less severe than, those reported in the guinea pigs. After 37

days on study, the rabbits had a concentration of 0.5 mg free phenol/100 mL blood and 0.7 mg

conjugated phenol/100 mL blood; similar concentrations were observed at the end of the exposure

period.

The rats did not show any clinical signs of toxicity during the 74-day exposure period

(approximately 11 weeks), and there was no histopathological evidence of any effect. Blood phenol

levels were not reported for the rats, but an analysis of carcass homogenate found 0.2 mg free phenol

and 0.35 mg conjugated phenol per 100 g carcass homogenate. These levels were reported to be

within the normal range in unexposed rats. This study is limited by the use of only one exposure

concentration, the absence of controls, the inadequate control of exposure levels, and the absence of

reporting of the primary data. However, the results do appear to show that rats are much less

sensitive than rabbits or guinea pigs to the inhalation effects of phenol. However, a comparison of

blood levels suggests that the interspecies differences are due to both toxicokinetic and

toxicodynamic differences.

In an unpublished 90-day study (Sandage, 1961), groups of 10 male rhesus monkeys, 50 male

Sprague-Dawley rats, and 100 male albino mice were exposed to average phenol concentrations of 0

or 4.72 ppm (18.2 mg/m3

) continuously for 90 days.5 Exposure was interrupted for 14 hours on day

39 and for 36 hours on days 68–69. The frequency of monitoring of the test atmosphere was not

reported, but the phenol concentration was reported to remain in the desired range of 4.5–5.5 ppm

“after the first three days.” No further information on the concentrations during the first 3 days was

reported.

During the exposure, no deaths were observed in the test animals. Body weight gain in mice

was comparable to that in controls but was slightly higher in exposed rats and monkeys. A complete

hematological examination showed no significant changes in the three test species following phenol
59

exposure. Blood biochemistry (alkaline phosphatase, cholinesterase, amylase, lipase, and glutamic

oxalacetic transaminase) was evaluated in monkeys only. Urinalysis was apparently conducted in all

species, but kidney function tests (urine volume and specific gravity) were conducted only in

monkeys and rats. The study authors reported that there were no effects on any of these endpoints

but did not provide any supporting data.

At the end of the exposure period, “approximately half” of the animals underwent a stress test

in which the animals swam in a smooth-walled tank until exhausted. These animals were sacrificed

immediately after the test, and the other animals were held for a 2-week recovery period prior to

sacrifice. Histopathological evaluations were conducted on only 5–8 organs (including the liver,

kidney, and lung). It appears that all of the monkeys and about half of the rats and mice were

evaluated, although it is not clear whether some of the rodents were evaluated after the recovery

period.

The authors considered the histopathology findings “essentially negative” and did not provide

any description of the observed lesions or the number of animals examined histopathologically. 

Liver and kidney pathology was observed in 30% and 20%, respectively, of the monkeys (compared

with 0% of the controls). However, the authors did not consider these changes to be significant, and

they noted that 6/7 reports of pathology in monkeys were considered “minimal or doubtful.” Liver

and kidney pathology was also reported in 20% of phenol-exposed rats (compared with 0% of the

controls) and lung pathology was reported in 20% of the phenol-exposed mice (compared with 6% of

the controls). The incidences of liver and kidney pathology in the rat and lung pathology in the

mouse were statistically significant in a Fisher’s exact test done for this assessment. Although the

incidence of lung pathology was not affected in monkeys and rats, a relatively high incidence of lung

pathology in the control animals (30% and 65%, respectively) decreased the sensitivity of the

evaluation. No other significant pathological changes were reported in the test animals. 

Although the authors concluded that there was no evidence that phenol exposure resulted in

significant damage, there is some indication of liver, kidney, and lung pathology in this study, but the

inadequate reporting precludes the determination of whether there was a treatment-related effect. For

the purposes of this assessment, the single exposure level tested, 18.2 mg/m3, should be considered a

free-standing LOAEL, although it might be considered a minimal LOAEL if additional

histopathology data were available. The LOAEL (HEC) for the kidney and liver lesions is also 18.2

mg/m3

. In the absence of additional information on the nature of the lung lesions, the LOAEL (HEC)

for the lung cannot be determined. The study is also limited by the poor control of exposure levels

and limited reporting of effects. 
60

In a study published by Dalin and Kristoffersson (1974), seven white rats of an in-house strain

were exposed to phenol at a concentration of 100 mg/m3

 continuously for 15 days. There is some

uncertainty about this exposure measure, however, because the exposure chamber was not set up

according to modern designs, and it does not appear that continuous monitoring of exposure levels

was conducted. Unexposed rats (n = 11–12) were used as controls. Nervous system effects were

observed from the first day after the start of exposure. These effects progressed from increased

activity to imbalance, twitches, and disordered walking rhythm on days 3–4. These signs disappeared

by day 5 and were replaced by sluggish behavior until the end of the exposure. A tilting-plane test

was conducted before and after exposure in both groups, and a significant effect was observed on the

exposed rats. There were no significant changes in food intake or water consumption during the

exposure period. Although there was no significant difference in body weight of the exposed group

compared with that of the controls, the average body weight of the exposed group decreased during

exposure, whereas the controls gained weight.

The serum biochemical evaluations showed large, statistically significant increases in SGOT,

SGPT, lactic dehydroganese (LDH), and glutamate dehydrogenase activities, indicating liver damage. 

Plasma potassium and magnesium levels were also increased. Although the significance of these

changes is unknown, the authors suggested that the increased magnesium levels may have caused

some of the nervous system effects. Hemoglobin and hematocrit were unaffected. No histopathology

examination was conducted. On the basis of the observed nervous system effects as well as the

serum enzyme changes indicating liver damage, the only exposure concentration was a free-standing

LOAEL. The LOAEL (HEC) is 100 mg/m3

, but the actual exposure measurement is of low quality.

4.2.3. Dermal

Phenol is quite irritating, and dermal exposure to liquid phenol or to concentrated phenol

vapor can result in inflammation and necrosis of the skin (Conning and Hayes, 1970; Patrick et al.,

1985; Pullin et al., 1978). As discussed in Section 3.1, phenol is readily absorbed from dermal

contact with phenol liquid or phenol vapor, so systemic effects can also result from dermal exposure. 

Several acute lethality assays have been reported. Conning and Hayes (1970) reported a dermal LD50

of 669.4 mg/kg for undiluted phenol applied for 24 hours to the skin of female Alderly Park rats. 

Acute dermal toxicity appears to be dependent on the concentration of phenol, with increased

lethality observed with decreased concentration when the same total dose is applied (Deichmann and

Witherup, 1944; Conning and Hayes, 1970). In addition to lethality, renal effects (severe
61

hemoglobinuria and hematin casts in the tubules), cardiovascular effects (cardiac arrhythmias and

ventricular tachycardia), and neurological effects (severe muscle tremors, marked twitching,

generalized convulsions, loss of consciousness, and prostration) were observed at 107.1 mg/kg in

female Alderly Park rats following dermal exposure to undiluted phenol for 24 hours (Conning and

Hayes, 1970). A similar array of effects have been reported in humans following accidental dermal

exposures to large volumes of phenol (ATSDR 1998). 

4.3. REPRODUCTIVE/DEVELOPMENTAL STUDIES

No studies of the reproductive or developmental toxicity of phenol following inhalation

exposure of laboratory animals were located. Several developmental toxicity studies in rats and mice

conducted via the gavage route are available (Argus Research Laboratories, 1997; NTP, 1983a,b;

Narotsky and Kavlock, 1995); the only developmental effect reported in these studies was decreased

fetal body weight. In a two-generation drinking water study in rats (Ryan et al., 2001; available in

unpublished form as IIT Research Institute, 1999), decreased pup weight and decreased survival of

pups pre-culling were observed, but these effects appeared to be secondary to decreased water

consumption.

In the Ryan et al study, 30 Sprague-Dawley rats/sex/group were exposed to 0, 200, 1000 or

5000 ppm phenol in drinking water. Parental (P1) rats were given phenol for 10 weeks prior to

mating, during a 2-week mating period, throughout gestation, lactation, and until sacrifice. The

males were sacrificed after successful mating. All of the P1 females were allowed natural parturition

and were sacrificed at F1 weaning. The authors calculated that the average daily phenol intake

during week 10 was 0, 14.7, 70.9, and 301.0 mg/kg-day for P1 males and 0, 20.0, 93.0, and 320.5

mg/kg-day for P1 females. 

For the F1 generation, the average phenol intake during week 10 was 0, 13.5, 69.8, and 319.1

mg/kg-day for males and 0, 20.9, 93.8, and 379.5 mg/kg-day for females. The F1 generation (20

rats/sex/group) was treated following a protocol similar to that used for the P1 generation, and F2

pups were sacrificed after weaning, on postnatal day (PND) 22. During treatment, rats were

monitored for mortality, clinical signs, body weight, and food and water consumption. At sacrifice,

the animals were necropsied, and reproductive organs from 20 animals per sex in the control and

high-dose groups from the P1 and F1 generations were examined microscopically. In addition, the

spleen, thymus, liver, and kidneys from 10 randomly selected P1 and F1 animals of each sex in the

control and high-dose groups were examined.
62

Most of the treatment-related changes in P1 rats were observed in the high-dose groups. No

treatment-related mortality was observed in P1 rats of either sex. Three high-dose F1 female pups

died shortly after weaning. The deaths appeared to be associated with decreased water consumption

associated with poor palatability, as these pups refused to drink the water containing the phenol. No

other treatment-related mortality was reported in the F1 generation. The only significant observed

clinical sign was redness around the nose fur, which occurred in the high-dose males and females of

the F1 generation before mating and in P1 dams during lactation. As noted in the context of the

Hoffman et al. (2001) inhalation study, this redness likely reflects a nonspecific stress response. A

significant decrease in water consumption was observed throughout the study in P1 animals of both

sexes (up to 23% for males and up to 39% for females) and was attributed to poor palatability. 

Decreased water consumption in the F1 rats at the high dose was of a similar or larger magnitude. 

The low water consumption at the high dose was accompanied by decreased body weights as

compared with the controls (9% in P1 males and 16% in P1 females at sacrifice). 

At birth, the high-dose F1 and F2 pups had decreased body weights as compared with the

controls, and the differences were larger by PND 21. The authors noted that pups began drinking the

water prior to weaning and that the decreased water consumption was also observed in the high-dose

pups. Decreased absolute organ weights and increased relative organ weights were observed for a

number of organs at the high dose in both the P1 and F1 generations. Most of these changes likely

reflected the lower body weight and overall dehydration in these groups.

F1 females had a statistically significant, dose-related decrease in uterine weights at all doses,

but P1 females were not affected. The authors suggested that some of this decrease may have been

related to a lower incidence of uterine dilatation at the high dose. Because the stage in the estrus

cycle can affect uterine weight, estrus cycle staging was also considered as an explanation, but the

authors did not consider the small decreases in the number of rats in estrus (16/24, 15/24, 13/25, 9/22

in the control, low-, mid-, and high-dose groups) sufficient to account for the decreased uterine

weight. Nonetheless, the decreased uterine weight was not considered adverse for several reasons. 

There was no evidence of a dose-response relationship for relative uterine weight across the three

dose groups (Table 8). There was no effect on reproductive function and no histopathological

changes in the uterus. Finally, the uterine weight was below the control range for only a few rats in

each dose group, and the control group appeared to have greater variability (particularly at the high

end) than each of the experimental groups. 
63

Table 8. Selected Results of Two-Generation Drinking Water Study 

(Ryan et al., 2001; IIT Research Institute, 1999)

Endpoint Control 200 ppm

(14-21 mg/kgday)

1000 ppm

(70-94 mg/kgday)

5000 ppm

(301-380

mg/kg-day)

P1 body weight - males week 10

(g)

414±39.8 434±34.4 423±40.3 382±40.8*

P1 body weight - females week

10 (g)

259±24.5 260±21.8 258±19.3 234±21.2*

P1 water consumption - males

week 10 (g/day)

30±4.6 32±4.6 30±3.3 23±3.1*

P1 water consumption - females

week 10 (g/day)

24±3.7 26±5.5 24±4.1 15±2.1*

F1Absolute uterine weight (g) 

Uterine weight relative to body

weight

0.81±0.33

0.27±0.11

0.62±0.13*

0.21±0.044*

0.60±0.17*

0.20±0.058*

0.53±0.17*

0.20±0.068*

F1 Absolute prostate weight (g)

Prostate weight relative to body

weight

0.90±0.20

0.18±0.039

0.77±0.17*

0.15±0.031

0.74±0.17*

0.15±0.027*

0.76±0.16*

0.18±0.036

F1 pnd0 litter weight (g) 6.73±0.42 6.85±0.51 6.63±0.51 6.38±0.27*

F2 pnd0 litter weight (g) 6.67±0.38 6.80±0.44 6.58±0.42 6.20±0.48*

Preputial separation (age, days)

Body weight at onset (g)

44.0±2.16

229±23.0

45.4±2.05

236±23.4

44.4±2.022

227±19.7

47.8±3.13*

195±23.6*

Vaginal patency (age, days)

Body weight at onset (g)

34.6±1.85

126±13.7

34.5±1.83

127±13.5

34.1±2.25

121±12.6

38.3±2.21*

112±7.8*

1

Mean ±Standard deviation

2

Standard deviation reported in Ryan et al. (2001) has a typographical error; correct standard deviation obtained from IIT

Research Institute (1999)

*Statistically significant, p 0.05
64

Absolute prostate weights were significantly reduced—by approximately 15%—in all dose

groups of the F1 generation but not the P1 generation. Relative prostate weight was decreased in the

low-dose (but not statistically significantly) and mid-dose F1 groups but not at the high dose. In the

absence of a dose-response, the changes in prostate weight were not considered adverse. Doserelated statistically significant decreases in absolute adrenal and spleen weights were also observed in

the mid- and high-dose F1 males. However, there was no effect on relative weights of these organs in

these groups, there was no effect on these organ weights at any dose in the P1 males, and P1 females

were only affected at the high dose (where decreased body weight would have played a role).

The pathological examinations showed no treatment-related lesions in the kidneys, spleen,

liver, thymus, or reproductive organs. An immunotoxicity screen conducted with 10 male P1 rats per

group found no significant effects on spleen weight, cellularity, or antibody-forming cells for any test

group when compared with the control group; the expected results were found with a positive control

group. The strength of this finding is reduced, however, by the relatively large standard variability in

response (based on the standard error as a percent of the mean). 

Complete hematological evaluations (including hematocrit, erythrocyte count, and differential

white cell count) and serum biochemical evaluations were conducted on 8–10 P1 males/group prior to

sacrifice. The only significant change in these evaluations was increased blood urea nitrogen (BUN)

in the high-dose group. Because this change was not accompanied by increased creatinine and there

was no associated kidney pathology, the BUN increase was not considered to be biologically

significant. The authors also noted that all but one of the high-dose BUN values were within the

control range. 

There was no effect on fecundity or fertility in either generation. In addition, there was no

effect on other indicators of reproductive toxicity, including the frequency of estrus, testicular sperm

count, sperm motility, and sperm morphology. The survival of the high-dose F1 pups was

significantly decreased on PND 4 (pre-culling), although there was no effect on overall F1 pup

survival. In the F2 generation, high-dose pup survival was significantly decreased throughout the

lactation period. This decreased survival of both generations of pups was likely secondary to the

decreased maternal water intake and associated decreases in milk production. In the F1 generation,

delayed vaginal patency and delayed preputial separation were observed at the high dose. The delay

was considered secondary to decreased fetal growth at the high dose resulting from decreased water

consumption due to poor palatability and associated decreased food consumption.

Thus, all of the adverse systemic and reproductive effects of phenol in this study occurred at

the high dose, and they appear to be secondary to decreased water consumption due to poor
6

This study has not been peer-reviewed, but it was conducted (with minor deviations) according

to EPA guidelines for developmental toxicity studies, it is well documented, and it contributes useful

information to the hazard identification and dose-response portions of the assessment. The study was

designed to meet the U.S. EPA Pesticide Assessment Guidelines, Subdivision F, 83-3.

65

palatability rather than to a toxic effect of phenol. On the basis of the decreased parental and pup

body weight (compared to the controls) and decreased pup survival, the high dose is a LOAEL. The

study NOAEL is 70.9 mg/kg-day (based on the NOAEL corresponding to the lowest LOAEL in this

study, in P1 males). BMD modeling was not conducted for this study because the observed effects

appeared to be secondary to decreased water consumption and not reflective of phenol toxicity.

In an unpublished developmental toxicity study conducted according to GLP guidelines

(Argus Research Laboratories, 1997), pregnant Crl:CDRBR VAF/Plus Sprague-Dawley rats (25 per

group) received phenol by oral gavage on GDs 6 through 15.6 Dosing was three times daily with 0,

20, 40, or 120 mg phenol/kg/dosage using a dosing volume of 10 mL/kg. The corresponding daily

doses were 0, 60, 120, and 360 mg/kg-day. The authors noted that the test material was 90% phenol

United States Pharmacopeia (USP); the authors adjusted the dosage calculations for test material

purity.

The exposed dams were observed twice a day for viability and daily for clinical signs,

abortions, and premature deliveries. In addition, the maternal body weights were recorded every day,

and food consumption was also recorded periodically (every 1–2 days). The rats were sacrificed on

GD 20, and gross necropsy of the thoracic, abdominal, and pelvic viscera was performed. The

number of corpora lutea in each ovary was recorded. The uterus of each rat was excised and

examined for number and distribution of implantations, live and dead fetuses, and early and late

resorptions. Each fetus was weighed, sexed, and examined for gross external alterations. One-half of

the fetuses were examined for soft tissue alterations, and the rest were examined for skeletal

alterations.

One high-dose dam died on GD 11. The study authors attributed this death to phenol

treatment because it occurred only at the high dose, although there were no adverse clinical

observations and no abnormal necropsy findings in this animal. Other high-dose animals exhibited

excess salivation and tachypnea (rapid breathing). There were no other treatment-related clinical

observations and no treatment-related necropsy findings. Dose-dependent decreases in body weight

of the exposed animals as compared with the controls were observed. Statistically significant

decreases in both maternal body weight (8%) and body weight gain (38% for GD 6–16) were

observed at the high dose; although a statistically significant decrease in body weight gain (11%) was
66

observed at the mid dose, the decrease at the mid dose (relative to controls) in absolute maternal

weight at the end of dosing (3%) was not statistically significant. Dose-dependent decreases in food

consumption were also observed during the dosing period (see Table 9). 

Fetal body weights in the high-dose group were significantly lower than those of the controls,

by 5–7%. The high-dose group had a statistically significant decrease in ossification sites on the

hindlimb metatarsals, but it is unlikely that this small change is biologically significant. The

incidence of litters with incompletely ossified or unossified sternal centra was 0/23, 0/25, 3/23, and

3/24; this increase was not statistically significant (Table 9). There were small, dose-related

increases in the number of litters with fetuses with “any alteration” and with “any variation” at 120

mg/kg/day and higher. However, neither of these changes was statistically significant, and the

response was not clearly dose-related. In addition, an increase in total variations is of questionable

significance in the absence of any increase in individual variations. There were no other treatmentrelated effects on uterine contents, malformations, or variations.

The maternal NOAEL was 60 mg/kg-day, based on small decreases in maternal body weight

gain at 120 mg/kg-day, and the developmental NOAEL was 120 mg/kg-day, based on decreased fetal

body weight and delayed ossification at 360 mg/kg-day. Benchmark dose (BMD) modeling was also

conducted for the decreased maternal weight. Defining the BMR as a one standard deviation

decrease in maternal body weight gain, the BMDL was 93 mg/kg-day. Details on the BMD modeling

are provided in Appendix B.

In a well-designed developmental toxicity study (NTP, 1983a), timed-mated CD rats were

administered phenol by gavage at 0, 30, 60, or 120 mg/kg-day in 5 mL/kg distilled water on GDs 6 to

15 and sacrificed on GD 20. Females were weighed on GDs 0, 6–15 (prior to daily dosing), and 20

(immediately following sacrifice), and they were also observed during treatment for clinical signs of

toxicity. A total of 20–22 females per group were confirmed to be pregnant at sacrifice on GD 20. 

The dams were evaluated at sacrifice for body weight, liver weight, gravid uterine weight, and status

of uterine implantation sites. Live fetuses were weighed, sexed, and examined for gross

morphological abnormalities and malformations in the viscera and skeleton. 

The results of this study did not show any dose-related signs of maternal toxicity or any

clinical symptoms of toxicity related to phenol treatment. The number of implantation sites was

slightly higher in the dosed groups, but this change could not be treatment related because

implantations in this strain
67

Table 9. Key Results in Argus Research Laboratories, (1997) Rat Developmental Toxicity

Study

Control 60 mg/kg-day 120 mg/kg/day 360 mg/kg-day

Maternal body weight 

GD 16 (g)

363.9±22.2 359.0±18.7 354.3±17.5 334.2±20.1**

Maternal body weight gain

(GD 6-16) (g)

64.0±10.7 58.0±9.4 56.8±10.8* 39.8±9.5**

Maternal food

consumption (GD 6-16)

(g/day)

26.7±2.7 26.0±2.6 24.8±1.9** 21.9±2.1**

Fetal body weight - all

fetuses (g)

3.62±0.30 3.63±0.24 3.60±0.30 3.41±0.35*

Fetal body weight - males

(g)

3.71±0.30 3.73±0.25 3.71±0.31 3.53±0.34

Fetal body weight -

females (g)

3.53±0.30 3.53±0.25 3.49±0.28 3.29±0.35**

% Fetuses with any

alteration/litter

1.0±2.6 2.4±4.6 3.7±4.4 4.1±7.8

% Fetuses with any

variation/litter

1.0±2.6 2.4±4.6 3.4±4.4 3.8±7.3

Incompletely ossified or

unossified sternal centra

litter incidence

0 0 37337 37338

Ossification sites per fetus

per litter, hindlimb

metatarsals

4.00±0.00 4.00±0.00 4.00±0.00 3.98±0.07*

1

Mean ±Standard deviation

*Statistically significant, p 0.05

**Statistically significant, p 0.01
68

take place prior to GD 6 (prior to dosing). Significant increases in the litters with nonlive (dead plus

resorbed) pups were observed in the low- and mid-dose groups but not in the high-dose group (Table

10). Because this response was not dose dependent, and the response in the high-dose group was

comparable to that in the control group, this observation is not considered treatment-related. In

addition, there was no effect on the more appropriate measure of nonlive pups per litter. 

There was also no effect on live fetuses, sex ratio, malformations, or variations, but there was

a clear dose-related downward trend in fetal body weight, although the changes at the two lower

doses were small and the effect was statistically significant only at the high dose (Table 10). Fetal

body weights in the high-dose group were 93% of the average of those in the control group; fetal

body weights were not reported separately for males and females. Historical control data from the

supplier report the average fetal body weight in this strain as being well below the weight in the highdose group (Charles River Laboratories, 1988). (Concurrent control weight was 4.14 g, high-dose

weight was 3.84 g, and historical control weight was 3.39 g.) The litter size in the high-dose group

was also somewhat higher (but not statistically significantly) than in the controls, possibly

contributing to the smaller fetal weight at the high dose. 

As shown in Table 10, the total pup burden (total fetal weight) and the gravid uterine weight

were highest in the low-dose group and then in the high-dose group; both values higher than those in

the control group. In addition, the treatment-period maternal weight gain was very similar in the

control and high-dose groups (but higher in the low-dose group), but the absolute maternal weight

gain (i.e., adjusted for the gravid uterine weight) was much lower in the high-dose group than in the

controls. The results from the low-dose group suggest that the dams could have borne a somewhat

higher burden of the total in utero package. However, the results also suggest that the dams were

near the limit of what they could carry, considering the lower absolute weight gain but unaffected

treatment-period weight gain in the high-dose group. No dose-related signs of maternal toxicity and

no clinical symptoms of toxicity related to phenol treatment were observed in this study. 

From these considerations and the potential for the decreased fetal weight to reflect primarily

the larger litter size, the decreased fetal weight in this study could be considered an equivocal

LOAEL. Thus, on the basis of decreased fetal body weight, the mid dose in this study of 60 mg/kgday was a NOAEL for developmental toxicity and the high dose of 120 mg/kg-day was an equivocal

LOAEL. The high dose was a maternal NOAEL. BMD modeling could not be done for the

decreased fetal weight because NTP did not have information on the fetal weight by sex,
69

Table 10. Key Results from Developmental Toxicity Study in Rats

Administered Phenol by Gavage (NTP, 1983a)

Control 30 mg/kg-day 60 mg/kg-day 120 mg/kg-day

Live fetuses/litter1 12.23±0.51 13.32±0.51 12.14±0.56 13.75±0.62

No. litters with nonlive (%) 0

(0%)

5

(22.7%)

7

(31.8%)

3

(15%)

Nonlive/liver 0 0.41±0.18 0.32±0.10 0.25±0.14

Historical control data for this strain on nonlive/litter (Implants/pregnant female minus live fetuses/pregnant female) (Charles River

Laboratories, 2001): 0.6-0.8, based on 62 studies.

Males/litter 6.23±0.38 7.36±0.52 5.14±0.56 7.35±0.57

Average fetal body weight per litter (g) 4.14±0.07 4.10±0.05 4.03±0.07 3.84±0.05**

Historical control fetal weight (from Charles

River Laboratories, 1988)

Mean 3.39 g

Range 3.04-

3.52 g

Total ave. fetal weight/dam (g)

(calculated)

50.6 54.6 48.9 52.8

Gravid uterine weight (g) 76.9±3.0 82.9±3.4 75.4±3.2 81.7±3.7

Treatment period maternal weight gain (g) 41.0±1.1 47.2±1.7** 40.2±1.6 41.2±2.9

Absolute maternal weight gain (adjusted for

gravid uterine weight) (g)

58.0±3.1 58.4±2.2 52.7±2.4 51.8±3.2

1

Mean ±Standard error of the mean. 

**Statistically significant, p<0.01
7

Michael Shelby, NTP, personal communication to Lynne Haber (TERA), March 13, 2002.

70

either in the report or in its archives.7

 Data on fetal weight by sex is needed for meaningful modeling

because the average weight of males and females is different and the number of males per group

varied.

The preliminary rat developmental toxicity studies (NTP, 1983a) found that phenol toxicity is

increased by the use of small dosing volumes. For example, when phenol was administered by

gavage on GDs 6–15 to pregnant CD rats at doses of 0, 125, 160, 200, or 250 mg/kg-day in a volume

of 1 mL/kg, the mortality was 0% (0/9), 70% (7/10), 78% (7/9), 100% (9/9), and 100% (9/9),

respectively. The deaths were preceded by dose-related signs of toxicity, including tremors,

convulsion, and respiratory distress; mottled liver and congested lungs were found on necropsy. In

contrast, when the same doses were administered in a volume of 5–7.5 mL/kg, the respective

mortality was only 0% (0/24), 0% (0/5), 17% (1/6), 17% (1/6), and 71% (5/7), respectively. On the

basis of these results, a volume of 5 mL/kg was used in the main developmental toxicity study. 

In preliminary toxicity studies conducted with doses of 60–250 mg/kg-day in a volume of

5–7.5 mL/kg, decreased maternal body weight gain (or body weight loss) during dosing was observed

at 160 mg/kg-day and up, doses at which mortality was also observed. In addition, tremors were

observed sporadically in the phenol-dosed groups, without any clear dose-response. There were no

treatment-related changes in prenatal viability, fetal sex ratio, or fetal structural development. 

The study authors stated that when results of all of the preliminary studies were pooled, a

statistically significant trend of decreasing fetal weight was observed, but there were no statistically

significant differences from controls in pairwise analyses. The power of the pairwise tests was

limited because only 4–6 litters were produced in the dose range 100–200 mg/kg-day.

In a standard mouse developmental toxicity study (NTP, 1983b), phenol was administered by

gavage in water at 0, 70, 140, or 280 mg/kg-day on GDs 6 to 15 in a volume of 10 mL/kg. Groups of

31–36 plug-positive female CD-1 mice were used in each treatment group. The pregnancy rate in the

controls was only 83%; the pregnancy rate in dosed animals ranged from approximately 83% in the

low- and mid-dose groups to 71% at the high dose. In addition, 4/36 high-dose mice died; no deaths

occurred in any other groups. The average maternal body weight gain during treatment was

statistically significantly reduced at the high dose, as was the maternal body weight at terminal

sacrifice on GD 17 (by 10%, compared with the control group). In addition, tremors were observed

at the high dose throughout the dosing period. As in the rat study, there was a highly statistically

significant decrease in fetal body weight per litter (18%) at the high dose. An increased incidence of

cleft palate was also reported at the highest dose level, although the incidence was not significantly
71

different from that of the other groups and there was no statistically significant increase in the

incidence of litters with malformations. There was no other evidence of altered prenatal viability or

structural development. 

Thus, the high dose of 280 mg/kg-day was a maternal frank effect level based on the observed

deaths; tremors and decreased body weight also occurred at this dose. The high dose was also a

developmental LOAEL based on decreased fetal body weight (accompanied by a possible increase in

the incidence of cleft palate) in the fetuses, an effect that was likely secondary to the severe toxicity

in the dams. The study NOAEL for maternal and developmental toxicity was 140 mg/kg-day. 

The series of oral screening studies mentioned above (Narotsky and Kavlock, 1995; Berman

et al., 1995; Moser et al., 1995; MacPhail et al., 1995) also included a developmental toxicity

screening study in which groups of pregnant F344 rats (15–20 animals/group) were given phenol at

doses of 0, 40, or 53.3 mg/kg by gavage once daily on GDs 6–19 and then sacrificed on postnatal day

(PND) 6 (Narotsky and Kavlock, 1995). The dosing volume was 1 mL/kg. Pups in each litter were

examined and counted on PND 1, 3, and 6 and were weighed on PND 1 and 6. Uterine implantation

sites were counted after the dams were sacrificed. Only minimal quantitative data were presented. 

No maternal deaths were observed. The authors reported that phenol treatment altered respiration

(rales and dyspnea) at both dose levels, but no quantitative data were presented. Decreased (but not

statistically significant) maternal body weight (compared with the controls) and decreased

(statistically significant) maternal body weight gain were also reported at both doses, but there was

no clear dose-response.

No statistically significant evidence of developmental toxicity due to phenol exposure was

observed. The only evidence of developmental toxicity came from dams that exhibited severe

respiratory signs. These signs of developmental toxicity included a dose-dependent increase in fulllitter resorptions (one at the low dose and two at the high dose) and an excessive incidence of

perinatal mortality and reduced pup weights on PND 1 in one litter at the high dose. However, these

changes as a group were not significantly different from those in the controls. Nonetheless, the

respiratory effects from oral dosing reported in this study are of interest, particularly as they were not

reported in the related study of systemic toxicity (Moser et al., 1995). This difference may reflect

differences in the completeness of the study reporting. Alternatively, it may suggest that pregnant

females may be more sensitive than nonpregnant females to the toxic effects of phenol.
72

4.4. OTHER STUDIES 

4.4.1 Initiation/PromotionStudies, Other Short-Term Tumorigenicity Assays, and Cancer

Mechanism Studies

Several studies have tested the promotion potential of dermally administered phenol. These

studies found that phenol promotes tumors initiated with dimethyl=benzanthracene (DMBA), but the

phenol doses tested caused ulceration (Salaman and Glendenning, 1957; Boutwell and Bosch, 1959)

and thus were well above the MTD. One study (Wynder and Hoffman, 1961) found promotion of

DMBA-initiated tumors by dermally administered phenol at a concentration that caused “no toxic

reactions.”

Salaman and Glendenning (1957) conducted an initiation/promotion study in which groups of

20 male “S” strain mice were initiated with a single dermal treatment with DMBA and promoted with

dermal treatment with 0.5 mg/mouse phenol in acetone using two different treatment concentrations

for the same applied dose. The phenol was applied beginning 3 weeks after the DMBA application

for either 24 weeks in a volume of 0.025 mL as a 20% solution or for 32 weeks in a volume of 0.1

mL as a 5% solution (rotating the weekly applications among four application sites for both

concentrations). The study did not report whether the application site(s) were covered or whether the

animals were restrained from licking the site. 

The high concentration produced local ulceration that healed just in time for the next

treatment 4 weeks later, whereas the low concentration produced only transient light crusting that

tended to decrease as the experiment progressed. It is unclear how severe the skin effects would have

been if the low concentration had been repeatedly applied to the same site rather than being rotated

among four sites. Tumors were observed in both treatment groups, with a shorter time to first tumor

and a higher tumor burden in the group treated with the higher phenol concentration. A few

histologically confirmed malignant tumors (primarily squamous epitheliomas) were observed in both

groups. In mice that underwent the same phenol treatment but were not pretreated with DMBA,

seven papillomas were observed at the high concentration. No tumors were observed at the low

concentration, even though the weekly dose was the same and the total dose per mouse was higher

(because the duration was longer).

This study had no control group on DMBA-only group, but the absence of tumors at the low

concentration indicates that the observed tumors were phenol related. The authors noted that the

observed tumors might have been related to the significant skin injury produced by phenol. This

suggestion is supported by the strong effect of the concentration of applied phenol at the same total

dose.
73

Boutwell and Bosch (1959) conducted a series of initiation/promotion studies with different

strains of mice. The mice were pretreated with a single application of 75 μg DMBA in benzene

followed by 5% or 10% phenol (1.25 or 2.5 mg per application) in benzene or with either dose of

phenol alone twice weekly for 52 to 72 weeks. An additional group received DMBA alone,

apparently followed by benzene vehicle, although there is some inconsistency between the text and

the summary tables regarding whether the control group received the benzene vehicle. At the high

phenol dose, dermal treatment with phenol resulted in decreased body weight (compared with the

controls) and decreased survival. Skin wounds, hair loss, and reparative hyperplasia were also seen

at the high dose, with the wounds predominantly seen in the first 6 weeks of treatment. By contrast,

the authors stated that there was no evidence of ill effects of 5% phenol except for its promoting

activity. This statement was based on external observation; no histopathology was conducted.

A dose-related increase in papillomas and in carcinomas was observed in the groups initiated

with DMBA and promoted with phenol. Increased papillomas were also observed in one strain

treated with the high dose of phenol alone. There was evidence of decreased activity when phenol

was further purified, indicating that the activity was not due to a contaminant. Because the benzene

vehicle is a defatting agent, it is unclear whether it would have contributed to the effect of phenol. 

Wynder and Hoffmann (1961) also found that dermally applied phenol is a promoter. Female

Swiss mice (28–30/group) were initiated with a single application of DMBA followed by treatment

with 5% phenol three times weekly or 10% phenol two or three times weekly. The dilution vehicle

was not reported. “No toxic reactions” were reported at 5% phenol, although the higher

concentration was reported to be “rather toxic.” Treatment was for 12 months, and the mice were

observed for an additional 3 months; the percentage of animals with papillomas and with cancers was

recorded monthly. 

At 10 months, papillomas were seen in 33% of the low-dose group and > 80% of the highdose group; cancer was seen in 3% of the low-dose animals and 30–60% of the high-dose groups. By

contrast, there were no papillomas or cancers in female Swiss mice treated with phenol alone and

only 10% papillomas (no cancer until week 12, and only 7% of the animals had cancer at study

termination) in the mice treated with DMBA alone. Survival decreased markedly after week 10 in

the high-dose groups but not the other groups. In another experiment, the onset of tumor formation

was much earlier in mice treated with 0.005% benzo[a]pyrene three times weekly plus 5% phenol

twice weekly than in mice treated with benzo[a]pyrene alone. Papillomas were observed by the

second week in the groups receiving benzo[a]pyrene and phenol and were present in at least 33% of

the animals by week 5, compared with 3% of the mice at week 5 in the benzo[a]pyrene-only group. 
74 

In a short-term assay, Stenius et al. (1989) found that phenol did not increase the production

of gamma-glutamyl transpeptidase (GGT)-positive foci. Groups of 7–9 male Sprague-Dawley rats

were partially hepatectomized and treated with diethylnitrosamine and then with 100 mg/kg phenol

by gavage for 5 days (gavage volume not reported). Phenobarbital, the positive control, produced a

marked increase in GGT-positive foci. This assay was based on the assumption that GGT-positive

cells in enzyme-altered foci represent initiated cells and the observation that these cells are often

resistant to toxic effects.

A decrease in tumor formation was seen in a co-carcinogenesis study of phenol and

benzo[a]pyrene (Van Duuren and Goldschmidt, 1976; Van Duuren et al., 1971, 1973). Phenol was

applied at a dose of 3 mg/application in acetone to the clipped skin of female ICR/Ha Swiss mice

(50/group) three times a week for 1 year simultaneously with 5 :g of benzo[a]pyrene. The resulting

number of tumors (both papillomas and squamous cell carcinomas) was markedly lower than in the

animals receiving the benzo[a]pyrene alone. Phenol alone did not cause skin tumors. Neither the

application volume nor the application surface area were reported, and no information was provided

on any skin effects other than tumors.

In a test of a (TG •AC) transgenic mouse line carrying a v-Ha-ras gene fused to a . globin

promoter, Spalding et al. (1993) found that phenol did not produce papillomas. This strain has

genetically initiated skin and has been shown to be sensitive to the known promoter 12-Otetradecanoylphorbol-13-acetate (TPA). Five male mice were dermally treated twice per week for 20

weeks with 3 mg phenol. Only one papilloma was observed; in contrast, strongly promoting agents

produced five or more papillomas per mouse. This dose did, however, cause chronic irritation and

hair loss.

A number of studies have investigated the reason why benzene is carcinojenic but orally

administered phenol is not, in light of the fact that when phenol and many of its metabolites are

significant products of benzene metabolism. Medinsky et al. (1995) noted that, on the basis of the

urinary metabolite profile, higher levels of hydroquinone are produced after benzene exposure than

after exposure to comparable doses of phenol. The potential production of other toxic metabolites,

such as muconaldehyde, following benzene exposure but not phenol exposure was also noted. The

authors explained the different metabolite profiles of phenol and benzene using the zonal distribution

of metabolizing enzymes in the liver. As described in Section 3.3, phenol is conjugated in the gut

and in zone 1 of the liver. This reduces the amount of phenol that reaches zone 3 of the liver, where

oxidative activity is highest, and so decreases hydroquinone production. By contrast, conjugation of

benzene in the gut and zone 1 is low, because benzene must be oxidized prior to conjugation. This
75

results in more free phenol reaching zone 3 of the liver following benzene exposure than after phenol

exposure, and hence more production of hydroquinone.

Equivocal or negative results were obtained with phenol in a well-conducted and wellcontrolled interlaboratory study evaluating the usefulness of the Chinese hamster V79 cell metabolic

cooperation assay for detecting tumor promoters (Bohrman et al., 1988). The study authors noted,

however, that the assay was conducted in the absence of exogenous metabolic activation, and V79

cells have low intrinsic metabolic capacity.

Miyagawa et al. (1995) conducted a validation test in male B6C3F1 mice of the in vivo-in

vitro replicative DNA synthesis test. The test was based on the hypothesis that nongenotoxic

carcinogens are likely to increase cell proliferation. Phenol was negative in this assay, which was

conducted at 0, 300, and 600 mg/kg administered via oral gavage.

4.4.2. Genotoxicity

The genotoxic potential of phenol appears to depend on the competing processes of activation

to a genotoxic form and metabolic inactivation (e.g., via conjugation). Phenol tended to be negative

in bacterial gene mutation assays (Pool and Lin, 1982; Rapson et al., 1980; Haworth et al., 1983) but

was positive or equivocal in mammalian cell gene mutation assays (McGregor et al., 1988a,b;

Paschin and Bahitova, 1982; Tsutsui et al., 1997) (Table 11). Phenol tended to induce micronuclei in

mice when administered intraperitoneally (Marrazzini et al., 1994; Chen and Eastmond,1995a;

Ciranni et al., 1988b) but was negative (or positive only at very high doses) when administered orally

(Ciranni et al., 1988b; Gocke et al., 1981). This difference is likely due to the first-pass conjugation

and inactivation of orally administered phenol. Phenol was also positive in in vitro micronucleus

tests with human lymphocytes (Yager et al., 1990) and Chinese hamster ovary (CHO) cells (Miller et

al., 1995) and caused chromosome aberrations in the presence of S9 activation in CHO cells (Ivett et

al., 1989). Results from DNA damage assays are inconsistent, but they tend to show that phenol can

cause sister chromatid exchanges (Erexson et al., 1985; Ivett et al., 1989) or cell transformation

(Tsutsui et al., 1997) if it is not metabolically inactivated. 

Phenol was negative in a well-conducted assay Salmonella typhimurium reverse mutation

assay performed with up to cytotoxic doses in the presence and absence of varying concentrations of

S9 activation with strains TA1535, TA1537, TA1538, TA98, and TA100 (Pool and Lin, 1982). 

Phenol was tested in two independent laboratories as part of a large-scale test by NTP in salmonella

strains TA1535, TA1537, TA98, and TA100 in the presence and absence of S9 activation (Haworth

et al., 1983). Both laboratories found that phenol was negative. Rapson et al. (1980) also reported
76

that phenol was negative in a test in strain TA100, although no primary data were presented and it

was unclear whether sufficiently high doses were tested. A weak positive response was reported with

phenol in strain TA98 in the presence of S9 but not in the absence of S9 (Gocke et al., 1981). Other

strains were also tested in that assay, but the results were unclear.

Positive or equivocal results have been reported in mammalian cell gene mutation assays. 

McGregor et al. (1988a, b) reported on a well-conducted mouse lymphoma L5178Y tk+/tk- assay of

phenol performed as part of a test of 72 coded chemicals. In the absence of S9, the results were

considered questionable or inconclusive in two independent assays because of the absence of a doserelated trend and increases occurring only in the presence of high cytotoxicity. In the presence of S9,

the first test was questionable (no dose-related trend but statistically significant results at several

doses), but a clear positive result was obtained in the confirmatory test. Overall, the study authors

concluded that no definitive conclusion was possible. 
77

Table 11. Summary of Genotoxicity Studies

Endpoint Assay system Results (wo/w

Activation)

Comments Reference

In Vitro Studies

Gene mutationbacteria

Salmonella

typhimurium

TA1535, TA1537,

TA1538, TA98,

TA100

-/- Tested to cytotoxic

doses, varying S9

concentrations

Pool and Lin,

1982

Salmonella

typhimurium

TA1535, TA1537,

TA98, TA100

-/- Part of NTP testing;

tested in 2

laboratories

Haworth et al.

1983

Salmonella

typhimurium

TA100

-1 Unclear if

sufficiently high

doses tested

Rapson et al.

1980

Salmonella

typhimurium

TA98

-/w 2 Other strains also

tested, but results

with them unclear

Gocke et al.

1981

Gene mutation -

mammalian cell

Mouse lymphoma

L5178Y cells

?/?2 Two independent

assays conducted +/-

S9

McGregor et

al. 1988a,

1988b

Chinese hamster

V79 cells

NT2/+ S9 from

phenobarbitalinduced mice

Paschin and

Bahitova 1982

Syrian hamster

embryo (SHE) cells

NT/+ None Tsutsui et al.

1997

Clastogenicity Micronuclei in

human lymphocytes

+/NT None Yager et al.

1990

Micronuclei in CHO

cells

+/+ S9 from

phenobarbital/betanaphthoflavone

induced rats

Miller et al.

1995

Chromosome

aberration

CHO cells -/+ Part of NTP testing Ivett et al.

1989
Endpoint Assay system Results (wo/w

Activation)

Comments Reference

78

DNA damage Sister chromatid

exchange, human

lymphocytes

-/NT Unclear if

sufficiently high

doses tested

Jansson et al.

1986

Sister chromatid

exchange, human

lymphocytes

+/NT Small increases Morimoto et

al. 1983;

Morimoto and

Wolff 1980

Sister chromatid

exchange, human

lymphocytes

+/NT None Erexson et al.

1985

Sister chromatid

exchange, CHO cells

+/w Part of NTP testing Ivett et al.

1989

Unscheduled DNA

synthesis, SHE cells

+/NT None Tsutsui et al.

1997

Single strand breaks

mouse lymphoma

cells

- None PellackWalker and

Blumer, 1986

Single strand breaks

CHO cells

- None Sze et al. 1996

Cell transformation Syrian hamster

embryo (SHE) cells

+/NT None Tsutsui et al.

1997

In Vivo Studies

Gene mutation Drosophila sexlinked recessive

lethal

- None Gocke et al.

1981

Drosophila sexlinked recessive

lethal

- None Sturtevant

1952

Drosophila sexlinked recessive

lethal

- None Woodruff et

al. 1985

Clastogenicity Mouse micronucleus

i.p.

 + None Marazzini et

al. 1994
Endpoint Assay system Results (wo/w

Activation)

Comments Reference

79

Mouse micronucleus

i.p.

 + Weak response Chen and

Eastmond

1995a

Mouse micronucleus

i.p.

 - No positive control,

unclear if

sufficiently high

doses tested

Barale et al.

1990

Mouse micronucleus

i.p.

 - Small sample size Gocke et al.

1981

Mouse micronucleus

oral

w None Ciranni et al.

1988b

Mouse micronucleus

i.p.

 + Clear effect at same

dose as oral study

Ciranni et al.

1988b

Mouse micronucleus

oral

- Number tested not

reported

Gad el-Karim

et al. 1985

Mouse micronucleus

oral

 + Pregnant females on

GD13

Ciranni et al.

1988a

Chromosome

aberration

Mouse,

spermatogonia and

spermatocytes 

 + Inconsistencies in

reporting

Bulsiewicz

1977

DNA damage Single strand breaks,

testicular cells, i.p.

- None Skare and

Schrotel 1984

1

Apparently in the absence of S9 - the presence of absence of S9 was not addressed. 

2

w = weak positive response; ? = questionable or inconclusive; NT = not tested
80 

Other authors have also reported positive results in mammalian gene mutation assays. 

Paschin and Bahitova (1982) found that phenol was mutagenic in an in vitro assay for mutagenicity

at the HGPRT locus of V79 Chinese hamster cells in the presence of S9 from the livers of

phenobarbital-induced mice. Tsutsui et al. (1997) also reported that phenol induced a dose-related

increase in mutation frequency in Syrian hamster embryo cells.

In contrast with these positive results in mammalian cells, a number of authors (Gocke et al.,

1981; Sturtevant, 1952; Woodruff et al., 1985) found that phenol was negative in sex-linked

recessive lethal assays conducted in drosophila using the feeding and injection routes.

The potential for phenol to induce micronuclei (a measure of clastogenicity) appears to be

related to the route of dosing, with generally positive results when phenol is administered

intraperitoneally but negative or equivocal results when it is administered orally. This route-related

difference is likely due to the potential for first-pass detoxification of phenol when it is administered

via the oral route but not when administered intraperitoneally. Several authors have suggested that

the chromosomal effects of phenol result from phenol interactions with the spindle apparatus

(Bulsiewicz, 1977; Yager et al., 1990). No in vitro studies of phenol clastogenesis were located.

Phenol was positive in the micronucleus test in male mice at an i.p. dose of 120 mg/kg

(Marrazzini et al., 1994). Similarly, Chen and Eastmond (1995a) reported a weak increase in bone

marrow micronuclei of male CD-1 mice treated with up to 160 mg/kg-day phenol intraperitoneally

for 3 days. By contrast, Barale et al. (1990) found that phenol was negative when administered at

i.p. doses of up to 160 mg/kg to male CD-1 mice. However, it is unclear whether sufficiently high

doses were tested in that study, since no cytotoxicity and no clinical signs of toxicity were reported;

also, the study did not include a positive control. Gocke et al. (1981) also found that phenol was

negative in male and female NMRI mice treated with i.p. phenol at two daily doses of up to 188

mg/kg per dose, although the sample size of tested animals was small. 

Ciranni et al. (1988b) found that an oral dose of 265 mg/kg phenol caused a slight increase in

micronuclei and some myelotoxicity (decreased ratio of polychromatic to normochromatic

erythrocytes, PCE/NCE ratio), but i.p. administration of the same dose caused clear increases in

micronuclei and stronger myelotoxicity. Gad-el-karim et al. (1985) reported that a single oral dose

of phenol (250 mg/kg) was negative for micronucleus formation in male CD mice, but they did not

report the number of animals tested. This difference between the effects of i.p. and oral

administration of phenol is also consistent with the metabolic effects of first-pass metabolism

mentioned in Section 3.3.
81 

Ciranni et al. (1988a) found that a single gavage dose of 265 mg/kg phenol caused a small

but statistically significant increase in bone marrow micronuclei—accompanied by cytotoxicity—in

pregnant female CD-1 mice treated on GD 13. There was no effect on fetal liver. Although no

positive control was included, benzene did cause micronuclei in fetal liver, confirming the

sensitivity of the assay. 

Phenol was positive in an in vitro assay for the development of micronuclei in human

lymphocytes in the absence of exogenous metabolic activation, although the dose-response was

weak (Yager et al., 1990). Miller et al. (1995) also found that phenol was positive in an in vitro

micronucleus test in CHO cells in the presence or absence of S9 from livers of phenobarbital/betanaphthoflavone induced rats, although a stronger response was observed in the presence of S9. 

Phenol was evaluated in the chromosome aberration assay in CHO cells as part of a series of

tests by NTP to evaluate genotoxicity assays (Ivett et al., 1989). No significant increase was

observed in the absence of S9 activation. In the presence of S9, significant increases in both simple

and complex aberrations were observed. A delayed harvest time was used due to cell cycle delay.

In a five-generation study of chromosome aberrations in spermatogonia and spermatocytes in

Porton strain inbred mice, Bulsiewicz (1977) observed dose-related increases in aberrations that

tended to increase with successive generations. Polyploidy was also observed. Three dose groups

and a control were treated by oral gavage. The dosing volume was reported as “2 mL of a solution

of phenol” (sic) for the low-dose group; volumes were not reported for the other groups. Phenol was

reported as being administered in a concentration of 0, 0.08, or 0.8 mg/L per day, or “8 g per liter”

(sic). 

A number of studies reported synergistic effects between phenol and hydroquinone in the

micronucleus assay in mice (Marrazzini et al., 1994; Barale et al., 1990; Chen and Eastmond,

1995a). At least part of this interaction is likely due to phenol enhancing the peroxidase-dependent

metabolic activation of hydroquinone. 

A number of studies have evaluated the potential for phenol to cause DNA damage. These

studies tend to show that phenol can produce effects when it is metabolized to an active form, but

that inactivation is likely to predominate over activation following oral dosing.

Jansson et al. (1986) found no effect on sister chromatid exchanges (SCEs) in an in vitro

assay with human lymphocytes, although it was unclear whether sufficiently high doses were tested. 

Small but statistically significant increases in SCEs in cultured human lymphocytes were reported by

Morimoto et al. (1983) and Morimoto and Wolff (1980). By contrast, Erexson et al. (1985) found a 
82

dose-related increase in SCEs in human lymphocytes exposed to phenol in vitro. They attributed the

difference between their results and negative results in other studies to differences in the procedure

used. In particular, Erexson and colleagues used mitogenic stimulation of the lymphocytes 24 hours

prior to the phenol exposure. This mean that the cells were blast-transformed and exposed in the G1-

S phase (and so there was less opportunity for repair prior to DNA replication), and cytochrome P450

activity was elevated as a result of the blast transformation. Negative controls showed that the

increases were not due to the mitogenic stimulation alone. As part of a series of tests by NTP to

evaluate genotoxicity assays, phenol was tested for the induction of SCEs in CHO cells (Ivett et al.,

1989). Phenol was positive in the absence of S9 and weakly positive in the presence of S9. Cell

cycle delay was observed at all positive doses.

In an assay with Syrian hamster embryo cells, Tsutsui et al. (1997) reported that phenol

induced a slight dose-related increase in transformed colonies as well as a dose-related increase in

unscheduled DNA synthesis in the same cell line.

In an evaluation of effects on germ cells in vivo, Skare and Schrotel (1984) found no effect on

single strand-breaks in testicular cells of Sprague-Dawley rats receiving a single i.p. injection of up to

79 mg/kg or five daily i.p. injections of up to 39.5 mg/kg-day. Phenol also did not induce single

strand breaks in mouse lymphoma L5178YS cells (Pellack-Walker and Blumer, 1986) or in CHO

cells in a test up to cytotoxic concentrations (Sze et al., 1996). 

Reddy et al. (1990) reported that DNA adducts were produced in cultured rat Zymbal glands

orally dosed with 750 μg/mL of either phenol or hydroquinone. The adducts were not chemically

characterized and their intensities were not quantified, but no spots were observed

autoradiographically in the untreated controls. By contrast, many different adducts were seen in the

analyzed tissues (bone marrow, Zymbal gland, liver, spleen) from both untreated female SpragueDawley rats and from rats treated for 4 days by oral gavage with a dose of 75 mg/kg-day phenol or

150 mg/kg-day of a 1:1 mixture of phenol and hydroquinone. The adduct patterns and levels of

adducts did not differ significantly between control and treated animals. 

The authors noted that endogenous adducts would interfere with the determination of

treatment-induced adducts that chromatograph similarly. To address this possibility, they compared

the chromatograms resulting from in vitro and in vivo treatments. The absence of the major in vitro

adducts of hydroquinone or benzoquinone in the in vivo samples suggested that these adducts were

not formed in the whole animal. Conversely, the primary adduct of phenol formed in vitro was also

observed in vivo, although the levels relative to controls were much higher under in vitro conditions. 
83

The authors suggested that the higher level of adducts following in vitro treatment versus in vivo

treatment could be attributed to detoxification of orally administered phenol, but they did not further

address the possibility that there may be a significant basal load of adducts formed by endogenously

produced phenol.

Using a fluorescence in situ hybridization approach, Chen and Eastmond (1995a) found that

treatment with phenol alone resulted in micronuclei and breaks in euchromatin, whereas

hydroquinone affected chromosome loss and chromosomal breakage, particularly in centromeric

heterochromatin. They suggested that the different pattern of effects with phenol and hydroquinone

indicates that the synergism between phenol and hydroquinone is not due solely to phenol-induced

increases in hydroquinone metabolism. Instead, they suggested, phenol or its metabolites may also

be inhibiting DNA repair. In a follow-up study, Chen and Eastmond (1995b) found that phenol alone

did not affect the DNA repair enzymes topoisomerase I or topoisomerase II in vitro. However,

mixing phenol with horseradish peroxidase to mimic the peroxidase metabolism of the bone marrow

resulted in complete inhibition of topoisomerase II; no effect was seen when glutathione was added to

the peroxidase mixture. 

In an assay with NCTC 929 mouse fibroblast cells, Yang and Duerksen-Hughes (1998) found

that phenol caused a dose-related increase in levels of the p53 protein. The authors noted that cells

increased p53 levels in response to DNA damage. 

4.4.3. Neurological Effects

As described above, tremors have been observed following relatively high exposures to

phenol via the oral (Dow Chemical Co., 1994; Moser et al., 1995) or inhalation (Dalin and

Kristoffersson, 1974) routes. Decreased motor activity and a statistically significant increase in 

rearing post-exposure were also reported in a screening study with rats (Moser et al., 1995), and

altered balance was reported in rats exposed continuously via inhalation for 15 days (Dalin and

Kristoffersson, 1974). However, in a 13-week drinking water neurotoxicity study that included

extensive neurohistological analyses (ClinTrials BioResearch, 1998), the only observed effects were

decreased motor activity and decreased body weight (compared with the controls), which were

probably secondary to decreased water consumption as a result of poor palatability. On the basis of

the results of a short-term screening study, neurotoxic effects do not occur at lower exposures than

other systemic effects of phenol (Berman et al., 1995; Moser et al., 1995).
84

4.4.4. Immunotoxicity

As described in Section 4.2, Hsieh et al. (1992) reported immune effects in CD-1 mice

administered phenol in drinking water for 28 days. The reported effects included decreased antibody

response (based on the PFC assay and direct antibody quantification using ELISA), with some

evidence of decreased lymphoproliferative response and decreased mixed lymphocyte response. The

clearest response was seen in the PFC and ELISA, which are highly predictive of effects on host

resistance (Luster et al., 1992, 1993). Confidence in the study results is decreased by the somewhat

small sample size (five rather than eight per group). Berman et al. (1995) reported atrophy of the

spleen or thymus of rats gavaged with phenol under conditions that greatly enhanced toxicity in

comparison with drinking water exposure.

Although no effects on spleen weight, cellularity, or antibody-forming cells in the spleen were

observed in a two-generation study of Sprague-Dawley rats exposed to phenol in drinking water at

concentrations of up to 5000 ppm (approximately 300 mg/kg-day) (Ryan et al., 2001; reported in

unpublished form as IIT Research Institute, 1999), qualitative and quantitative differences between

rats and mice in the effects of chemicals on the immune response are not unusual (e.g., U.S. EPA,

2000c). There is also no consistent pattern between mice and rats regarding which species is more

predictive of immunotoxic effects in humans. In light of Hsieh et al. (1992) who reported

immunotoxic effects of phenol at unusually low doses, it would be useful to confirm the results of

that study in mice using a protocol compliant with EPA immunotoxicity test guidelines (U.S. EPA,

1998c).

The National Institute of Environmental Health Sciences (NIEHS) and the NTP have elected

to conduct a comprehensive series of tests to evaluate the potential of phenol to adversely affect the

immune system (verbal communication from Dori Germolec, NIEHS to Monica Barron, EPA, 2002). 

Using test protocols designed to be consistent with EPA testing guidelines and GLP standards, the

first assay considered the same dosing regimen used in the Hsieh et al. (1992) study. That is, inbred

female B6C3F1 mice–rather than outbred CD-1 male mice–were exposed to drinking water

concentrations of 0, 5, 20, and 100 mg/L phenol (approximately 0, 1.7, 6.7, and 33 mg/kg-day). The

study also included a positive control. Preliminary results demonstrated immunosuppression

(reduced antibody response) at all levels of exposure, confirming the Hsieh study outcome. 

In order to characterize a wider range of response levels, a second assay was initiated using

drinking water concentrations of 0, 1.25, 2.5, 5.0, 20, and 40 mg/L (approximately 0, 0.4, 0.8, 1.7, 

6.7, and 13.3 mg/kg-day), overlapping the previous study’s range of exposures. In addition,

NIEHS/NTP has initiated host resistence studies. Depending on the final outcome of this series of 
85

tests, including NIEHS/NTP peer review, EPA may reconsider and, if appropriate, reopen this

assessment.

4.4.5. Other Studies

Eastmond et al. (1987) investigated the role of phenol in benzene-induced myelotoxicity. No

suppression of bone marrow cellularity was observed in male B6C3F1 mice treated intraperitoneally

with doses as high as 150 mg/kg twice daily for 12 days (daily doses up to 300 mg/kg). Only minimal

suppression was observed in mice dosed with hydroquinone at up to 100 mg/kg twice per day. By

contrast, marked statistically significant, dose-related suppression was seen in mice treated with 75

mg/kg phenol and 75 mg/kg hydroquinone under the same conditions. In further in vitro studies, the

authors showed that phenol stimulates the horseradish peroxidase-mediated metabolism of

hydroquinone, and they hypothesized that similar stimulation of local peroxidases occurs in the bone

marrow. The observation of myelotoxicity following benzene treatment–but only minimally or not at

all following phenol or hydroquinone treatment– was therefore explained by a more-than-additive

interaction between phenol and hydroquinone.

Corti and Snyder (1998) evaluated gender- and age-specific differences in cytotoxicity of

benzene metabolites in vitro. Bone marrow cells were harvested from adult unexposed male and

female Swiss Webster mice as well as from pregnant females and from fetal males and females. 

Cultures of CFU-e (colony forming units-erythroid, an erythroid precursor cell particularly susceptible

to benzene toxicity) were prepared and then exposed to different concentrations of the metabolites. 

Although most of the benzene metabolites caused marked cytotoxicity, only minimal toxicity (0–20%

cytolethality) of phenol was observed up to the highest concentration tested (40 M), compared with

nearly 100% cytolethality at the same concentration of catechol or hydroquinone. The effects were

strongest in cells isolated from fetal females or from virgin adult females, but the dose-response was

inconsistent, and it appears that no statistical comparisons with the untreated control of the same life

stage were done.

Zamponi et al. (1994) studied the mechanism of phenol-induced cardiac arrhythmia, including

ventricular tachycardia. In an abbreviated report, the authors suggested that phenol caused cardiac

arrhythmia by blocking batrachotoxin-activated cardiac sodium channels. Testing conditions,

including doses tested, were not provided in the abbreviated report. 

Bishop et al. (1997) investigated the effect of phenol exposure on total reproductive capacity

in mice. Groups of 26 female hybrid (SEC x C57BL6) F1 mice were given a single i.p. injection of 0

(buffer solution) or 350 mg/kg of phenol, and the females were caged individually with an untreated

male hybrid (C3H/R1 x C57BL10) F1 mouse following the day of injection for 347 days. The
86

animals were observed daily for producing newborn mice; the young mice were counted and

discarded immediately after birth. Female reproductive performance was evaluated on the basis of

the total number of offspring per female and the average number of litters per female. The numbers

of offspring per female and litters per female in phenol-treated mice were comparable to those in the

controls. Thus, phenol had no measurable detrimental effect on the parameters used for evaluating

long-term reproductive effects in this study. 

4.5. SYNTHESIS AND EVALUATION OF MAJOR NONCANCER EFFECTS AND MODE

OF ACTION

Studies investigating the effects of orally administered phenol are summarized in Table 2. 

When phenol is administered in drinking water, the most common effect is decreased water

consumption, presumably due to poor palatability. Effects seen concurrently with decreased water

consumption, and probably secondary to dehydration, include body weights lower than those of the

controls, decreased maternal body weight compared with that of developmental toxicity studies, and

decreased pup survival prior to culling. Decreased motor activity was also seen in a drinking water

neurotoxicity study, but it does not appear to be secondary to the decreased water consumption. The

data also indicate that phenol causes immune effects in mice, but not in rats.

Oral exposure to phenol also affects the kidney and liver. Kidney inflammation was observed

in the chronic drinking water study in rats (NCI, 1980). Liver and kidney pathology (tubular

degeneration, kidney necrosis, and vacuolar degeneration) in rats also has been observed in shortterm and subchronic toxicity studies using gavage dosing (Berman et al., 1995; Moser et al., 1995;

Dow Chemical Co., 1945). These results from experimental animal studies support the observations

in case studies and epidemiology studies (Shamy et al., 1994; Merliss, 1972) that the liver is a target

of phenol in humans. 

A number of nervous system effects have been observed following phenol dosing. Tremors

were observed in one animal that later died (apparently of dehydration) following dosing in drinking

water (ClinTrials BioResearch, 1998). Tremors have also been observed in several gavage studies in

rats and mice (NTP, 1983a; Dow Chemical Co., 1994; Moser et al., 1995). However, in a specialized

13-week neurotoxicity study in male and female rats that included an FOB and a detailed

neurohistopathology evaluation (ClinTrials BioResearch, 1998), the only observed nervous system

effects were tremors in one animal and decreased motor activity in females. A short-term gavage

screening study (Moser et al., 1995) found that the only effect in an FOB was a marginal decrease in

motor activity and increased rearing post-exposure.
87

Headaches and weakness were reported in workers exposed to a mixture of phenol and a

number of other aromatic compounds (Baj et al., 1994). Due to the mixed nature of the exposure, the

effects cannot be clearly attributed to phenol. Muscle pain in a laboratory technician who developed

phenol marasmus after being frequently exposed to phenol vapor (Merliss, 1972) may have reflected

neurological damage.

The data regarding the hematotoxic potential of phenol are conflicting. No hematological

effects were observed in rats in a well-conducted two-generation study in Sprague-Dawley rats (Ryan

et al., 2001; available in unpublished form as IIT Research Institute, 1999). By contrast, decreased

hematocrit and erythrocyte counts were seen at much lower doses in a 28-day drinking water study in

mice (Hsieh et al., 1992), although this study is limited by the use of only five males per dose. Data

from these two studies are contrasted in Table 6. The differences between the two studies cannot be

resolved by considering the results of the chronic drinking water studies conducted by NCI (1980) in

mice and rats, because no hematological evaluation was conducted in those studies. 

The negative finding in rats following oral exposure are supported by the absence of

hematological effects in rhesus monkeys, male Sprague-Dawley rats, and male albino mice (strain not

further identified) exposed to an average phenol concentration of 18.2 mg/m3

 continuously for 90

days (Sandage, 1961). Hemoglobin and hematocrit were also unaffected in a small study of rats

exposed to phenol in air at 100 mg/m3

 continuously for 15 days (Dalin and Kristoffersson, 1974). 

Neither of these inhalation studies used modern exposure protocols, and both were limited by

inadequate exposure monitoring. However, Dalin and Kristoffersson (1974) did both report systemic

effects (nervous system and liver effects) in rats at exposure levels that did not cause hematological

effects. Sandage (1961) found an indication of liver and kidney histopathology in the monkeys and

rats, although not in the mice. Assuming that all of the inhaled phenol was absorbed (see Section

3.2), the systemic dose to mice in the Sandage (1961) study can be estimated at approximately 30

mg/kg-day, based on a body weight of 0.03 kg and an inhalation rate of 0.052 mg/m3

/day. This dose

is comparable to the high dose in the Hsieh et al. (1992) study.

The negative results in the inhalation studies raise further questions about the reliability of the

hematotoxicity effects seen by Hsieh et al. (1992) in mice–particularly in the light of the small sample

size–as well as the relevance of those results to humans. Because portal-of-entry conjugation is more

efficient following ingestion rather than following inhalation of phenol (see Section 3.3), it is not

surprising that the systemic toxicity (i.e., liver and kidney effects) of a given absorbed dose may be

higher for inhaled phenol. Human data on hematotoxic effects of phenol are limited. Baj et al. (1994)

reported a small but statistically significant decrease in erythrocytes in workers exposed to a mixture
88

containing phenol, chlorinated benzenes, and other compounds. Due to the mixed nature of the

exposure, the effect, if any, cannot be clearly attributed to phenol.

The results of Hsieh et al. (1992) also indicate that phenol can be immunotoxic to mice. The

investigators observed a clear dose-related decrease in two related measures of antibody formation

(the PFC assay and a direct measure of antibody titer using ELISA), along with some evidence of a

decreased cell-mediated response at the high dose. Confirmation of their results in a repeat assay

conducted according to EPA test guidelines would be useful in light of the small number of animals

used and the limited number of risk assessments that have been based on in vivo/in vitro

immunotoxicity assays. 

No effect on spleen weight, cellularity, or antibody-forming cells (in the PFC assay) were

observed in a two-generation study of Sprague-Dawley rats exposed to phenol in drinking water at

much higher doses (Ryan et al., 2001; reported in unpublished form as IIT Research Institute, 1999),

but qualitative and quantitative differences in effects of chemicals on the immune response of rats and

mice are not unusual (e.g., see dioxin, as described in U.S. EPA, 2000c). “Necrosis or atrophy in the

spleen or thymus” (not further described) was observed in a 14-day screening study of rats gavaged

with phenol (Berman et al., 1995), supporting the immune organs as targets of phenol. 

Baj et al. (1994) reported in an epidemiology study of Polish workers that exposure to

Ksylamit® vapor resulted in immune effects, but it is unclear whether phenol is the causative agent

because Ksylamit® contains a number of different aromatic compounds. Overall, the data indicate

that phenol by itself may cause immunotoxicity in humans, but more data are needed to address this

possibility. Interaction between phenol and benzene metabolites may also cause immune effects, as

described below.

Benzene (which is metabolized to phenol) among other compounds, causes immunological

effects, including lymphopenia and leukopenia (reviewed in ATSDR, 1998). However, although

benzene is a leukemogen in humans, it has not been shown to induce leukemias in experimental

animals. For example, in the NTP gavage studies of benzene (NTP, 1986), it was carcinogenic to both

male and female F344 rats and B6C3F1 mice, inducing tumors at multiple sites. There was a

statistically significant increase in lymphomas in male and female mice but not in rats of either sex. 

No significant increase in leukemias was noted in either species. 

One proposed mechanism by which this immunotoxicity is induced involves the interaction

between phenol and hydroquinone, in which phenol stimulates the metabolism of hydroquinone. 

Eastmond et al. (1987) observed decreased bone marrow cellularity in male mice dosed

intraperitoneally with phenol and hydroquinone but not with phenol alone at doses of up to 300 mg/kgday for 12 days and only minimally with hydroquinone. These results appear to contradict those seen
89

at much lower doses (Hsieh et al., 1992), although the target tissue examined by Eastmond et al. (1987)

was bone marrow, whereas Hsieh et al. examined the spleen. The former study was also conducted via

the i.p. rather than oral route, but toxicity might be expected to be higher via the i.p. route because firstpass metabolism would be lower.

It is not known with certainty whether the toxic effects of phenol are due to the parent

compound or to its metabolites. Distinguishing between effects of each is complicated by the lack of

adequate data on concurrent blood levels of parent compound and metabolites. Phenol and metabolite

levels in blood would be expected to rise in parallel as portal-of-entry metabolism becomes saturated. 

However, several lines of reasoning suggest that many of the toxic effects are due to the oxidative

metabolites of phenol. In an in vitro study of the dysmorphogenic and embryotoxic effects of benzene

and phenol and their metabolites on whole rat conceptuses, Chapman et al. (1994) found that phenol

toxicity was much higher in the presence of S9.

The target tissues of phenol toxicity (kidney, liver, lung, and possibly bone marrow) are also

those in which phenol can be oxidatively metabolized. In a 2-week inhalation study, Dalin and

Kristoffersson (1974) observed altered balance and twitches in the absence of increased amounts of

phenol in the blood, suggesting that a phenol metabolite rather than phenol itself is the toxic agent. 

Alternatively, the analytical method used may not have been sensitive enough or specific enough to

detect any changes in blood levels of phenol. Phenol could produce portal-of-entry and systemic

toxicity as a result of its ability to react with and to denature protein.

A key point with regard to the evaluation of the toxicity of orally administered phenol is

whether gavage studies accurately represent the toxicity under environmental exposure conditions. 

Gavage studies are typically done using a single bolus dose per day, whereas environmental exposure is

more likely to involve exposure distributed over the course of the day. Although laboratory animals

consume drinking water in a few larger doses primarily during the active period rather than in

continuous small sips, the toxicokinetics of environmental exposure are more closely modeled by a

drinking water study or a gavage study using divided dosing, than by a study using a single gavage

dose per day.

Figure 2 compares the doses and observed severity of effects in drinking water and gavage
90

Figure 2. Plot of severity with dose for drinking water (DW) (open symbols) or gavage 

symbols). Values of 1, 2, 3 correspond with NOAEL, AEL, or FEL, respectively

012345

1 10 100 1000 

Dose (mg/kg-day)

Severity

DW NDW ADW FGavagGavagGavag
8

Although this unpublished study is not a primary reference for this assessment, it is presented

here because it contributes some useful information to the overall hazard identification phase of the

phenol assessment.

91

studies. Of particular interest is the number of gavage studies in which death, a frank effect, was

observed, whereas drinking water studies at comparable or higher daily doses produced only tremors,

kidney inflammation, and effects secondary to decreased water intake. The gavage NOAELs that

occur at the same doses as gavage adverse effect levels are for maternal and developmental toxicity. 

The sole exception to the large difference between gavage and drinking water studies in doses that

cause effects is the 28-day drinking water study by Hsieh et al. (1992), which reported hematological

and immune effects at doses comparable to NOAELs in drinking water studies. As noted above, it

would be useful to obtain independent confirmation in mice of the results reported by Hsieh et al.

(1992).

Toxicokinetic data support this difference between gavage and drinking water studies and

indicate that toxicity is correlated with peak blood concentrations rather than the area under the

curve. Dow Chemical Co. (1994) observed “phenol twitching behavior” (including tremors and eye

blinking) in rats gavaged with 150 mg/kg phenol; the behavior disappeared in less than an hour postdosing, as phenol blood levels declined below peak values. By contrast, no twitching behavior was

observed following a similar daily dose of phenol in drinking water. Unfortunately, blood levels of

phenol or its metabolites were not determined in the drinking water phase of the study, but they are

likely much lower than in the gavage phase, in light of the rapid blood clearance. 

The higher systemic toxicity of gavaged phenol when it is administered in smaller volumes

(NTP, 1983a) also supports the idea that toxicity is related to peak blood concentrations, because

smaller dosing volumes would be expected to enhance the absorption rate. An unpublished GLP

range-finding study for maternal toxicity (International Research and Development Corp., 1993) also

found higher systemic toxicity for phenol when it was administered in smaller dosing volumes.8

 Data

on the relationship between dosing volume and peak phenol blood concentrations are not available. 

Data on the relationship between peak blood concentrations and effects also are not available for the

endpoints relevant to the critical effect.

The inhalation data on the effects of phenol are very limited (Table 7). Only one study

conducted according to modern toxicological methods was located (Hoffman et al., 2001; available

in unpublished form as Huntingdon, 1998), and the exposures in this study were for only 2 weeks. 
92

Other studies ranged from 2 weeks (Dalin and Kristoffersson, 1974) to 90 days (Deichmann et al.,

1944; Sandage, 1961), but they included incomplete documentation of the study results, and they

did not use modern methods for controlling exposure levels. In addition, the authors of some of the

studies (e.g., Sandage, 1961) appear to have been looking for marked effects and thus dismissed

statistically significant incidences of organ pathology of lesser severity. Nonetheless, the studies

are fairly consistent with regard to the target organs and the effects observed. Exposure to high

concentrations produced nervous system effects, and liver, kidney, and lung pathology occurred at

lower concentrations. Rats were reported to be much less sensitive than rabbits or guinea pigs

(Deichmann et al., 1944). The systemic targets observed following inhalation exposure to phenol

are supported by data from the oral exposure route.

Information on the mode of action of inhaled phenol toxicity is also quite limited, but some

extrapolation from other routes is possible. On the basis of the irritative and corrosive effects seen

following dermal exposure to phenol, respiratory tract effects are likely due to direct contact of

phenol with the respiratory tract tissue. As noted in Section 3, phenol is extensively absorbed

following inhalation exposure. The lung can metabolize phenol prior to absorption, but the

efficiency of metabolism in the lung is lower than that for the gut or liver (Cassidy and Houston,

1984). After the inhaled phenol (and its metabolites) reaches the blood stream, the same points

described above for the oral route are relevant. In brief, it is not known whether the systemic toxic

effects of inhaled phenol are due to phenol itself or to its metabolite(s), but at least some of the

toxic effects appear to be attributable to phenol metabolite(s) (Chapman et al., 1994). Systemic

toxicity appears to be related to peak concentrations in blood rather than to total daily intake.

4.6. WEIGHT OF EVIDENCE EVALUATION AND CANCER

CHARACTERIZATION–SYNTHESIS OF HUMAN, LABORATORY ANIMAL,

AND OTHER SUPPORTING EVIDENCE, CONCLUSIONS ABOUT HUMAN

CARCINOGENICITY, AND LIKELY MODE OF ACTION

Chronic drinking water bioassays of phenol have been conducted in rats and mice (NCI,

1980). In these studies, NCI concluded that phenol was “not carcinogenic” in male or female F344

rats or B6C3F1 mice. However, the report also noted that leukemia and lymphoma were statistically

significantly increased in low-dose male rats, although there was no significant increase at the high

dose. The increases in leukemia are of particular interest in light of the leukemogenic effects of

benzene (for which phenol is a metabolite) in humans. (In experimental animals, benzene has not
93

been shown to induce leukemia, although increases in lymphoma have been observed [e.g., NTP,

1986].)

The MTD was clearly reached in the rat study (NCI, 1980), on the basis of decreased body

weight compared to controls and on kidney histopathology. Although the only sign of toxicity in the

mouse study was decreased body weight (compared to the controls) secondary to decreased water

consumption, higher doses probably could not have been tested, because of the decreased water

consumption. Higher toxicity probably could have been achieved in a gavage study in mice at lower

doses. These considerations suggest that an MTD was also reached in mice, although a definitive

conclusion is difficult.

No other long-term oral carcinogenicity studies of phenol are available. No inhalation

studies of phenol were of a sufficient duration to assess phenol carcinogenicity. The only long-term

study that has assessed the carcinogenicity of phenol applied dermally without initiation was that of

Boutwell and Bosch (1959), in which increased papillomas were seen at a dose that also caused

ulceration.

In contrast with these negative results for oral carcinogenicity, dermally administered phenol

has been consistently observed to be a promoter. Several authors (Salaman and Glendenning, 1957;

Boutwell and Bosch, 1959; Wynder and Hoffmann, 1961) observed that dermally applied phenol

promoted DMBA-initiated skin tumors. These studies have generally reported significant skin

ulceration at all phenol doses tested. The exception is Wynder and Hoffman (1961), who reported

that 5% phenol promoted DMBA-initiated tumors in mice in the absence of any toxic reactions. 

When the same phenol dose was administered in different volumes, higher promotion activity was

exhibited by the more concentrated solution, which also produced severe skin ulceration, suggesting

that some of the promotion activity may have been related to the rapid cell division in the repairation

of skin damage (Salaman and Glendenning, 1957). The observed response was dose-related

(Boutwell and Bosch, 1959), but marked systemic toxicity was also observed at these doses. Cocarcinogenesis with dermally administered benzo[a]pyrene has also been observed (Wynder and

Hoffmann, 1961). Because the benzo[a]pyrene was co-administered with the phenol, this assay

cannot be classified as a true initiation/promotion assay. Production of papillomas by dermally

administered phenol (in the absence of an initiator) was observed only at a concentration that caused

ulceration, and hence was above the MTD. 

Genotoxicity studies have found that phenol tends not to be mutagenic in bacteria (Pool and

Lin, 1982; Rapson et al., 1980; Haworth et al., 1983), but positive or equivocal results have been
94

obtained in gene mutation assays in mammalian cells (McGregor et al., 1988a, 1988; Paschin and

Bahitova, 1982; Tsutsui et al., 1997). Increases were larger in the presence of S9 activation. Phenol

tended to induce micronuclei in mice when administered intraperitoneally (Marrazzini et al., 1994;

Chen and Eastmond,1995a; Ciranni et al., 1988b), but it produced negative (or positive only at very

high doses) results when administered orally (Ciranni et al., 1988b; Gocke et al., 1981). This

difference is likely due to the first-pass conjugation and inactivation of orally administered phenol. 

Phenol was also positive in in vitro micronucleus tests with human lymphocytes (Yager et al., 1990)

and CHO cells (Miller et al., 1995), and it caused chromosome aberrations in the presence of S9

activation in CHO cells (Ivett et al., 1989). Phenol has been observed to act synergistically with

hydroquinone in the production of genotoxic effects (Marrazzini et al., 1994; Barale et al., 1990;

Chen and Eastmond, 1995a).

Epidemiology data do not shed further light on the carcinogenic potential of phenol. Some

studies (Kauppinen et al., 1986; Dosemeci et al., 1991) have reported elevated risks in phenolexposed workers, whereas others have observed no effect (Wilcosky et al., 1984). However, the

usefulness of each of these studies for risk assessment is limited by (depending on the study) an

absence of an effect when latency was considered, a lack of a dose-response, and potential for

confounding.

Although phenol was negative in oral bioassays conducted in rats and mice (NCI, 1980),

questions remain regarding its carcinogenic potential in light of the positive results in 

initiation/promotion assays (albeit at exposures typically above the MTD), the increases in leukemia

in low-dose male rats in the oral bioassay, and the observation of gene mutations in mammalian cells

in vivo and micronuclei in vivo following i.p. dosing. No inhalation studies of sufficient duration to

assess phenol carcinogenicity have been conducted. Dermal carcinogenicity or initiation/promotion

studies with phenol at exposures below the MTD have not been conducted. The carcinogenic

potential of phenol via inhalation exposure has not been evaluated at all. Under the draft revised

Guidelines for Carcinogen Risk Assessment (U.S. EPA, 1999), the data regarding the carcinogenicity

of phenol via the oral, inhalation, and dermal exposure routes are inadequate for an assessment of

human carcinogenic potential. Under the current guidelines (U.S. EPA, 1986a), phenol falls in

Category D: not classifiable as to human carcinogenicity.

Similar conclusions have been reached by other groups in recent assessments of the

carcinogenicity of phenol. IARC (1999) concluded that there is inadequate evidence in humans 

and in experimental animals for the carcinogenicity of phenol. Overall, IARC concluded that phenol
95

is not classifiable as to its carcinogenicity to humans (Group 3). Phenol is not included in the 9th

Report on Carcinogens (NTP, 2000a), and it is not listed as being considered for inclusion in the 10th

Report on Carcinogens (NTP, 2000b). This report contains only chemicals and substances that have

been reviewed and classified as known human carcinogens or as reasonably anticipated to be

carcinogens.

4.7. SUSCEPTIBLE POPULATIONS

People with decreased ability to conjugate and eliminate phenol, such as those with low

activities of PST or glucuronyltransferase, may be more susceptible to phenol toxicity. If the

toxicity of phenol is due to oxidative metabolites such as hydroquinone or catechol, individuals with

increased oxidative activity would also be more sensitive to phenol toxicity. The activity of

CYP2E1, the enzyme responsible for the oxidation of phenol, can be increased by exposure to a

variety of agents, including ethanol and chlorinated solvents, so people with high exposures to these

agents may be more sensitive to phenol.

4.7.1. Possible Childhood Susceptibility

As discussed in Section 4.3, a gavage study in rats (NTP, 1983a) reported decreased fetal

body weight at a dose below that at which maternal toxicity occurred. In contrast, maternal toxicity

occurred at a lower dose than did decreased fetal body weight in a gavage developmental toxicity

study that used a divided-dosing protocol (Argus Research Laboratories, 1997). The observation of

a fetal effect at a dose as low as 120 mg/kg-day (NTP, 1983a) suggests that the developing fetus is a

possible susceptible population. The strength of this conclusion is weakened, however, by the small

magnitude of the fetal weight decrease together with the increased litter size, which led to the high

dose being identified as an equivocal LOAEL (see Section 4.3). The strength of the conclusion is

also weakened by the observation in another gavage rat study (Argus Research Laboratories, 1997)

that decreased fetal weight occurs only at doses above those that cause decreased maternal weight

gain, and the observation of decreased fetal weight in a drinking water study (Ryan et al., 2001) only

at concentrations that also resulted in reduced water consumption.

Only one study was located that specifically addressed age-related differences in the

systemic toxicity of phenol. Deichmann and Witherup (1944) compared the lethality of an oral dose

of 600 mg/kg phenol (administered as a 5% aqueous solution) in 10-day-old, 5-week-old, and adult

rats. Mortality was 90%, 30%, and 60% in the neonates, young rats, and adult rats, respectively. 

Although the young and adult rats died within 1.5 hours of dosing, the neonates died 12–24 hours
96

after dosing. The data from this study suggest that neonates are more sensitive than adults and

young rats may be less sensitive than adults. Alternatively, the age-related differences observed in

this study could reflect inter-individual variability that was not a consequence of age. This study has

not been replicated; therefore, definitive conclusions are not possible.

Data from humans and rats are consistent in showing very little fetal expression of CYP2E1,

which is rapidly induced shortly after birth and rises though childhood (reviewed in Hakkola et al.,

1998). The age after parturition at which CYP2E1 levels peak has been studied in laboratory

animals, with inconsistent results. Some studies suggest that peak levels are reached during

childhood, with a subsequent decrease to adult levels (Schenkman et al., 1989), whereas others have

shown a rapid rise in CYP2E1 levels during childhood to a maximum level that is sustained during

adulthood (Song et al., 1986). If the toxic moiety is a metabolite, decreased P450 metabolism could

be expected to result in decreased toxicity. However, sulfate conjugation is also lower early in life

(Iwasaki et al., 1993), so more of the dose is available for oxidative metabolism. Indeed, Heaton and

Renwick (1991) found higher production of oxidative metabolites in young rats. This difference was

higher in males, with the percentage of the administered dose recovered as hydroquinone conjugates

decreasing from 38% of the administered dose in 3-week-old males to 8.2% in 16-week-old rats. 

Smaller decreases with age (from 17.8% of the administered dose in 4-week-old rats to 10.5% in 15-

week-old rats) were observed in females. These data suggest the potential for children to be more

sensitive than adults to the systemic effects of phenol.

4.7.2. Possible Gender Differences

Kenyon et al. (1995) (in mice) and Heaton and Renwick (1991) (in rats) reported higher

excretion of hydroquinone conjugates in males than in females, suggesting higher levels of

hydroquinone production in males. By contrast, Meerman et al. (1987) reported only slightly faster

metabolism in male rats. These data would tend to suggest that if hydroquinone is the toxic moiety,

phenol would be more toxic in males. However, few differences in phenol toxicity between males

and females were identified; differences in NOAELs reflect differences in water consumption per

unit weight, resulting in differences in estimated intake. Acute oral lethality data do suggest that

phenol is more toxic to males (Thompson and Gibson, 1984). 

5. DOSE RESPONSE ASSESSMENTS
97

5.1 Oral Reference Dose (RfD) 

5.1.1. Choice of Principal Study and Critical Effect 

An extensive database for the effects of orally administered phenol is available. The studies

relevant to the development of the RfD are summarized in Table 2. Two-year drinking water studies

conducted in rats and mice are available (NCI, 1980). Hematology and serum biochemical

evaluations were not included in those chronic studies, but they were included in a recent twogeneration drinking water study conducted in rats (Ryan et al., 2001; available in unpublished form

as IIT Research Institute, 1999). The only study evaluating hematological effects in mice is a 28-

day drinking water study (Hsieh et al., 1992). A specialized subchronic neurotoxicity study was

conducted with rats exposed to phenol in drinking water (ClinTrials BioResearch, 1998). A number

of developmental toxicity studies have been conducted in rats and mice, all using the gavage route

(Argus Research Laboratories, 1997; NTP, 1983a; NTP, 1983b, Narotsky and Kavlock, 1995).

As shown in Table 2, the study with the lowest NOAEL/LOAEL boundary is the 28-day

drinking water study in mice by Hsieh et al. (1992). In this study, the NOAEL was 6.3 mg/kg-day,

and the LOAEL was 33.6 mg/kg-day, based on decreased antibody response, supported by decreases

in hematocrit and red blood cells. Statistically significant decreases in erythrocyte counts were

observed at the low and mid doses, but these results were considered biologically questionable in the

absence of effects on hematocrit, in the absence of data addressing whether the apparent

inconsistency was due to macrocytosis, and in light of the lack of support from other studies. 

As noted in Section 4.5, this study is useful for hazard identification. However, confirmation

of the study results in an independent assay in mice would be useful before using the data for doseresponse assessment, considering the small number of animals tested and the limited number of risk

assessments that have been based on in vivo/in vitro immunotoxicity assays. In addition, although

qualitative differences between rats and mice in immune assays are not unusual (e.g., dioxin, see

U.S. EPA, 2000c), it is of interest that Hsieh et al. (1992) observed immune effects in mice at very

low doses, but the only other study evaluating similar immune parameters (Ryan et al., 2001; IIT

Research Institute, 1999) did not observe effects in rats at 10-fold higher doses. No other studies in

mice have directly evaluated effects on antibody forming cells. 

In another study that evaluated immune effects of phenol on mice (Eastmond et al., 1987), no

effect on bone marrow cellularity was observed at phenol doses of up to 300 mg/kg-day in mice 

dosed intraperitoneally for 12 days. Similarly, Corti and Snyder (1998) evaluated the effects of

benzene metabolites on CFU-e cells (an erythroid progenitor cell sensitive to benzene) harvested from
98

mice and found that the cytotoxicity of phenol was much lower than that of other benzene metabolites. 

The effects in mice and rats were not compared in that study. 

In light of these issues, and in the absence of other data supporting the observed effects at such

low doses, the results of Hsieh et al. (1992) are too preliminary to be used as the basis for the phenol

RfD. However, this study does raise concerns regarding the potential of phenol to cause

hematological and immune effects, and it would be worthwhile to address these questions in a mouse

immunotoxicity study conducted according to modern methods. The uncertainties regarding these

endpoints and the use of a database uncertainty factor to address these uncertainties are further

addressed in Section 5.1.3.

The next lowest NOAEL/LOAEL combination was observed in a 14-day gavage study in rats

conducted as part of a large-scale screening study of a number of chemicals (Berman et al., 1995;

Moser et al., 1995). Tremor, kidney tubular degeneration, and increased rearing in an FOB were

observed at the high dose of 40 mg/kg-day but not at the next lower dose of 12 mg/kg-day. Although

the incidence of kidney histopathology was not statistically significant, the high dose can be

considered a LOAEL in light of the low statistical power of the study (only eight female rats per

group) and the rarity of these lesions in short-term studies. The corresponding NOAEL is 12 mg/kgday.

The relevance of this NOAEL to environmental exposures is questionable, however, due to the

markedly higher toxicity observed in gavage studies than in drinking water studies, as discussed in

Section 4.5, and the absence of supporting toxicity in drinking water studies of much longer duration. 

In particular, drinking water studies found no kidney histopathology in rats exposed to 260 mg/kg-day

for 2 years (NCI, 1980) (although kidney inflammation was observed at higher doses), in mice

exposed to doses up to 660 mg/kg-day for 2 years (NCI, 1980), or in parental rats in a two-generation

reproduction study (Ryan et al., 2001; IIT Research Institute, 1999).

The only other study reporting kidney histopathology at low doses was a poorly documented

and unpublished 6-month gavage study in rats (Dow Chemical Co., 1945). The very small dosing

volume used by Berman et al. (1995) also may have contributed to the high toxicity, considering the

findings of NTP (1983a).

The principal study for development of the RfD is Argus Research Laboratories (1997), in

which decreased maternal weight gain was observed in rats gavaged on GD 6–15 with 120 mg/kg-day

phenol; the maternal NOAEL was 60 mg/kg-day, based on decreased body weight gain, and the

developmental NOAEL was 120 mg/kg-day. The BMDL was 93 mg/kg-day. No effect on body
99

weight was reported at 120 mg/kg-day in systemic toxicity studies using drinking water as the

exposure route (NCI, 1980; ClinTrials BioResearch, 1998), but it is not surprising that maternal

toxicity following 9 days of dosing occurs at a lower dose than does chronic systemic toxicity,

considering the different physiological status during pregnancy. Although the principal study was

conducted via the gavage route, measures were taken to decrease the impact of bolus dosing by

dividing the daily dose into three administrations per day.

It is of interest that rats consume drinking water not in many little sips, but in a few larger

doses primarily associated with food consumption during the active period of the day. Therefore, the

toxicokinetic profile of the divided-dose gavage study may actually be fairly similar to the

toxicokinetic profile that would be observed with drinking water exposure. In addition, a more precise

measurement of administered dose is possible in gavage studies, because spillage can occur in

drinking water studies.

The NOAEL identified in the Argus Research Laboratories (1997) study is supported by a 

developmental toxicity study (NTP, 1983a) in which decreased fetal weight was observed in CD rats

gavaged on GDs 6–15 with 120 mg/kg-day; the NOAEL was 60 mg/kg-day. The high dose of 120

mg/kg-day was considered an equivocal LOAEL for developmental effects, in light of the small

magnitude of the weight decrease, the increased litter size (which can result in decreased fetal weight),

and the absence of an effect on fetal weight at a maternally toxic dose in another gavage

developmental study in rats (Argus Research Laboratories, 1997). In the NTP (1983a) study, the

maternal toxicity NOAEL was the high dose, 120 mg/kg-day.

Because of the uncertainties regarding identification of the critical effect level for the NTP

(1983a) study, it was not considered to be an appropriate co-principal study. BMD modeling could

not be conducted on the fetal weight endpoint, because fetal weights were reported only as an average

across both sexes for all litters; no individual animal data were available. Because fetal weights of

male fetuses tend to be heavier than those of females, and because the number of fetuses per litter

affects the fetus weight, it was not appropriate to model the pooled data.

Although the decreased maternal weight gain (Argus Research Laboratories, 1997) was a

mild effect and was possibly confounded by the gavage dosing, these results are supported by a

drinking water study. Decreased motor activity was seen in female rats consuming the high

concentration of phenol (5000 ppm, corresponding to 360 mg/kg-day) in the 13-week neurotoxicity

study (ClinTrials BioResearch, 1998). The NOAEL in females was 107 mg/kg-day; no adequate fit

could be obtained using BMD modeling. As discussed in Section 4.2, the authors considered the
100

decreased motor activity to be secondary to dehydration, but an analysis of the individual animal

data and comparison with the literature could not confirm this assumption.

Ryan et al. (2001) conducted a two-generation drinking water study (also available in

unpublished form as IIT Research Institute, 1999) in rats in which decreased parental and pup

weight occurred at a LOAEL of 301 mg/kg-day, with a NOAEL of 71 mg/kg-day. However, these

lower body weights, compared with control, are likely to be secondary to decreased water

consumption and not an indication of phenol toxicity.

5.1.2. Method of Analysis: Benchmark Dose 

The RfD was derived by the BMD approach using BMDS Version 1.3, which downloaded

from the National Center for Environmental Assessment’s web site. The BMR was defined as the

default of a change of one standard deviation (U.S. EPA, 2000d). A BMDL of 93 mg/kg-day was

derived for decreased maternal weight gain (Argus Research Laboratories, 1997) using the

polynomial model. Similar BMDL values of 125 and 129 mg/kg-day were calculated using the

power and Hill models, respectively, although the fit (based on the Akaike Information Criterion

[AIC]) was slightly better using the polynomial model, and a more conservative BMDL was

obtained using this model. 

An alternative BMDL for this endpoint could be calculated using the geometric mean of the

BMDLs from all three models, 114 mg/kg-day, on the rationale that the small difference in AICs

was not meaningful. Other measures of fit (based on the goodness-of-fit p value and on visual fit)

also indicated that all three models are comparable. However, in this case the slightly more

conservative approach was used, in part as an added degree of protection because of the

uncertainties regarding immunotoxicity. Details of the model results are presented in 

Appendix B.

5.1.3. RfD Derivation 

The data on the within-human variability in the toxicokinetics and toxicodynamics of

ingested phenol are insufficient to adjust the default uncertainty factor for intraspecies variability

(UFA). In a sample of liver fractions from 10 people, Seaton et al. (1995) found that the kinetics of

phenol sulfation and hydroquinone conjugation varied by up to approximately threefold. Much

larger variability in CYP2E1 has been found, particularly between neonates and adults (Vieira et al.,

1996). These data on inter-individual variability in enzymatic metabolism are not adequate to move
101

from the default UFH of 10, because they do not reflect potential variability in portal-of-entry

metabolism of phenol or uncertainty regarding the identity of the toxic moiety. Furthermore,

variability in CYP2E1 does not necessarily translate directly into variability in tissue dose of

because metabolism by CYP2E1 may be limited by blood flow to the liver.

The absorption, distribution, and metabolism of ingested phenol in rats and humans appear to

be generally qualitatively similar, although the data are insufficient for a quantitative comparison. 

Comparison of laboratory animal and human phenol toxicokinetics is also limited by incomplete

information regarding the identity of the toxic moiety. As discussed in Section 4.5, the data suggest

that most of the toxic effects of phenol can be attributed to its oxidative metabolites, but the data are

insufficient to rule out the possibility that some effects may be attributable to phenol itself. In the

absence of adequate data on which to based a toxicokinetic or toxicodynamic comparison of rodents

and humans, the default UFA of 10 is used for interspecies extrapolation. However, it may be

possible to reduce this default value of 10 after review and evaluation of data (perhaps supplemented

by a PBPK model) that compare the toxicokinetics of phenol and its metabolites in the placenta and

fetus of rats and humans, if such data become available. 

The BMDL was based on an effect of minimal severity (decreased maternal weight gain),

and a higher BMDL and NOAEL were obtained for effects on maternal weight. The BMDL is also

within 50% of the NOAEL identified for the decreased maternal weight endpoint. Therefore, no

uncertainty factor (UF) is required for extrapolation from a NOAEL to a LOAEL. No UF for

extrapolation across duration is needed because this developmental study is supported by chronic

bioassays in two species in which toxicity was observed only at higher doses.

The database for phenol by the oral route can be considered complete. It includes 2-year

drinking water studies conducted in rats and mice (NCI, 1980), a two-generation drinking water

study conducted in rats (Ryan et al., 2001; available in unpublished form as IIT Research Institute,

1999), and gavage developmental toxicity studies in rats (Argus Research Laboratories, 1997; NTP,

1983a; Narotsky and Kavlock, 1995) and mice (NTP, 1983b). However, the range of endpoints

evaluated in the chronic toxicity studies was limited and did not include hematological or serum

biochemistry evaluations. Immunological and hematological effects in mice were observed by

Hsieh et al. (1992) in a 28-day drinking water at low doses. These endpoints were evaluated, and no

significant hematological or serum biochemistry effects were observed at doses of up to >300

mg/kg-day in the two-generation rat study (IIT Research Institute, 1999; Ryan et al., 2001). The

difference in these results suggest species differences between mice and rats, but confirmation of the
102

immunological and hematological effects in an assay done according to modern test methods would

be useful.

An i.p. study of the effects of phenol on bone marrow cellularity in mice at doses of up to

300 mg/kg-day (Eastmond et al., 1987) and an in vitro study with mouse bone marrow cells (Corti

and Snyder, 1998) also do not indicate that mouse blood cells are highly susceptible to effects of

phenol. However, these studies did not evaluate the same parameter measured by Hsieh et al.

(1992), and significant interspecies differences in immunotoxicity are not unusual. It is of interest

that the endpoints affected in the Hsieh et al. (1992) study (two measures of effects on antibodyforming cells, PFC and ELISA) are the immune endpoints most highly predictive of effects on host

resistance (Luster et al., 1992; Luster et al., 1993). Therefore, to account for the uncertainties

regarding the immunological and hematological effects in mice, a database uncertainty factor of 3 is

used.

An additional degree of public health protection may also be provided by the use of a gavage

study rather than the more environmentally relevant route of drinking water. This is because gavage

administration results in a higher peak blood level–presumably even using a divided-dosing

protocol–than does ingestion of the same daily dose in drinking water. Because at least some toxic

effects of phenol are related to peak blood levels rather than to total intake, toxicity would be

expected to be higher following gavage exposure than drinking water exposure.

A composite UF of 300 results. No modifying factor is applied because the existing

uncertainties have been addressed with the standard uncertainty factors.

RfD = 93 mg/kg-day /300 = 0.3 mg/kg-day, or 3E-1 mg/kg-day.

Note that this RfD is applied to ingested phenol in addition to the normal daily endogenous

production of phenol, as discussed in further detail in Section 6.1.2.

An additional uncertainty factor for sensitive populations such as infants and children is not

needed for phenol because sufficient studies of reproductive and developmental toxicity have been

performed.

5.2. Inhalation Reference Concentration (RfC) 

5.2.1. Choice of Principal Study and Critical Effect 
103

The minimal database needed for the development of an RfC is a well-conducted subchronic

inhalation study that has adequately evaluated a comprehensive array of endpoints, including the

respiratory tract, and established a NOAEL and a LOAEL (U.S. EPA, 1994b). This criterion was

not met for phenol. Neither of the two available subchronic studies (Deichmann et al., 1944;

Sandage, 1961) are adequate for exposure-response assessment, because neither included adequate

documentation of the histopathology results, and neither used modern methods for generating or

monitoring exposure levels. These studies can, however, be used for hazard identification, and they

identify the respiratory tract, liver, and kidney as targets of inhalation exposure to phenol.

The phenol database also includes a well-conducted 2-week inhalation study with rats that

used modern exposure methods, evaluated a wide array of endpoints, and included a thorough

histopathology evaluation of the respiratory tract (Hoffman et al., 2001; the full unpublished study

report is available as Huntingdon, 1998). The only treatment-related effect observed was a red nasal

discharge in male rats, which was observed with a statistically significant duration-related and

concentration-related incidence in the mid- and high-concentration groups. However, because the

red nasal discharge was likely due to a nonspecific response to stress, this response is not considered

adverse.

In the absence of an inhalation study of sufficient duration, no RfC for phenol can be

derived. A route-to-route extrapolation is not appropriate, because phenol can be a direct-contact

irritant, and so portal of entry effects are a potential concern. 

5.2.2. RfC Derivation

No RfC could be derived, due to insufficiencies of the database.

5.3. CANCER ASSESSMENT

As discussed in Section 4.6, the data regarding the carcinogenicity of phenol are

inadequate for assessment of human carcinogenic potential. Phenol was negative in oral

carcinogenicity studies in rats and mice, but questions remain regarding increased leukemia in

male rats in the bioassay as well as the positive gene mutation data and the positive results in

dermal initiation/promotion studies at doses at or above the MTD. No inhalation studies of an

appropriate duration exist. Therefore, no quantitative assessment of carcinogenic potential via

any route is possible.
104

6. MAJOR CONCLUSIONS IN CHARACTERIZATION OF HAZARD AND DOSE 

 RESPONSE

6.1. HUMAN HAZARD POTENTIAL

6.1.1. Oral Noncancer

In most studies of phenol administered in drinking water, water consumption was markedly

decreased at the highest dose, presumably due to poor palatability. A number of toxic effects

secondary to the decreased water consumption have been observed, including decreased 

body weight compared to controls, decreased pup weight, and decreased pup survival pre-culling. 

Other effects that may not have been secondary to decreased water consumption were kidney

inflammation (NCI, 1980) and decreased motor activity (ClinTrials BioResearch, 1998). Gavage

studies found more severe effects and reported these effects at lower doses. Observed effects

included lung, liver, and kidney pathology; tremors and other nervous system effects; and, at

sufficiently high doses, death.

These data suggest that the toxicity of phenol is higher via gavage dosing than via

administration in drinking water. The suggestion is supported by the finding that a series of

behaviors termed “phenol twitching behavior” correlate with peak blood levels rather than area

under the curve (Dow Chemical Co., 1994). For a given daily dose, peak blood levels would be

much higher following gavage dosing than following continuous administration in water. A direct

comparison of the toxicity of phenol when administered via these two routes could determine

definitively whether route-specific differences exist. Nonetheless, the data supporting the higher

toxicity of phenol administered by gavage were considered sufficiently strong to consider it

inappropriate to use the Berman et al. (1995) study (which also used small dosing volumes) as the

principal study. 

Developmental toxicity studies have been conducted only via the gavage route. In the

principal study (Argus Research Laboratories, 1997), decreased maternal weight gain was observed

in rats administered 120 mg/kg-day in a divided-dosing protocol. The BMDL for this study was 93

mg/kg-day and the NOAEL was 60 mg/kg-day. Although exposure in this study was for only 9 days,

comparison with the entire database for phenol via the oral route indicates that this study

appropriately identifies the critical effect. Because a maternal effect (decreased maternal body

weight gain during gestation) is considered the critical effect for phenol (i.e., the first adverse effect

or its known precursor that occurs to the most sensitive species as the dose rate of an agent increases),
105

protection from this effect would also be protective from systemic effects of chronic exposure, which

occur at higher doses. In light of the different physiological status during pregnancy, it is not

unreasonable for maternal toxicity following 9 days of dosing to occur at a lower dose than does

chronic systemic toxicity. 

Although a NOAEL of 60 mg/kg-day was identified for decreased fetal weight in the NTP

(1983a) study, this study was not considered adequate to be a co-principal study. The high dose of

120 mg/kg-day was considered an equivocal LOAEL for developmental effects, in light of the small

magnitude of the weight decrease and the absence of an effect on fetal weight at a maternally toxic

dose in another gavage developmental study in rats (Argus Research Laboratories, 1997). In

addition, although the observed decrease in fetal weight was small (but statistically significant),

increased litter size was also seen at this dose. It is possible, therefore, that the dams were near the

limit of what they could carry in terms of pup burden (total fetal weight).

The NOAEL was supported quantitatively by the NOAEL of 107 mg/kg-day for decreased

motor activity in a 90-day drinking water neurotoxicity study (ClinTrials BioResarch, 1998). A

NOAEL of 71 mg/kg-day for decreased parental and pup body weights was also identified in a

drinking water two-generation reproduction study (Ryan et al., 2001; available in unpublished form

as IIT Research Institute, 1999), although these effects are likely secondary to decreased water

consumption. There is, however, some uncertainty in the identification of the NOAEL for this latter

study. As described in Section 4.3, a statistically significant decrease in uterine weight was observed

at all three doses in this study. The decrease was not considered adverse for a number of reasons,

including the absence of a dose-response and the small number of animals outside the control range. 

This consideration, however, is rather subjective and is based on considerable professional

judgement.

A key uncertainty in the development of the RfD is the interpretation of the study by Hsieh et

al. (1992). Immunotoxicity (decreased response of antibody-forming cells) and hematotoxicity

(decreased red blood cells and hematocrit) were observed in this 28-day drinking water study in mice

at doses much lower than the doses than produced toxicity in other studies. No immunological or

hematotoxic effects were seen at much higher doses in a two-generation drinking water study in rats

(Ryan et al., 2001; IIT Research Institute, 1999). These differing results suggest species differences

between mice and rats, but confirmation of the immunological and hematological effects in an assay

done in mice according to modern test methods would be needed before using the data for doseresponse assessment, considering the small number of animals tested and the limited number of risk
106

assessments that have been based on in vivo/in vitro immunotoxicity assays. 

Similarly, in the absence of hematotoxicity in monkeys, rats, and mice following inhalation

exposure to phenol levels resulting in comparable or higher systemic doses of phenol (Sandage, 1961;

Dalin and Kristoffersson, 1974), confirmation of the reported hematological effects would also be

useful. A database uncertainty factor of 3 was used to account for the uncertainty regarding

immunotoxicity. This factor could be removed if an immunotoxicity study conducted according to

U.S. EPA (1998c) guidelines became available.

Although it does not directly affect the determination of the RfD, uncertainty also exists

regarding whether the decreased motor activity in females reported by ClinTrials BioResearch,

(1998) was due to dehydration only or whether phenol exposure also contributed to the effect. The

NOAEL from this study was used as supporting data for the principal study. The study authors

attributed the decreased motor activity to dehydration, because of the marked decrease in water intake

and the absence of supporting changes in the FOB. By contrast, this assessment concluded that

phenol at least contributed to the effect, because there was no clear correlation between individual

animals with dehydration and those with decreased activity and because the limited literature on the

topic reports no effect on motor activity of water deprivation for several days. A neurotoxicity study

in which the controls were allowed only limited access to drinking water would also address this

issue.

6.1.2. Inhalation Noncancer

The database for inhalation toxicity of phenol is very limited. A well-conducted 2-week study

is available (Hoffman et al., 2001; available in unpublished form as Huntingdon, 1998), but the

duration is less than that appropriate for serving as the basis for the RfC. Longer-term studies have

been conducted (Deichmann et al., 1944; Sandage, 1961), but they are limited by inadequate control

of exposure levels, unclear sensitivity of the evaluation, and limited reporting. 

However, the inhalation toxicity studies are sufficient however, to identify the respiratory

tract, liver, kidney, and nervous system as targets of inhaled phenol toxicity. A significant

uncertainty exists regarding which species is the most appropriate for extrapolation to humans. 

Deichmann et al. (1944) reported marked systemic toxicity in rabbits and deaths in guinea pigs at

exposure concentrations that caused no histopathology in rats. No inhalation studies in guinea pigs or

rabbits have been conducted to confirm these findings. In addition, it is unclear which of these

species is most like humans.
107

The primary data need for developing an RfC is a 90-day inhalation study that includes a

thorough examination of the respiratory tract. Pharmacokinetic studies of inhaled phenol would also

aid in the extrapolation from experimental animals to humans.

6.1.3. Cancer

Several epidemiology studies have evaluated the carcinogenesis of phenol, but they have not

found a consistent dose-related association. Because all of the subjects were also exposed to other

chemicals and there was no correction for smoking, these studies are not adequate to reach

conclusions on the carcinogenic potential of phenol. 

Phenol was negative in drinking water bioassays with rats and mice (NCI, 1980), although an

increased incidence of leukemias was observed in low-dose male rats. No inhalation studies of

sufficient duration to assess carcinogenicity were found. In short-term dermal assays, tumorigenicity

(production of papillomas in the absence of treatment by an initiating agent) was observed only at a

dose/concentration combination that produced ulceration and thus was well above the MTD (Salaman

and Glendenning, 1957). Similarly, although phenol was a promoter when tested in

initiation/promotion studies, the doses tested typically caused ulceration (Salaman and Glendenning,

1957; Boutwell and Bosch, 1959) and death (Boutwell and Bosch, 1959). There were two

exceptions. First, the low concentration tested by Salaman and Glendenning) caused promotion as

well as “transient light crusting.” Because the site of the weekly treatment was rotated across four

sites on the body, it is unclear whether more severe effects would have been observed if the same site

had been treated for the entire study. The second exception was that the low concentration tested by

Wynder and Hoffmann (1961) was reported as causing no toxicity, although the sensitivity of the

evaluation is unclear. On the basis of the high observed toxicity, it is not clear whether the promoting

activity observed for phenol in several studies was secondary to the repeated injury and healing of the

skin. From these considerations, the data regarding the carcinogenic potential of phenol are

inadequate for an assessment of human carcinogenic potential.

6.2. DOSE-RESPONSE

No human data that are adequate for the derivation of a phenol RfD were located. Therefore,

laboratory animal data were used.

The RfD of 0.3 mg/kg-day was based on a BMDL of 93 mg/kg-day for decreased maternal

body weight gain in a gavage rat developmental toxicity study that used a divided-dosing protocol
108

(Argus Research Laboratories, 1997). There was a corresponding NOAEL of 60 mg/kg-day and a

LOAEL for maternal toxicity of 120 mg/kg-day. A composite UF of 300 was used. This factor is

based on a default factor of 10 for extrapolation from laboratory animals to humans, a default factor

of 10 to account for intrahuman variability, and a factor of 3 to account for database insufficiencies.

Although the database for phenol can be considered “complete,” there are uncertainties

regarding the immunotoxicity potential of phenol in light of the immunotoxicity (decreased antibody

forming cells) reported by Hsieh et al. (1992). The database factor may be reconsidered with results

from an immunotoxicity study in mice that is compliant with EPA immunotoxicity test guidelines

(U.S. EPA, 1998c). This RfD is at least twice the endogenous rate of phenol formation in humans,

estimated as 0.014–0.14 mg/kg-day (Bone et al., 1976; Lawrie and Renwick, 1987; Renwick et al.,

1988), based on total phenol (free plus conjugated) levels in urine. This means that endogenous

production is approximately 5–50% of the RfD. 

 Note also that the RfD is meant to apply to ingested phenol in addition to the endogenous

formation of phenol. Endogenous phenol is produced by bacteria in the gut, so endogenous phenol

and ingested phenol would have similar toxicokinetics. Both humans and laboratory animals

efficiently conjugate and excrete phenol at low doses, resulting in only a small degree of systemic

exposure to free phenol (or any of its oxidative metabolites) at these low levels. The primary

difference between endogenous and exogenous phenol would result, because endogenous phenol is

formed in the intestines and some phenol may reach the colon and rectum, where some will escape

the hepatic portal circulation and be absorbed directly without conjugation. By contrast, a smaller

amount of ingested phenol would be expected to reach the colon and rectum.

The data are insufficient to determine the degree of conjugation of endogenously formed

phenol in humans because the available data are based on analysis of daily urinary excretion of total

phenol (i.e., phenol conjugates plus any trace amounts of free phenol) (Lawrie and Renwick, 1987;

Renwick et al., 1988); oxidative metabolites of phenol were not measured. The phenol conjugation

capacity of the liver is an important determinant of the ingested dose that would result in toxicity, but

there is no information on the degree of phenol conjugation by humans at doses in the range of the

RfD. Human variability exists in both the levels of endogenous phenol production and in the

conjugative capacity of the liver.

In the absence of more detailed information, it is reasonable to assume that humans have

adapted by having adequate conjugation capacity for the range of endogenous phenol production. 

Therefore, the default total uncertainty factor of 10 for human variability in toxicokinetics and
109

toxicodynamics is considered adequate. Determining whether oxidative metabolites are formed in

individuals who have high endogenous levels of phenol formation would enhance the confidence in

determining the intraspecies uncertainty factor.

The principal study (Argus Research Laboratories, 1997) used an adequate number of animals

and evaluated an appropriate array of endpoints for a developmental toxicity study. Although gavage

dosing was used, the divided-dosing protocol provided a significant enhancement that made the

gavage dosing more closely resemble an environmentally relevant route of exposure. The principal

study is judged to have medium confidence. Although the use of gavage dosing lowers the

confidence in the study, the dosing frequency may be fairly similar to that in drinking water studies,

in which rodents typically consumed water in a few larger doses and often in association with food

consumption. 

Confidence in the supporting database is medium to high. Although the oral toxicity database

meets the minimal criteria for a high-confidence database (chronic studies in two species,

developmental toxicity studies in two species, and a multigeneration reproduction study), the chronic

studies did not evaluate a sufficient array of endpoints. In particular, the chronic mouse study (NCI,

1980) did not evaluate hematological and immunological effects, making interpretation of the results

of the Hsieh et al. (1992) study difficult. Consideration of the above issues results in medium to high

confidence in the RfD.

The RfD developed in this document can be compared with other limits on phenol exposure,

partially as a test of the reasonableness of the RfD. Phenol is used in a number of industrial products,

as well as in over-the-counter medicines such as cough drops, throat sprays, and mouthwashes (e.g.,

Cepastat® and Chloraseptic® brands). Use of these consumer products can result in short-term, highlevel phenol exposures, but prolonged exposure (more than a week) at these levels is not

recommended. The short duration of exposures to the cough medicines suggests that safe exposure

levels would be higher than those for lifetime exposure. On the other hand, at least some aspects of

phenol toxicity appear to be related to peak concentrations in blood, and higher peak blood

concentrations could result from the consumer product exposure. 

ATSDR (1998) estimated that intake of the maximum recommended dosage of 300 mg

phenol/day would result in an approximate dose of 4–8 mg/kg-day. No documentation of this

maximum recommended dosage could be located. Use of these products on a daily basis over the

course of a lifetime would result in a dose approximately 10-fold higher than the RfD derived in this

document; however, these products are not intended for use over a prolonged period. Evaluation of
110

potential health effects in individuals who do consume these products in large amounts or over long

periods of time could provide additional information about human health effects of phenol and safe

exposure levels.

The use of a higher dose of phenol than the RfD in over-the-counter medicines suggests that

this RfD provides an adequate degree of public health protection. However, the maximum

recommended dosage may not be protective of pregnant women and fetuses, as consumers are

advised “as with any drug, if you are pregnant or nursing a baby, seek the advice of a medical

professional before using this drug.”

Although a substantial amount of data on phenol toxicokinetics are available, they are not

sufficient to move away from the default UFs for interspecies extrapolation and intraspecies

variability (IPCS, 2001). Data on how blood levels of phenol and its metabolites relate to doses 

in rats and humans would be useful in addressing the interspecies UF, as would data on the potential

for phenol to cross the placenta. Similarly, data on how differences in enzyme activities relate to

phenol and metabolite blood levels would be useful in addressing intrahuman variability. Finally, a

drinking water study compliant with EPA test guidelines (U.S. EPA, 1998c) that evaluated

hematological effects and immunological effects in mice could address the uncertainties associated

with the Hsieh et al. (1992) study and lead to reconsideration for the need for a database UF.

The available data are inadequate to derive an RfC. As noted above, a 90-day inhalation

study that evaluated the respiratory tract would be necessary for development of an RfC.

Because the data were considered inadequate to assess the carcinogenicity of phenol, no

quantitative assessment was conducted.
111

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APPENDIX A. SUMMARY OF EXTERNAL PEER REVIEW

COMMENTS AND DISPOSITION

The support document and IRIS summary for phenol have undergone both internal peer

review by scientists within EPA and a more formal external peer review by scientists outside EPA in

accordance with EPA guidance on peer review (U.S. EPA, 1998b, 2000a). Comments made by the

internal reviewers were addressed prior to submitting the documents for external peer review and are

not part of this appendix. The three external peer reviewers were tasked with providing written

answers to general questions on the overall assessment and on chemical-specific questions in areas of

scientific controversy or uncertainty. A summary of significant comments made by the external

reviewers and EPA’s response to these comments follows.

(1) General Comments

A. Data Presentation

All three reviewers agreed that the document is well organized. Two reviewers recommended

specific changes to increase the clarity of certain sections. 

Response to comments: The specific changes requested by the reviewers to clarify the text

were made. 

B. Are there additional data/studies that should be included?

One reviewer was not aware of any other data/studies that should be included. Another

reviewer recommended a developmental toxicity study but noted that it uses an unconventional study

design. A third reviewer criticized EPA for superficially exploring human health literature and

relying heavily on animal experiments. This third reviewer also suggested that EPA expand its

discussion of studies looking at phenol usage in over-the-counter drugs.

Response to comments: We reviewed the developmental toxicity study (Minor and Becker,

1971) recommended by one reviewer and concluded that it should not be included in the document

because it would not contribute any significant information to the database. The specific reasons for

the exclusion are: (1) this study used intraperitoneal (i.p.) dosing, a route of exposure of limited

relevance for human environmental exposure; because several well-conducted oral developmental

toxicity studies of phenol are available, it is not useful or necessary to supplement the database with

an i.p. study; (2) phenol was administered on gestation days 9–11 or days 12–14; such studies of

short windows of exposure can be useful for mechanistic purposes, but they are inadequate to fully

assess developmental toxicity; (3) this study is presented only as an abstract published 30 years ago,

and the full study was never published, raising questions about the reliability of the results, and in

addition, the abstract provides insufficient information for evaluation of the study. 

In response to the request for additional human data, additional reviews of literature databases

and secondary review articles were conducted, but no additional data were identified. Trade

associations were also contacted, with the same results. However, information relating the effects

observed in animals to effects observed in the available human studies was added to Section 4.5. Due

to limitations such as confounding exposures and uncertain exposure estimation, the available human
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data are useful for hazard identification but not for dose-response assessment. 

Manufacturers of over-the-counter drugs containing phenol, the relevant trade associations

were contacted, and reference books were consulted, in order to obtain information on the basis for

the recommended maximal dosage, but the information was not available.

C. For the RfD, has the most appropriate critical effect (decreased fetal body weight in the

NTP (1983a) study and decreased maternal weight gain in the Argus Research Laboratories

(1997) study) been chosen?

One reviewer agreed with the selection of decreased fetal body weight as the critical effect,

with a NOAEL of 60 mg/kg-day. This reviewer noted that although this is a relatively nonspecific

toxic effect, it could be appropriately considered adverse. In contrast, a second reviewer stated that

the reduction in body weight was a weak basis for the RfD because it was nonspecific and because of

other changes affecting body weight (e.g., increased litter size). On the basis of these considerations,

the reviewer did not consider the observed reduction in body weights to be an adverse effect and

suggested that the corresponding dose might be considered a NOAEL. Neither of these reviewers

commented on the appropriateness of the co-critical effect of decreased maternal weight gain

endpoint, also with a NOAEL of 60 mg/kg-day (and a BMDL of 93 mg/kg-day).

A third reviewer asserted that the critical effect occurs at a lower dose and that the RfD should

be based on a NOAEL lower than 60 mg/kg-day. The reviewer stated that the data from the NTP

(1983a) study are consistent with a LOAEL of 30 mg/kg-day, based on statistically significant

increases in the number of litters with nonlive fetuses in the low- and mid-dose groups and a trend

comparison of the fetal weight data. The reviewer also stated that data from the Argus Research

Laboratories (1997) study support the conclusion that 60 mg/kg-day is a LOAEL, based on an

increase in the percent litters with fetuses with alterations. The reviewer also recommended that the

LOAEL in the IIT Research Institute (1999) study should be 20 mg/kg-day (with no NOAEL

identified), based on decreased uterine weight and decreased prostate weight, rather than the NOAEL

of 70.0 mg/kg-day identified in the Toxicological Review. The reviewer asked for additional

discussion of the interpretation of these results but did not recommend a specific critical effect. This

reviewer also recognized the uncertainties in the Hsieh et al. (1992) study but suggested that

macrocytosis could explain the inconsistency in that study between erythrocyte count and hematocrit. 

Response to comments: The relationship between litter size and fetal weight was noted, and

the analysis of the NTP (1983a) study notes that 120 mg/kg-day is an equivocal LOAEL, with a

NOAEL of 60 mg/kg-day. Due to the uncertainties in the identification of the NOAEL/LOAEL

boundary for this study, the NTP (1983a) study was then judged not appropriate as a co-principal

study. If the NOAEL in the NTP (1983a) study were changed to 120 mg/kg-day, there would not be

a significant effect on the RfD because the Argus Research Laboratories (1997) study identified a

NOAEL of 60 mg/kg-day and a LOAEL of 120 mg/kg-day, with a BMDL of 93 mg/kg-day. In

addition, the NOAEL identified in the principal study (Argus Research Laboratories, 1997) is

supported by a NOAEL of 107 mg/kg-day in the ClinTrials BioResearch (1998) study, a value that is

very close to the BMDL used to derive the RfD.
126

Additional information about the litters with nonlive fetuses, decreased uterine weight, and

other findings mentioned by the third reviewer were added to the document (Section 4.3). The

former endpoint was not considered treatment-related due to the absence of a dose-response. The

data on percent litters with fetuses with alterations were independently analyzed using the chi-square

test, and no significant effect was observed. Text was also added to Section 6.1.1 regarding

uncertainties in the interpretation of the uterine weight data.

D. Has the noncancer assessment been based on the most appropriate studies?

Two of the reviewers agreed that the noncancer assessment is based on the most appropriate

studies, although, as indicated in the previous question, one of the reviewers raised some concerns

about the interpretation of those studies. Two of the reviewers also agreed with EPA’s assessment of

the deficiencies of the study by Hsieh et al. (1992). One reviewer agreed with the selection of

significant endpoints in the phenol document; however, this reviewer disagreed with the use of

gavage studies for the derivation of the RfD. This reviewer asserted that EPA should have used the

two-generation drinking water study by IIT Research Institute (1999) to derive the RfD. 

Response to comments: As described in the Toxicological Review, the effects at the LOAEL

(the high dose) of the IIT Research Institute (1999) study appear to be secondary to decreased water

consumption due to poor palatability, and so do not appropriately reflect phenol toxicity. Text was

added to the Toxicological Review noting that rats consume water only intermittently during the day,

and so a divided-dose gavage protocol is fairly similar to drinking water consumption.

E. For the noncancer (RfD) assessment, are there other data that should be considered in

developing uncertainty factors or the modifying factor? Do the data support the use of

different values from those proposed?

All three reviewers agreed with the uncertainty factors presented in the phenol document and

EPA’s rationale for selecting these factors.

Response to comments: None.

F. Do the confidence and weight-of-evidence statements present a clear rationale and

accurately reflect the utility of the studies chosen, the relevancy of the effects (cancer and

noncancer) to humans, and the comprehensiveness of the database? Do these statements make

sufficiently apparent all the underlying assumptions and limitations of these assessments?

The reviewers agreed that the confidence statements are carefully reasoned and clearly stated. 

Two of the reviewers agreed that the weight-of-evidence statements are appropriate, and a third

reviewer noted that no weight-of-evidence statement is used for noncarcinogenicity.

Response to comments: None needed.

(2) Chemical-Specific Comments

A. When endogenously produced phenol is taken into account, can the RfD be supported? 

Note that the RfD is applied to ingested phenol in addition to the normal daily endogenously
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produced phenol. Are there differences in endogenous phenol production between rats and

humans that should be taken into account in the development of the RfD?

The reviewers agreed that the RfD applied to ingested phenol in addition to the normal daily

endogenously produced phenol is appropriate. One reviewer noted that endogenous production is a

relatively small fraction (~2–20%) of the RfD. (This fraction became 5–50% after addition of the

database uncertainty factor during consensus review.) Another reviewer noted that both the

experimental animals and humans would have similar baseline levels of endogenous phenol

production.

Response to comments: Additional supporting information provided by the reviewers was

incorporated into the text. Toxicokinetic considerations regarding quantitatively accounting for

endogenous production were also incorporated into the text.

B. Do you agree/disagree with the recommendation that there are not sufficient data to

generate a scientifically defensible RfC and cancer slope factor?

Two reviewers agreed that data are insufficient to generate an RfC and cancer slope factor. 

One of these reviewers also commented on statements made by EPA related to red nasal discharge

identified in study animals. This reviewer stated that the secretion/discharge is not directly

suggestive of, or a precursor to, a nasal or ocular lesion but is simply the animal’s response to stress. 

A third reviewer did not comment on this question because it was beyond this person’s area of

expertise.

Response to comments: The information on red nasal discharge provided by the reviewer

was incorporated into the text.

C. Was the interpretation of the decreased fetal body weight in rats in the National Toxicology

Program (NTP) study (NTP, 1983a) appropriate?

One reviewer raised some issues regarding the adversity of the effect but generally agreed

with EPA’s decision to designate the decreased fetal body weight finding at 120 mg/kg-day as an

equivocal LOAEL, resulting in a NOAEL of 60 mg/kg-day. A second reviewer disagreed that

decreased fetal body weight is an adverse effect and asserted that 120 mg/kg-day should be classified

as a NOAEL instead. A third reviewer supported interpreting the data using trend analysis rather

than pairwise comparisons, resulting in a LOAEL of 30 mg/kg-day for fetal body weight reduction.

Response to comments: On the basis of the weight of evidence, the weight of the reviewers’

comments, and the supporting data from the Argus Research Laboratories (1997) study, the high dose

in the NTP (1983a) study was retained as an equivocal LOAEL, but the study was removed from

being a co-principal study for the derivation of the RfD. Although there was a dose-response trend at

the low and mid doses, the decreases in the fetal body weight at these doses were marginal (1%–2%)

and were not considered biologically significant. Only the response (7%) at the high dose was

significantly different from the control. A meaningful benchmark dose could not be calculated for

these data, because fetal body weight by sex were not in the published study and not available in
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NTP’s archives.

D. Please comment on the choice of gavage developmental toxicity studies as the co-critical

studies in light of the differences between phenol toxicity when administered in drinking water

and by gavage.

One reviewer agreed with EPA’s willingness to use gavage studies but suggested that the IIT

Research Institute (1999) drinking water study represents a more relevant exposure scenario and

supports a lower LOAEL, based on decreased absolute and relative uterine weight. A second

reviewer asserted that the drinking water study is more appropriate because a divided gavage dose is

not equivalent to more extended intake during ad libitum water consumption. A third reviewer

supported the use of the gavage study and asserted that the divided dose administered in the Argus

Research Laboratories study (1997) is not unrealistic because animals in drinking water studies tend

to drink when they eat, not continuously throughout the day. This reviewer believed that the NOAEL

and the LOAEL are not overly conservative.

Response to comments: Additional information about the interpretation of the decreased

uterine weight in the IIT Research Institute (1999) drinking water study and associated uncertainties

was added Section 4.3. In particular, even though the decrease in relative uterine weight was

statistically significant at all dose levels, there was no dose-response. Information was also added to

the text noting that rats drink water in a small number of periods during the day rather than

continuously through the day, so a divided-dose gavage study is fairly similar to drinking water

exposure.

E. Was the interpretation of decreased motor activity in the 13-week oral neurotoxicity study

appropriate?

One reviewer agreed with EPA’s interpretation of decreased motor activity in the 13-week

oral neurotoxicity study. A second reviewer found the interpretation difficult to comment on, given

the confounded results of the study, and a third reviewer did not respond to this question because it

was beyond this person’s area of expertise.

Response to comments: The issues potentially confounding the neurotoxicity were noted in

the document. 

OVERALL RECOMMENDATION

All three reviewers stated that the document is acceptable with revisions.

New Reference:

Minor, JL; Becker, BA. (1971). A comparison of the teratogenic properties of sodium salicylate,

sodium benzoate, and phenol. Toxicol Appl Pharmacol 19:373.
129

Appendix B. Benchmark Dose Modeling Results

Benchmark dose (BMD) modeling was performed to identify potential critical effect levels for

derivation of the RfD for phenol. The modeling was conducted according to draft EPA guidelines

(U.S. EPA, 2000d) using Benchmark Dose Software Version 1.3 (BMDS), which is available from

EPA (U.S. EPA, 2001). The BMD modeling results are summarized in Table B-1, and the output is

attached as Appendix C. A brief discussion of the modeling results for each endpoint is presented

below.

Because all the following endpoints are continuous variables, the continuous models available

with BMDS (power, polynomial, and Hill models) were used. The hybrid model was not used,

because the hybrid model software in BMDS is still undergoing Beta-testing and was not considered

sufficiently validated to provide a BMDL as the basis for the quantitative dose-response assessment. 

(The hybrid modeling approach defines the benchmark response [BMR] directly in terms of risk,

whereas the standard approach, defines the BMR in terms of a change in the mean.) For all of the

modeling conducted, the BMR was defined as a 1.0 SD change in the mean because this is the default

measure recommended by the EPA (U.S. EPA, 2000d) in the absence of a clear biological rationale

for selecting an alternative response level.

Argus Research Laboratories, 1997 

Two endpoints were modeled from this study: decreased maternal body weight gain and the

related endpoint of decreased maternal body weight. The decrease in weight gain was the more

sensitive endpoint, with a NOAEL of 60 mg/kg/day and a LOAEL of 120 mg/kg/day. 

As summarized in Table B-1, the BMD and BMDL estimates for the endpoint of decreased

maternal body weight gain was similar for all three models. The model fit was also generally similar. 

A visual analysis of the data fit in the regions of the BMDLs indicated that the data fitting from the

three model was adequate and comparable across all models. The goodness-of-fit p-values calculated 

for the power and polynomial models were very similar, but no p-value could be computed for the

Hill model because there were no degrees of freedom for the calculation. (This was because the

number of parameters included in the model was equal to the number of data points.) The Akaike

Information Criterion (AIC), a measure of goodness of fit that takes into account the number of

degrees of freedom, was very similar for all three models but was marginally better (i.e., lower) for

the polynomial model. The polynomial model was chosen as the basis for the BMDL for this

endpoint, based on the slightly better fit and as a slightly more health-protective value. An

alternative BMDL for this endpoint could be calculated using the geometric mean of the BMDLs

from all three models, 114 mg/kg-day, based on the rationale that the small difference in AIC

observed was not meaningful. 

For the endpoint of decreased maternal body weight, all three continuous models gave a

similar BMD estimate and provided adequate data fits, with goodness-of-fit p-values larger than 0.1. 

However, the best data fit was obtained with the polynomial model, which had a p-value of 0.92,

compared with a p-value of 0.30 obtained with the power model. The p-value for the Hill model was

not computed due to insufficient degrees of freedom. A visual analysis of the data fit in the region of
130

the BMDLs also indicated that good fit was obtained with all three models, with the best fit obtained

using the polynomial model. In addition, the AIC analysis also indicated that the best data fit was

obtained with the polynomial model. Comparable BMDLs of 143 and 147 mg/kg/day were obtained

using the polynomial and Hill models, respectively. Thus, the BMDL chosen for this endpoint was

143 mg/kg-day, obtained with the polynomial model, with a corresponding BMD of 345 mg/kg. This

BMDL is higher than the BMDL for maternal body weight gain, the other endpoint modeled for this

study.

ClinTrials BioResearch (1998) 

Only one endpoint from this study was modeled: decreased total activity counts in a motor

activity assay in females at week 4. Acceptable fits (p=0.35 and p=0.17) were obtained with the

polynomial and power models, respectively. The visual fit for both models in the region of the

BMDL was acceptable but not as good as would be desired. Much better fit was obtained with the

Hill model, based on visual fit, but no BMDL could be calculated for this model. The same BMDL

of 219 mg/kg-day was calculated with both the power and polynomical models and was chosen as the

study BMDL.

Hsieh et al. (1992) 

Three related endpoints were modeled for this study: plaque-forming cells, plaque-forming

cells/106

 spleen cells, and antibody titer. The study NOAEL was 6.2 mg/kg-day, with a LOAEL of

33.6 mg/kg-day. 

As summarized in Table B-1, unacceptable fits were obtained with the power and polynomial

models for the endpoint of plaque-forming cells/106

 spleen cells. No p-value could be obtained for

the Hill model due to insufficient degrees of freedom, but a visual analysis of the results indicated

that the Hill model provided an acceptable fit. Based on the Hill model, the BMD and BMDL for

decreased plaque-forming cells/106

 spleen cells were 1.26 mg/kg-day and 0.38 mg/kg-day,

respectively.

Similar results were obtained for plaque forming cells/spleen. Inadequate fits were obtained

with the power and polynomial models, and the Hill model provided no p-value at all due to

insufficient degrees of freedom. A visual analysis of results indicated that the Hill model provided an

overall adequate fit, but not a very good fit in the range of the BMD. In addition, this model failed to

estimate BMD or BMDL. Therefore, no BMD and BMDL can be identified for this particular

endpoint.

For the endpoint of decreased antibody titer, the power and polynomial models both had

marginal fit, based on the goodness-of-fit p-values; visual inspection of the data indicated that these

models had inadequate fit. No p-value could be calculated for the Hill model, due to insufficient

degrees of freedom, but a visual analysis of the model results indicated an adequate fit. Therefore,

based on the Hill model, the BMD and BMDL for decreased antibody titer were 3.51 mg/kg-day and

0.73 mg/kg-day, respectively.

The lowest BMDL for this study was 0.38 mg/kg-day, calculated for plaque-forming cells/106
131

spleen cells. However, this BMDL is not used for risk assessment due to uncertainties in the

appropriate BMR for this in vivo/in vitro study design.

Table B-1. Benchmark Dose Modeling Results for Phenol

Model p-value AIC BMDa BMDLa

Argus Research Laboratories (1997)

Maternal Body Weight Gain

Power 0.3165 545 152 125

Polynomial 0.3191 543 157 93

Hill N/A 545 151 129

Maternal Body Weight

Power 0.3013 731 354 244

Polynomial 0.9188 729 345 143

Hill N/Ab 733 345 147

ClinTrials BioResearch Ltd. (1998)

Motor Activity - Total Activity Counts in Females in Week 4

Power 0.1701 629 337 219

Polynomial 0.3477 625 336 219

Hill N/A 630 246 ---

Hsieh et al., 1992

Plaque-Forming Cells

Power 0.008 236 15.7 10.5

Polynomial 0.003 227 14.3 9.7

Hill NA 229 1.26 0.38

Plaque-Forming Cells/Total Spleen

Power 0.054 469 23 15.3

Polynomial <0.0001 509 23.4 5.87

Hill N/A 467 6.14 --- c
132

Antibody Titer

Power 0.102 -66.2 21.4 13.1

Polynomial 0.102 -70.2 21.4 13.1

Hill N/A -66.9 3.51 0.73

a. BMD and BMDL are based on benchmark response of 1.0 SD. Results are presented in units of mg/kg/day.

b. NA: the information is not available because there are insufficient degrees of freedom for the test

c. ---: failed to estimated this value.
133

Appendix C. Benchmark Dose Modeling Output