NTP-CERHR EXPERT PANEL REPORT 

on

Di-n-Butyl Phthalate

October 2000 NTP-CERHR-DBP-00

Center For The Evaluation Of Risks 

To Human Reproduction

National Toxicology Program

U.S. Department of Health and Human Services

NTP-CERHR Monograph on the 

Potential Human Reproductive and 

Developmental Effects 

of Di-n-Butyl Phthalate (DBP)
Table of Contents

Preface.............................................................................................................................................. i

Introduction..................................................................................................................................... ii

NTP Brief on Di-n-Butyl Phthalate (DBP)......................................................................................5

References........................................................................................................................................6

Appendix I. NTP-CERHR Phthalates Expert Panel 

Preface...............................................................................................................................I-1

Expert Panel......................................................................................................................I-2

Appendix II. Phthalates Expert Panel Report on DBP

Preface.............................................................................................................................. II-i

Chemistry, Usage and Exposure ......................................................................................II-1

General Toxicological and Biological Parameters .........................................................II-6

Developmental Toxicity Data .........................................................................................II-2

Reproductive Toxicity...................................................................................................II-20

Data Summary & Integration........................................................................................II-25

References.....................................................................................................................II-39

Tables............................................................................................................................II-44

Appendix III. Public Comments on the Phthalates Expert Panel Reports

AdvaMed........................................................................................................................ III-1

American Chemistry Council (12-7-2000).................................................................... III-5

American Chemistry Council (12-11-2000) .................................................................. III-7

American Chemistry Council (4-13-2001).................................................................. III-58

Discovery Medical, Inc................................................................................................ III-66

Environmental Working Group (11-3-2000) ............................................................... III-67

Environmental Working Group (12-8-2000) ............................................................... III-69

William Faber .............................................................................................................. III-71

Healthy Environments & Product Safety Branch ........................................................ III-81

Health Care Without Harm .......................................................................................... III-83

Beverly Smith .............................................................................................................. III-87

Swedish Chemical Inspection Agency......................................................................... III-89
i

The National Toxicology Program (NTP) established the NTP Center for the Evaluation 

of Risks to Human Reproduction (CERHR) 

in 1998. The CERHR is a publicly accessible 

resource for information about adverse reproductive and/or developmental health effects 

associated with exposure to environmental 

and/or occupational chemicals. The CERHR 

is located at the National Institute of Environmental Health Sciences (NIEHS) of the 

National Institutes of Health and Dr. Michael 

Shelby is the director.1

The CERHR broadly solicits nominations of 

chemicals for evaluation from the public and 

private sectors. The CERHR follows a formal 

process for review and evaluation of nominated 

chemicals that includes multiple opportunities 

for public comment. Chemicals are selected for 

evaluation based upon several factors including the following: 

• potential for human exposure from use 

and occurrence in the environment.

• extent of public concern.

• production volume.

• availability of scientific evidence for 

reproductive and/or developmental toxicity. 

The CERHR convenes a scientific expert panel 

that meets in a public forum to review, discuss, 

and evaluate the scientific literature on the selected chemical. Public comment is invited prior to and during the meeting. The expert panel 

produces a report on the chemical’s reproductive and developmental toxicities and provides 

its opinion of the degree to which exposure 

to the chemical is hazardous to humans. The 

panel also identifies areas of uncertainty and 

where additional data are needed. The CERHR 

expert panels use explicit guidelines to evaluate the scientific literature and prepare the 

expert panel reports. Expert panel reports are 

made public and comments are solicited. 

Next, the CERHR prepares the NTP-CERHR 

monograph. The NTP-CERHR monograph includes the NTP brief on the chemical evaluated, 

the expert panel report, and all public comments. The goal of the NTP brief is to provide 

the public, as well as government health, regulatory, and research agencies, with the NTP’s 

interpretation of the potential for the chemical 

to adversely affect human reproductive health 

or children’s health. The NTP-CERHR monograph is made publicly available electronically 

on the CERHR website and in hard copy or 

CD-ROM from the CERHR.

Preface

1 Information about the CERHR is available on the 

web at <http://cerhr.niehs.nih.gov> or by contacting the director:

P.O. Box 12233, MD EC-32, NIEHS, 

Research Triangle Park, NC 27709 

919-541-3455 [phone] 

919-316-4511 [fax]

shelby@niehs.nih.gov [email] 

Information about the NTP is available on the web 

at <http://ntp-server.niehs.nih.gov> or by contacting the NTP Office of Liaison and Scientific Review at the NIEHS:

liaison@starbase.niehs.nih.gov [email]

919-541-0530 [phone]
ii 

In 1999, the CERHR Core Committee, an advisory committee composed of representatives 

from NTP member agencies, recommended 

seven phthalates for expert panel review. 

These chemicals were selected because: 

(a) there is the potential for human exposure 

from their widespread use and occurrence within the environment, 

(b) they have a high production volume, 

(c) there is substantial scientific literature 

addressing the reproductive and/or developmental toxicities of these chemicals, and 

(d) they are of concern to the public. 

These seven phthalates are as follows:

• di(2-ethylhexyl)phthalate (DEHP) 

• di-isononyl phthalate (DINP) 

• di-isodecyl phthalate (DIDP)

• di-n-butyl phthalate (DBP) 

• butyl benzyl phthalate (BBP) 

• di-n-octyl phthalate (DnOP)

• di-n-hexyl phthalate (DnHP)

Phthalates are a group of similar chemicals 

widely used to soften and increase the flexibility of plastic consumer products such as 

shower curtains, medical devices, upholstery, 

raincoats, and soft squeeze toys. They are not 

bound to the plastics and can leach into the surrounding environment. DEHP has the greatest 

production volume of the selected phthalates 

(approximately 260 million pounds [1994]), 

followed by DIDP (approximately 240 million pounds [1994]), and DINP (approximately 

215 million pounds [1994]). The scientific literature on the reproductive and developmental 

toxicities of several phthalates is extensive. In 

addition, there is widespread public concern 

about the safety of phthalates. 

As part of the evaluation of phthalates, the 

CERHR convened a panel of scientific experts 

(Appendix I) to review, discuss, and evaluate 

the scientific evidence on the potential reproductive and developmental toxicities of each 

phthalate. There were three public meetings 

of this panel (August 17-19 and December 15-

17, 1999 and July 12-13, 2000). The CERHR 

received numerous public comments on the 

phthalates throughout the evaluation process. 

The NTP has prepared an NTP-CERHR monograph for each phthalate. This monograph 

includes the NTP brief on DBP, a list of the 

expert panel members (Appendix I), the expert 

panel’s report on DBP (Appendix II), and all 

public comments received on the expert panel’s 

reports on phthalates (Appendix III). The NTPCERHR monograph is intended to serve as a 

single, collective source of information on the 

potential for DBP to adversely affect human 

reproduction or development. Those interested 

in reading this report may include individuals, 

members of public interest groups, and staff of 

health and regulatory agencies. 

The NTP brief included within this report presents the NTP’s interpretation of the potential 

for exposure to DBP to cause adverse reproductive or developmental effects in people. It is 

based upon information about DBP provided in 

the expert panel report, the public comments, 

and additional scientific information available 

since the expert panel meetings. The NTP brief 

is intended to provide clear, balanced, scientifically sound information on the potential for 

DBP exposures to result in adverse health effects on development and reproduction. 

Introduction
iii

While there are biological and practical reasons for considering developmental toxicity 

and reproductive toxicity as 2 separate issues, it is important to keep in mind that life 

in mammals, including humans, is a cycle. 

In brief, the cycle includes the production 

of sperm and eggs, fertilization, prenatal development of the offspring, birth, post-natal 

development, sexual maturity, and, again, 

production of sperm and eggs. 

In the past, toxic effects were often studied in a “life stage specific” manner. Thus, 

concerns for developmental toxicity were 

addressed by exposing pregnant mothers 

and looking for adverse effects in fetuses. 

Developmental toxicity was detected as 

death, structural malformations, or reduced 

weights of the fetuses just prior to birth. Reproductive toxicity was studied by exposing 

sexually mature adults to the chemical of interest and effects were detected as impaired 

capacity to reproduce. Over the years, toxicologists realized that exposure during one 

part of the life cycle could lead to adverse 

effects that might only be apparent at a different part of the life cycle. For example, 

exposure of a sexually mature individual to 

an agent capable of inducing genetic damage in eggs or sperm might have no apparent 

effect on the exposed individual. However, 

if a genetically damaged egg or sperm from 

that individual is involved in fertilization, 

the induced genetic damage might lead to 

death or a genetic disorder in the offspring. 

In this example, chemical-induced damage 

is detected in the next generation. In contrast, the reproductive system begins developing well before birth and continues until 

sexual maturity is attained. Thus, exposure 

of sexually immature animals, either before 

or following birth, to agents or conditions 

that adversely affect development of the 

reproductive system can result in structural 

or functional reproductive disorders. These 

effects may only become apparent after the 

exposed individual reaches the age of puberty or sexual maturity.

Thus, in the case of genetic damage induced 

in eggs or sperm, what might be considered 

reproductive toxicity gives rise to developmental disorders. Conversely, in the case 

of adverse effects on development of the 

reproductive tract, developmental toxicity 

results in reproductive disorders. In both 

these examples it is difficult to make a clear 

distinction between developmental and reproductive toxicity. This issue is important 

in considering the phthalate evaluations 

because evidence of developmental toxicity affecting reproductive capacity in later 

stages of the life cycle is reported for at least 

3 of the phthalates - BBP, DBP, and DEHP.

Developmental Toxicity versus 

Reproductive Toxicity
1

NTP Brief

What is DBP?

DBP is a clear, oily liquid with the chemical 

formula C16H22O4 and the structure shown 

in Figure 1. It is one of a group of industrially important chemicals known as phthalates. 

Phthalates are used primarily as plasticizers to 

add flexibility to plastics. Unlike many phthalates, DBP is not currently used as a plasticizer 

in polyvinyl chloride (PVC) plastics. Typically, 

DBP is used as a component of latex adhesives. 

It is also used in cosmetics and other personal 

care products, as a plasticizer in cellulose plastics, and as a solvent for dyes. 

DBP is produced by reacting n-butanol with 

phthalic anhydride. The most recent information available indicates that approximately 7.7 

million kilograms (17 million pounds) of DBP 

were produced in the United States in 1994 

(ATSDR, 2001). 

Are People Exposed to DBP?

Yes. Blount et al. (2000) reported that more 

than 75% of the people they studied were exposed to DBP. There are several ways that people may be exposed to DBP at home or at work. 

Human exposure to DBP can occur during the 

manufacture of DBP, during the manufacture 

of DBP-containing products such as latex adhesives, cellulose-based plastics, during the 

use of such products, or through the presence 

of DBP in the environment.

Environmental exposures can occur through 

air, water, or food. Most people are exposed to 

DBP primarily through food. DBP migrates into 

foods, particularly fatty foods, from DBP-containing materials that are used to process and 

package food. Possible, sources of DBP found in 

food are latex adhesives used in food processing 

equipment and food wraps made of cellulosebased plastics. Cosmetics and other personal 

care products may be another important source 

of exposure through inhalation or contact with 

the skin. Studies to determine the extent of such 

exposures have not been conducted.

The expert panel estimated that the U.S. general population is exposed to approximately 2-10 

μg/kg bw/day (micrograms per kilogram body 

weight per day). This reflects a total daily exposure of approximately 140-700 μg per person 

per day. By comparison, a small drop of water 

weighs approximately 30,000 μg and a grain of 

table salt weighs approximately 60 μg. 

A recent study not available to the expert panel 

determined the amount of DBP metabolites in 

human urine (Blount et al., 2000). Kohn et al. 

(2000) and David (2000) used the data from 

that study to estimate daily exposure levels of 

DBP. Kohn et al. estimated that 95% of people 

are exposed to less than 10 μg DBP/kg bw/day, 

consistent with the expert panel’s estimate. 

However, they found that some women of 

reproductive age (20-40 years) are exposed to 

higher DBP levels than other age or sex groups. 

The majority of women in the age group were 

exposed to DBP levels well within the range of 

exposures estimated by the expert panel.

However, a small percentage was exposed to 

NTP Brief on Di-n-butyl Phthalate (DBP)

O

O

O

O

Figure 1. Chemical structure of DBP
2

NTP Brief

30 μg/kg bw/day or greater and one individual 

was exposed to over 100 μg/kg bw/day. Neither 

the sources nor circumstances of these apparently higher exposures are known. It has been 

suggested that these higher exposures might be 

related to the use of DBP-containing personal 

care products such as perfumes, nail polishes, 

and hair spray (Blount et al., 2000). 

Workers producing DBP can be exposed 

through skin contact or inhalation. The expert 

panel estimated that such exposures might be 

as high as 143 μg/kg bw/day, but are generally 

thought to be well below this level. Information is not available on exposure of workers 

who use DBP to manufacture other products.

Can DBP Affect Human Development or 

Reproduction?

Probably. Although there is no direct evidence 

that exposure of people to DBP adversely affects reproduction or development, studies 

with laboratory rodents show that exposure to 

DBP can cause adverse effects (Fig. 2). Based 

on recent data on the extent to which humans 

absorb, metabolize and excrete DBP, the NTP 

believes it is reasonable and prudent to conclude that the results reported in laboratory 

animals indicate a potential for similar or other 

adverse effects in humans. 

Scientific decisions concerning health risks are 

generally based on what is known as “weightof-the-evidence.” In this case, recognizing the 

lack of human data and the clear evidence of 

effects in laboratory animals (Fig. 2), the NTP 

judges the scientific evidence sufficient to conclude that DBP may adversely affect human 

reproduction or development if exposures are 

sufficiently high. 

Summary of Supporting Evidence 

As presented in the expert panel report (see 

report for details and literature citations), 

many of the DBP studies in rodents addressed 

both developmental and reproductive effects. 

These studies have reported that exposure of 

pregnant dams to high doses of DBP (greater 

than 500,000 μg/kg bw/day) causes reduced 

fetal survival and reduced birth weights among 

surviving offspring. In some instances, this 

exposure was also associated with skeletal 

malformations and abnormalities of the reproductive systems and organs in both male and 

female offspring. Exposure to DBP has also 

been shown to reduce fertility in female rats 

and mice. It is clear from studies with laboratory animals that rodents in prenatal and early 

postnatal stages of development are more sensitive to the reproductive effects of DBP than 

are adult animals. It is important to note that 

Clear evidence of adverse effects

Some evidence of adverse effects

Limited evidence of adverse effects

Insufficient evidence for a conclusion

Limited evidence of no adverse effects

Some evidence of no adverse effects

Clear evidence of no adverse effects

Developmental and Reproductive Toxicity

Figure 2. The weight of evidence that DBP causes adverse developmental or 

reproductive effects in laboratory animals 
3

NTP Brief

DBP exposure levels that lead to these adverse 

effects in rodents are generally far higher than 

those experienced by people.

The developing male reproductive system of 

rodents appears particularly sensitive to the adverse effects of DBP exposure. There is growing evidence that this male sensitivity may 

result from a reduction in the level of the male 

sex hormone, testosterone, by DBP.

In a study published after the expert panel report 

was completed, Shono et al. (2000) showed that 

exposure to the monoester metabolite of DBP, 

monobutyl phthalate (MBP), is toxic to the 

male reproductive tract. Pregnant dams were 

given an oral dose of 300,000 μg/day of MBP 

on various days during pregnancy; fetuses were 

obtained by Caesarean section. 

On gestation day 20, significant inhibition of 

testis migration was reported for male fetuses 

exposed to MBP on gestation days 11-14 and 

or 15-18, with the greatest inhibition observed 

in the latter group. There were also treatmentrelated effects on the male reproductive tract 

along with a reduction in testicular testosterone 

levels. This study supports the role of MBP in 

mediating DBP toxicity to the male reproductive tract.

Another recent report by Foster et al. (2000) 

proposes that the use of rat data to assess human 

risks for reproductive or developmental effects 

may be inappropriate because humans might be 

much less efficient at producing the active DBP 

metabolite, MBP. However, a recent study (Anderson et al., 2001) supports using DBP rodent 

data for evaluating potential effects in humans. 

It offers evidence that people efficiently absorb 

and metabolize DBP. The results show that human volunteers given an oral dose of 255 or 

510 μg DBP excrete approximately 70% of it 

as MBP in urine after 24 hours. 

Are Current Exposures to DBP High Enough 

to Cause Concern?

Possibly. More data are needed to better un1 Based on Kohn et al. (2000) estimated exposure of some women of reproductive 

age to ~100 μg/kg bw/day (median 1.7; 95th percentile, 32; maximum, 113)

2 Based on the experts panel’s estimate of general population exposure from

2-10 μg/kg bw/day

Figure 3. NTP conclusions regarding the possibilities that human development 

or reproduction might be adversely affected by exposure to DBP 

Reproductive effects in adults2

Developmental effects2

Serious concern for adverse effects

Concern for adverse effects

Some concern for adverse effects

Minimal concern for adverse effects

Negligible concern for adverse effects

Insufficient hazard and/or exposure data

Developmental effects at high exposures1
4

NTP Brief

derstand human DBP exposure levels and how 

these exposures vary across the population. The 

general U.S. population presently appears to be 

exposed to DBP at levels that are not of immediate concern for causing adverse reproductive 

or developmental effects. However, data are 

not available to permit conclusions regarding 

the possibility of effects in various age groups, 

occupations, or socioeconomic strata. Thus, the 

NTP offers the following conclusions (Fig. 3).

The NTP concurs with the CERHR Phthalates Expert Panel that there is minimal concern for developmental effects when pregnant 

women are exposed to DBP levels estimated 

by the Panel (2-10 μg/kg bw/day).

Based upon recent estimated DBP exposures 

among some women of reproductive age, the 

NTP has some concern for DBP causing adverse effects to human development, particularly development of the male reproductive 

system.

This level of concern is greater than that expressed by the Phthalates Expert Panel and is 

based on recent estimates that some women of 

childbearing age are exposed to levels of DBP 

that are approximately 10 times higher than 

general population exposures.

The NTP concurs with the CERHR Phthalates 

Expert Panel that there is negligible concern 

for reproductive toxicity in exposed adults.

However, further data and evaluation are needed to determine if the higher DBP exposure 

levels reported for some women of reproductive age justify a higher level of concern for 

effects on their reproductive system.

These conclusions are based on the information available at the 

time this brief was prepared. As new information on toxicity and 

exposure accumulate, it may form the basis for either lowering or 

raising the levels of concern expressed in the conclusions.
5

References:

NTP Brief

Anderson WAC, Castle L, Scotter MJ, Massey 

RC, Springall C. A biomarker approach to mea-

suring human dietary exposure to certain 

phthalate diesters. Food Additives & Contami-

nants 18:1068-1074 (2001).

ATSDR Toxicological Profile for DI-N-BU-

TYL PHTHALATE (Update), September 

(2001).

Blount BC, Silva MJ, Caudill SP, Needham LL, 

Pirkle JL, Sampson EJ, Lucier GW, Jackson 

RJ, Brock JW. Levels of seven urinary phthal-

ate metabolites in a human reference popula-

tion. Environmental Health Perspectives 108: 

979-982 (2000).

David RM. Exposure to phthalate esters. En-

vironmental Health Perspectives 108: A440 

(2000).

Foster PMD, Cattley RC, Mylchreest E. Effects of di-n-butyl phthalate (DBP) on male 

reproductive development in the rat: Implica-

tions for human risk assessment. Food and 

Chemical Toxicology 38: S97-S99 (2000).

Kohn MC, Parham F, Masten SA, Portier CJ, 

Shelby MD, Brock JW, Needham LL. Human 

exposure estimates for phthalates. Environ-

mental Health Perspectives 108: A440-A442 

(2000).

Shono T, Kai H, Suita S, Nawata H. Time-spe-

cific effects of monon- butyl phthalate on the 

transabdominal descent of the testis in rat fe-

tuses. BJU International 86: 121-125 (2000).
I-1

Appendix I

Appendix I. NTP-CERHR Phthalates 

Expert Panel Report on DBP

A 16-member panel of scientists covering dis-

ciplines such as toxicology, epidemiology, and 

medicine was recommended by the Core Com-

mittee and approved by the Associate Director 

of the National Toxicology Program. Over the 

course of a 16-month period, the panel criti-

cally reviewed more than 500 documents on 7 

phthalates and identified key studies and issues 

for plenary discussions. At three public meet-

ings

2

, the expert panel discussed these studies, 

the adequacy of available data, and identified 

data needed to improve future assessments. At 

the final meeting, the expert panel reached con-

clusions on whether estimated exposures may 

result in adverse effects on human reproduction 

or development. Panel assessments were based 

on the scientific evidence available at the time 

of the final meeting. The expert panel reports 

were made available for public comment on 

October 10, 2000, and the deadline for public 

comments was December 11, 2000 (Federal 

Register 65:196 [10 Oct. 2000] p60206). The 

Phthalates Expert Panel Report on DBP is 

provided in Appendix II and the public com-

ments received on that report are in Appendix 

III. Input from the public and interested groups 

throughout the panel’s deliberations was in-

valuable in helping to assure completeness and 

accuracy of the reports. The Phthalates Expert 

Panel Reports are also available on the CERHR 

website (http://cerhr.niehs.nih.gov). 2

Phthalate Expert Panel meeting dates were: 

August 17-19, 1999, in Alexandria, VA; December 

15-17, 1999, in Research Triangle Park, NC; and 

July 12-13, 2000, in Arlington, VA. 
I-2

Appendix I

Robert Kavlock, Ph.D. (chair)

EPA/ORD

Research Triangle Park, NC

Kim Boekelheide, M.D., Ph.D.

Brown University

Providence, RI

Robert Chapin, Ph.D.

NIEHS

Research Triangle Park, NC

Michael Cunningham, Ph.D.

NIEHS

Research Triangle Park, NC

Elaine Faustman, Ph.D.

University of Washington

Seattle, WA

Paul Foster, Ph.D.

Chemical Industry Institute of 

Toxicology, Research Triangle Park, NC

Mari Golub, Ph.D.

Cal/EPA

Davis, CA

Rogene Henderson, Ph.D.

Inhalation Toxicology Research Institute

Albuquerque, NM

Irwin Hinberg, Ph.D.

Health Canada

Ottawa, Ontario, Canada

Ruth Little, Sc.D.

NIEHS

Research Triangle Park, NC

Jennifer Seed, Ph.D.

EPA/OPPT

Washington, DC

Katherine Shea, M.D.

North Carolina State University

Raleigh, NC

Sonia Tabacova, M.D., Ph.D.

FDA

Rockville, MD

Shelley Tyl, Ph.D.

Research Triangle Institute

Research Triangle Park, NC

Paige Williams, Ph.D.

Harvard University

Cambridge, MA

Tim Zacharewski, Ph.D.

Michigan State University, East 

Lansing, MI

Appendix I. NTP-CERHR Phthalates Expert Panel 

(Name and Affiliation)
NTP-CERHR EXPERT PANEL REPORT 

on

Di-n-Butyl Phthalate

October 2000 NTP-CERHR-DBP-00

Center For The Evaluation Of Risks 

To Human Reproduction

National Toxicology Program

U.S. Department of Health and Human Services

Appendix II
TABLE OF CONTENTS

1.0 CHEMISTRY, USAGE, AND EXPOSURE .................................................................................................1

1.1 Chemistry ....................................................................................................................................1

1.2 Exposure and Usage ....................................................................................................................1

2.0 GENERAL TOXICOLOGICAL AND BIOLOGICAL PARAMETERS .....................................................6

2.1 General Toxicity..........................................................................................................................6

2.1.1 Human Data ......................................................................................................................6

2.1.2 Experimental Animal Data................................................................................................6

2.2 Toxicokinetics .............................................................................................................................8

2.3 Genetic Toxicity ........................................................................................................................11

3.0 DEVELOPMENTAL TOXICITY DATA....................................................................................................12

3.1 Human Data...............................................................................................................................12

3.2 Experimental Animal Toxicity ..................................................................................................12

3.2.1 Prenatal Development.....................................................................................................12

3.2.2 Postnatal Development ...................................................................................................14

4.0 REPRODUCTIVE TOXICITY...................................................................................................................20

4.1 Human Data...............................................................................................................................20

4.2 Experimental Animal Toxicity ..................................................................................................20

5.0 DATA SUMMARY & INTEGRATION......................................................................................................25

5.1 Summary .............................................................................................................................25

5.1.1 Human Exposure.............................................................................................................25

5.1.1.1 Utility of Data to the CERHR Evaluation...........................................................25

5.1.2 General Biological and Toxicological Data ....................................................................26

5.1.2.1 Utility of Data to the CERHR Evaluation...........................................................27

5.1.3 Developmental Toxicity..................................................................................................27

5.1.3.1 Utility of Data to the CERHR Evaluation...........................................................31

5.1.4 Reproductive Toxicity.....................................................................................................31

5.1.4.1 Utility of Data to the CERHR Evaluation...........................................................36

5.2 Integrated Evaluation ................................................................................................................36

5.3 Expert Panel Conclusions..........................................................................................................37

5.4 Critical Data Needs ...................................................................................................................38

6.0 REFERENCES............................................................................................................................................39

7.0 TABLES ......................................................................................................................................................44

Appendix II
Appendix II

II-i

PREFACE

The National Toxicology Program (NTP) and the National Institute of Environmental Health Sciences 

established the NTP Center for the Evaluation of Risks to Human Reproduction (CERHR) in June, 

1998. The purpose of the Center is to provide timely, unbiased, scientifically sound evaluations 

of human and experimental evidence for adverse effects on reproduction, including development, 

caused by agents to which humans may be exposed.

The following seven phthalate esters were selected for the initial evaluation by the Center: butyl 

benzyl phthalate, di(2-ethylhexyl) phthalate, di-isodecyl phthalate, di-isononyl phthalate, di-n-butyl 

phthalate, di-n-hexyl phthalate, and di-n-octyl phthalate. Phthalate esters are used as plasticizers in 

a wide range of polyvinyl chloride-based consumer products. These chemicals were selected for the 

initial evaluation by the CERHR based on their high production volume, extent of human exposures, 

use in children’s products, published evidence of reproductive or developmental toxicity, and public 

concern.

This evaluation is the result of three public Expert Panel meetings and 15 months of deliberations by 

a 16-member panel of experts made up of government and non-government scientists. This report 

has been reviewed by the CERHR Core Committee made up of representatives of NTP-participating 

agencies, by CERHR staff scientists, and by members of the Phthalates Expert Panel. This report is 

a product of the Expert Panel and is intended to (1) interpret the strength of scientific evidence that 

a given exposure or exposure circumstance may pose a hazard to reproduction and the health and 

welfare of children; (2) provide objective and scientifically thorough assessments of the scientific 

evidence that adverse reproductive/development health effects are associated with exposure to specific chemicals or classes of chemicals, including descriptions of any uncertainties that would diminish confidence in assessment of risks; and (3) identify knowledge gaps to help establish research and 

testing priorities.

The Expert Panel Reports on phthalates will be a central part of the subsequent NTP report that will 

also include public comments on the Panel Reports and any relevant information that has become 

available since completion of the Expert Panel Reports. The NTP report will be transmitted to the 

appropriate Federal and State Agencies, the public, and the scientific community.

The NTP-CERHR is headquartered at NIEHS, Research Triangle Park, NC and is staffed and 

administered by scientists and support personnel at NIEHS and at Sciences International, Inc., 

Alexandria, Virginia.

Reports can be obtained from the website <http://cerhr.niehs.nih.gov/> or from:

CERHR

Sciences International, Inc.

1800 Diagonal Road, Suite 500

Alexandria, VA 22314-2808

Telephone: 703-838-9440
II-ii

Appendix II

A Report of the CERHR Phthalates Expert Panel: 

Name Affiliation

Robert Kavlock, PhD (Chair) National Health and Environmental Effects Research 

Laboratory/USEPA, Research Triangle Park, NC

Kim Boekelheide, MD, PhD Brown University, Providence, RI

Robert Chapin, PhD NIEHS, Research Triangle Park, NC

Michael Cunningham, PhD NIEHS, Research Triangle Park, NC

Elaine Faustman, PhD University of Washington, Seattle, WA

Paul Foster, PhD Chemical Industry Institute of Toxicology, RTP, NC

Mari Golub, PhD California Environmental Protection Agency, Sacramento, CA

Rogene Henderson, PhD Lovelace Respiratory Research Institute, Albuquerque, NM

Irwin Hinberg, PhD Health Canada, Ottawa, Ontario, Canada

Ruth Little, ScD NIEHS, Research Triangle Park, NC

Jennifer Seed, PhD Office of Toxic Substances/USEPA, Washington, DC

Katherine Shea, MD, MPH Duke University, Durham, NC

Sonia Tabacova, MD, PhD Food and Drug Administration, Rockville, MD

Rochelle Tyl, PhD, DABT Research Triangle Institute, Research Triangle Park, NC

Paige Williams, PhD Harvard University, Boston, MA

Timothy Zacharewski, PhD Michigan State University, East Lansing, MI

With the Support of CERHR Staff:

NTP/NIEHS

Michael Shelby, PhD Director, CERHR

Christopher Portier, PhD Acting Associate Director, NTP

Gloria Jahnke, DVM Technical Consultant

Lynn Goldman, MD Technical Consultant

Sciences International, Inc. 

 

John Moore, DVM, DABT Principal Scientist 

Annette Iannucci, MS Toxicologist 

Ann Walker, MS, ELS Information Specialist and Technical Editor
Appendix II

II-1

1.0 CHEMISTRY, USAGE, AND EXPOSURE

1.1 Chemistry

Figure 1: Chemical Structure of Di-n-Butyl Phthalate

Di-n-butyl phthalate (DBP) (CAS RN 84-74-2) is produced through the reaction of n-butanol with 

phthalic anhydride (1).

Table 1: Physicochemical Properties of DBP

Property Value

Chemical Formula C16H22O4

Molecular Weight 278.35

Vapor Pressure 2.7 x 10-5 mmHg at 25 oC

Melting Point -35 o

C

Boiling Point 340 o

C

Specific Gravity 1.042

Solubility in Water Slight: 11.2 mg/L

Log Kow 4.45

(2)

1.2 Exposure and Usage

Overview

According to the American Chemistry Council (ACC, formerly CMA) (1), DBP is used mainly 

as a coalescing aid in latex adhesives. DBP is also used as a plasticizer in cellulose plastics and as 

a solvent for dyes. Although there was limited use of DBP in poly-vinyl chloride (PVC) plastics 

during the 1970’s and 1980’s, it is not currently used as a plasticizer in PVC. Release of DBP to 

the environment can occur during its production and also during the incorporation of the phthalate 

into plastics, adhesives, or dyes. Because DBP is not bound to the final product, it can be released 

during the use or disposal of the product. Phthalates that are released to the environment can be 

deposited on or taken up by crops intended for consumption by humans or livestock and can thus 

enter the food supply. 
II-2

Appendix II

General Population Exposure

Exposure of the general population to DBP has been estimated by at least four authoritative 

sources: the International Program on Chemical Safety (3), the UK Ministry of Agriculture, 

Fisheries, and Food (MAFF) (4, 5), Health Canada (6), and the US Agency of Toxic Substances and 

Disease Registry (7). Levels of DBP in exposure media, assumptions used in exposure calculations, 

and estimated exposure levels are detailed in Table 2 (3), Table 3 (7), and Table 4 (6). 

Table 2: IPCS Exposure Estimates for Adults

Ambient Air Indoor Air Drinking Water Food

DBP Concentration 

in Media

0.0045−0.0062 

μg/m3

0.420 μg/m3 <1.0 μg/L Various levels in a Canadian 

market basket survey. (See 

text)

Assumptions 22 m3 inhaled/day; 

64 kg bw; 4/24 

hours outdoors

22 m3 inhaled/day; 

64 kg bw; 20/24 

hours indoors

1.4 L/day intake; 

64 kg bw

Various intake rates for 

different food types; 64 kg 

bw

Estimated Doses

μg/kg bw/day

0.00026−0.00036 0.120 <0.02 7 

(3)

Table 3: ATSDR Exposure Estimates for Adults

Ambient Air Drinking Water Fish

DBP Concentration in Media 0.003−0.006 μg/m3 0.2 μg/L 78−200 μg/kg

Assumed Intake Rate 20 m3/day/70 kg adult 2 L/day/70 kg adult 6.5 g/day/70 kg adult

Assumed Absorption Fraction 0.5 0.9 0.9

Estimated Dose 

(μg/kg bw/day)

0.0005−0.0009 0.005 0.007−0.02

(7)

Table 4: Health Canada DBP Exposure Estimates

ESTIMATED INTAKE DBP (μg/kg bw/day)

Substrate/Medium 0.0−0.5 years old 0.5−4 years old 5−11 years old 12−19 years old 20−70 years old

Ambient Air* 0.00030 0.00040 0.00041 0.00038 0.00034

Indoor Air 0.68 0.91 0.1 0.87 0.78

Drinking Water 0.11 0.062 0.033 0.022 0.021

Food 1.6 4.1 3.2 1.4 1.1

Soil* 0.0070 0.0054 0.0018 0.00049 0.00040

Total Estimated 

Intake

2.4 5.0 4.3 2.3 1.9

* Value represents the upper range of the estimates. (6)
Appendix II

II-3

As noted in exposure estimates by the IPCS, Health Canada, and ATSDR, the largest source of 

DBP exposure to the general population is food. Sources of DBP in food include environmental 

uptake during crop cultivation or migration from processing equipment or packaging materials. 

IPCS (3) and Health Canada (6) conducted more comprehensive exposure estimates. Both exposure 

estimates were based on a 1986 Canadian market-basket survey of 98 different food types. Foods 

reported to contain DBP included butter (1.5 mg/kg), margarine (0.64 mg/kg), freshwater fish 

(0.5 mg/kg), cereal products (0−0.62 mg/kg), baked potatoes (0.63 mg/kg), bananas (0.12 mg/kg), 

coleslaw (0.11 mg/kg), gelatin (0.09 mg/kg), and white sugar (0.2 mg/kg). DBP exposure through 

food intake in adults was estimated at 7 μg/kg bw/day by IPCS (3) and at 1.9 μg/kg bw/day by 

Health Canada (6). DBP exposures in children were also estimated by Health Canada by applying 

appropriate assumptions such as intake rates of different food types per age group. Estimated DBP 

exposure levels from food ranged from 2.3 μg/kg bw/day in children aged 12−19 years to 5.0 μg/kg 

bw/day in children aged 6 months to 4 years. 

MAFF (4) estimated adult DBP exposure through dietary intake based on a 1993 survey of fatty 

foods in the UK. DBP was detected in carcass meat (0.09 mg/kg), poultry (0.2 mg/kg), eggs (0.1 

mg/kg), and milk (0.003 mg/kg). In calculating dietary food exposures, MAFF assumed that these 

types of food likely account for 85% of dietary phthalate intake. Food intake levels were obtained 

from the Dietary and Nutritional Study of British Adults, but the values were not reported by 

MAFF. Mean and high level DBP intakes were estimated at 13 μg DBP/person/day and 31 μg 

DBP/person/day, respectively. Specific details describing the calculations and assumptions used 

were not provided. Using the IPCS-assumed (3) adult body weight of 64 kg, the exposure values 

were converted to 0.20−0.48 μg/kg bw/day.

MAFF also addressed DBP exposure in infants resulting from the consumption of infant formula. 

A survey published in 1996 reported DBP levels of 0.08−0.4 mg/kg in infant formulas purchased 

in the UK, while a later survey reported DBP levels of <0.05–0.09 mg/kg (5, 8). It is speculated 

that the drop in DBP concentration occurred because infant formula manufacturers were urged 

to reduce phthalate levels after MAFF published the results of the 1996 survey. Exposure levels 

were estimated for infants based on the results from the 1998 survey using assumed body weights 

of 2.5–3.5 kg at birth and 7.5 kg at 6 months of age. Formula intake rates were determined from 

manufacturer instructions. Exposure levels for infants were estimated at 2.4 μg/kg bw/day at birth 

and 1.4 μg/kg bw/day at 6 months of age. Infants in the US are likely exposed to lower levels of 

DBP through formula than are infants in the UK. In a survey of infant formulas conducted in 1996, 

DBP levels in the US were approximately 10-fold lower than concentrations measured in the UK 

and ranged from <5 to 11 ppb (<0.005 to 0.011 mg/kg) (9). DBP has also been reported in baby 

food and breast milk samples collected from Germany and Japan; average values were within 

ranges reported by MAFF. DBP was measured in 7 German baby food samples (average 0.033 mg/

kg), 8 baby formulas (<0.2–0.9 mg/kg; average 0.042 mg/kg), and in the breast milk of 5 mothers 

from Germany (average 0.035 mg/kg) and 3 from Japan (0.02–0.08 mg/kg). The time period when 

these samples were collected was not specified (1). 

In their estimates of dietary exposure, ATSDR (7) only considered fish intake because at that time it 

was the only food source for which reliable data were available. The dietary estimate of 0.007−0.02 

μg/kg bw/day was based on DBP levels of 78–200 μg/kg that were reported for fish in studies 
II-4

Appendix II

published between 1973 and 1987. 

Levels of DBP in drinking water were estimated to be minimal. DBP exposure to adults through 

drinking water was estimated at 0.02 μg/kg bw/day by IPCS (3) and Health Canada (6) based 

upon a survey of drinking water supplies in Ontario, Canada. Health Canada also estimated DBP

exposures through drinking water intake in children and those values ranged from 0.022 μg/kg 

bw/day in children aged 12–19 years to 0.11 μg/kg bw/day in infants aged 0–6 months. Adult DBP 

exposure through drinking water was estimated by ATSDR (7) at 0.005 μg/kg bw/day. The value 

was based on a survey of drinking water in 10 unspecified cities prior to 1986. 

Mouthing of toys is another potential source of oral phthalate exposure in children. However, 

use of DBP in toys appears to be rare. In an analysis of 17 plastic toys, DBP was only detected in 

1 polyvinyl chloride doll’s head at 0.01% by weight (10). 

Although off-gassing from building materials has been reported as a potential source of DBP 

exposure through inhalation, exposure has been postulated to be minimal because of the low vapor 

pressure of DBP. The available data, though minimal, support this view. IPCS (3) estimated that 

adults are exposed to 0.120 μg/kg bw/day through inhalation of indoor air. The estimate was based 

on the mean air concentration of DBP measured within 125 homes in California in 1990. Health 

Canada also estimated indoor inhalation exposure to DBP based on a survey of DBP air levels in 9 

homes in Montreal (reported in 1985). Exposure to adults was estimated at 0.78 μg/kg bw/day and 

exposures in children ranged from 0.68 μg/kg bw/day in 0–6 month-old infants to 1.1 μg/kg bw/

day in 5–11 year-old children. Exposures to DBP through ambient air was also estimated by IPCS 

(3) and Health Canada (6); the values were roughly 2–3 orders of magnitude lower than the indoor 

air estimates. 

Dermal contact with products containing DBP is possible, but absorption through skin is most 

likely minimal. Studies in rats have demonstrated that absorption of DBP through skin is fairly slow 

(11). An in vitro study conducted with rat and human skin has demonstrated that human skin is 

much less permeable to DBP than is rat skin (12).

Caution is required to interpret exposure data for the general population. IPCS has emphasized 

that dietary intake can vary widely depending on the types of food eaten and the types of material 

in which the foods are packaged. In addition, the majority of data used to estimate exposure levels 

was collected 15–20 years ago and may not reflect current exposure levels. Lastly, exposures in 

children may be higher due to non-dietary intake through mouthing of DBP-containing objects. 

Medical Exposure

According to IPCS (3), a DBP level of 5 mg/g was measured in plastic tubing used for oral/nasal 

feeding. There are no other known uses of DBP in medical equipment. 

Occupational Exposure

Exposure in occupational settings can occur through skin contact and by inhalation of vapors 

and dust. Phthalates are manufactured within closed systems, but workers can be exposed during 
Appendix II

II-5

filtering or loading/unloading of tank cars (1). Higher exposures to phthalates can occur during the 

incorporation of the phthalate into the final product if the process is run at a higher temperature. In 

a limited number of surveys, DBP levels in US plants have ranged from concentrations below the 

detection limit (0.01–0.02 mg/m3) to 0.08 mg/m3 (3). OSHA established a permissible exposure 

limit of 5 mg/m3 for DBP. Following a review of six studies, the ACC has estimated exposure to 

DBP in the workplace based upon an assumed level of 1 mg/m3 during the production of phthalates 

(1). Exposure levels during the incorporation of DBP into plastics are not known. An exposure 

level was estimated by using assumptions of a 10 m3/day inhalation rate and a 70 kg body weight. 

The resulting exposure estimate was 143 μg/kg bw/workday for workers employed in phthalate 

manufacturing. The maximum exposure, by regulation, would be five-fold greater. As stated in the 

General Exposure section, absorption of DBP through skin is expected to be minimal. 

Conclusion

Exposure estimates varied between authoritative bodies. However, in all cases it was evident 

that food was the primary source of exposure to DBP. ATSDR only considered fish intake, and 

their exposure estimate therefore provides no information on total dietary exposure. The dietary 

exposure estimate by MAFF is approximately one order of magnitude lower than estimates by 

IPCS and Health Canada. The basis for discrepancies in dietary exposure estimates is difficult to 

determine for several reasons, including: use of different food types in calculations (e.g., fatty 

foods vs a variety of foods); use of different assumptions in calculations; varying DBP levels in 

foods from different countries; and changing DBP levels in food over time. Table 5 lists the dietary 

DBP estimates calculated by the different agencies for infants and adults. 

Table 5: Comparison of DBP Dietary Estimates

Agency Exposure in Infants (0–6 months)

(μg/kg bw/day)

Exposure in Adults

(μg/kg bw/day)

IPCS (3) N/A 7

MAFF (4, 5, 8) 1.4–2.4 0.2–0.48

ATSDR (7) N/A 0.007–0.02

Health Canada (6) 1.6 1.1

The summary for Section 1 is located in Section 5.1.1.
II-6

Appendix II

2.0 GENERAL TOXICOLOGICAL AND BIOLOGICAL PARAMETERS

2.1 General Toxicity

2.1.1 Human Data

There were no human data located for Expert Panel review.

2.1.2 Experimental Animal Data

Multiple evaluations are available for assessing the effects of oral exposure to DBP. A few 

inhalation and dermal evaluations have also been conducted; these studies are primarily in rats with 

a few assessments in mice, rabbits, hamsters, and guinea pigs.

Acute studies

The oral LD50 for DBP appears to be between 8,000 and 20,000 mg/kg bw in rats (3) and the 90-day 

dermal LD50 is 4,200 mg/kg bw in rabbits. Slight irritation was observed in rabbit dermal occlusion 

studies at 520 mg/kg bw.

Repeat-dose studies

In a 3-month sub-chronic study, 6-week old Wistar rats, 10 of each sex per dose, were fed a 

diet containing 0, 400, 2,000 or 10,000 ppm DBP (13) (Table 7-1). In addition to developing a 

toxicological profile of DBP, a stated purpose of the study was to evaluate possible neurological or 

testicular toxicity. A battery of standard hematological and clinical chemistry parameters (including 

thyroid function) was evaluated at points approximately halfway through and at the end of the 

study. Cyanide insensitive palmityl-CoA oxidation (PCoA) was also determined as a measure of 

peroxisome proliferation. Urinalyses were performed at the midpoint and at the end of the study. 

Neurological function, using the EPA functional observation battery, was assessed prior to DBP 

administration, and on days 34, 59, and 90 of the study. 

Dietary consumption was not a factor in the study; nominal daily doses were calculated to be 27

(M) and 33 (F) mg/kg bw/day, 142 (M) and 162 (F) mg/kg bw/day, and 688 (M) and 816 (F) mg/kg 

bw/day for the three dose groups. Effects were observed only in the high-dose group, 688 (M), 

and 816 (F) mg/kg bw/day. Statistically significant increases in liver and kidney to body weight 

ratios were observed in the absence of body weight changes in females. Histologically, a decrease 

in lipid deposition was noted in hepatocytes; this effect was possibly due to peroxisome-related 

enzyme increases in the liver. An increase in PCoA activity was confirmed. Serum triglycerides 

and triiodothyronine were both decreased. RBC, hemoglobin, and hematocrit were transiently 

decreased in males. No histological effects on testes appropriately preserved in Bouin’s fixative 

were observed.

Neurological function was assessed at three time points during the study and no effects were 

observed. A LOAEL was observed at 688 (M) and 816 (F) mg/kg bw/day based on multiple 

impacts and a NOAEL was determined at 142 (M) and 162 (F) mg/kg bw/day. 
Appendix II

II-7

Marsman (14) reported two 13-week, sub-chronic NTP studies using male and female F344 rats. 

One of the studies was of traditional design; 5–6 week-old rats were exposed to either control or 

one of four test diets. In the second study, rats placed in a standard sub-chronic design were born 

and reared by mothers exposed to 10,000 ppm DBP during pregnancy and nursing; at weaning, they 

were further exposed to a 10,000 ppm DBP diet until 8 weeks of age.

In the standard study, 10 F344 rats/sex were exposed to DBP in their diet for 13 weeks starting 

at 5–6 weeks of age (Table 7-2). The dietary levels were 0, 2,500, 5,000, 10,000, 20,000, and 

40,000 ppm (M: 0, 176, 359, 720, 1,540, and 2,964 mg/kg bw/day; F: 0, 177, 356, 712, 1,413, 

and 2,943 mg/kg bw/day). At the end of the study the rats were killed and necropsied with 

extensive tissue examination (testes preserved in 10% neutral buffered formalin), hematology and 

clinical chemistry, sperm morphology, and vaginal cytology parameters were evaluated. Zinc and 

testosterone levels were measured in sera and testes of all males. An increase in serum albumin was 

observed in exposed males at 176 mg/kg bw/day, the lowest dose tested. No other effects were seen 

in either sex at this dose. Adverse effects in males seen at the next highest dose (359 mg/kg bw/day) 

were evidenced by a decrease in hemoglobin and erythrocyte counts. Severity of the hematological 

effects, seen only in males, progressed in a dose-response manner at all other doses. Platelets and 

serum albumin were increased, as were liver and kidney organ to body weight ratios. An increase in 

PCoA activity was seen in both sexes, and an increase in bile acid was seen in females. Decreases 

in body weight occurred in males at the 720 mg/kg bw/day dose, the third highest out of 5 treatment 

levels. Males exposed to 359 mg/kg bw/day and males and females exposed to 712−720 mg/kg 

bw/day had increased liver and kidney organ to body weight ratios. Hepatic lesions in males 

and females and testicular lesions were first noted at 712-720 mg/kg bw/day. Testicular lesions 

consisted of focal seminiferous tubule atrophy in 4/10 males. The chemistry changes noted at the 

next lower dose (356−359 mg/kg bw/day) continued at this dose (712−720 mg/kg bw/day) with the 

addition of increases in alkaline phosphatase activity. The histologic hepatic lesions persisted and 

testicular lesions increased in severity at the higher doses with all males of that dose group affected. 

Hypospermia of the epididymis was observed at the two highest doses. Decreases in testicular 

organ weight ratios, testicular zinc, and testosterone were not observed until the 1,540 (M) mg/kg 

bw/day exposure level. Peroxisome proliferation was noted histologically at the highest dose tested 

(2,964 [M] and 2,963 [F] mg/kg bw/day). Good dose-response data was available for almost all 

parameters in this study. A NOAEL of 176 mg/kg bw/day was identified by the Expert Panel. 

In the second NTP sub-chronic study, F344/N rats were born and reared by mothers exposed to 

10,000 ppm DBP in diet throughout prenatal development and lactation; the weaned rats were 

then fed a 10,000 ppm diet until 8 weeks of age (14) (Table 7-3). At that time, the male rats, 10 

per sex per group, were placed on 1 of 5 diets for an additional 13 weeks that contained 0, 2,500, 

5,000, 10,000, 20,000, or 40,000 ppm DBP (M: 0, 138, 279, 571, 1,262, or 2,495 mg/kg bw/day; 

F: 0, 147, 294, 593, 1,182, or 2,445 mg/kg bw/day) (14). The sub-chronic exposure doses and the 

protocols for histopathology, hematology, and chemistry were the same as the NTP sub-chronic 

study discussed above. The authors concluded that developmental exposure to DBP resulted in 

neither increased sensitivity nor resistance to DBP exposure during adulthood (compare results 

in Tables 2 and 3). The Expert Panel notes that there were significant increases in organ to body 

weight ratios for kidney and liver in females and in testes at the lowest exposure group, 138 (M) 

and 147 (F) mg/kg bw/day. Such findings were not observed at this dose level in the other sub-
II-8

Appendix II

chronic study.

The NTP also conducted a sub-chronic study in 6-week-old B6C3F1 mice where 10 mice per sex 

were fed DBP in the diet for 13 weeks at levels of 0, 1250, 2,500, 5,000, 10,000, and 20,000 ppm 

(M: 0, 163, 353, 812, 1,601, and 3,689 mg/kg bw/day; F: 0, 238, 486, 971, 2,137, and 4,278

mg/kg bw/day (14) (Table 7-4). Experimental design in this study was similar to the 13-week subchronic study in rats. There were no clinical signs related to exposure and all mice survived until 

the end of the study. Decreases in body weight gain were observed in both sexes fed levels of 812 

mg/kg bw/day or higher. Increases in absolute and relative kidney weight were seen in all treated 

female groups, but absolute kidney weight was decreased in high-dose males. There was no report 

of histological change in the kidney nor did weights increase with increasing dose. The liver was 

the only organ identified as a site of DBP toxicity by the study authors. Relative liver weights were 

increased at doses of 812 mg/kg bw/day and higher. Cytoplasmic alterations consisting of fine 

eosinophilic granules, more intensely-staining cytoplasm, and increased lipofuschin were observed 

at the 2 highest doses in males (1,601 and 3,869 mg/kg bw/day) and at the highest dose in females 

(4,278 mg/kg bw/day). A reduced hematocrit level was observed in high-dose females. Based 

on decreased body weight gain, the NOAEL is 353 mg/kg bw/day in males. A LOAEL based on 

increased kidney weight in females is 238 mg/kg bw/day, the lowest dose tested according to the 

Expert Panel.

In a series of three identical experiments, Walseth and Nilsen (15) examined lung and liver effects 

in groups of five male Sprague-Dawley rats. The rats were exposed for 6 hours/day for 5 days 

to DBP vapors at 0, 0.5, 2.5, or 7.0 ppm (0, 5.7, 28.4, and 79.5 mg/m3 as calculated by authors). 

There were no effects on lung or liver weights. In the lung, there were dose-related decreases 

in microsomal cytochrome P-450 and cytochrome c-reductase levels in the two highest dose 

groups. There were no dose-related changes in liver cytochrome levels. A significant decrease 

in serum levels of alanine aminotranferase (ALAT) and significant increases in serum aspartate 

aminotranferase and albumin levels were observed, but the authors indicated that there was 

no evidence of liver cell damage. The authors concluded that the lung is the main target organ 

following inhalation exposure to DBP. 

2.2 Toxicokinetics

Phthalate Moiety

Absorption

Humans: Dermal. In an in vitro study, human skin absorption rate was reported as 0.07 μg/cm2/

hour (12) which was considered “slow.” 

Humans: Oral. DBP was detected in blood from humans following ingestion of foodstuffs 

containing DBP (3). Background levels of DBP in human blood were much higher following 

exposure. Unfortunately, the authors measured only the parent compound so there is no estimate of 
Appendix II

II-9

total DBP equivalents absorbed in this study. Similarly, levels of DBP in human adipose tissue were 

studied (16); again total DBP equivalents were not calculated. 

Rodents: Dermal. Dermal absorption of DBP was studied in Fischer 344 rats by applying 30−40 

mg/kg radiolabeled DBP to the skin (administration site occluded) and measuring the radioactivity 

in urine (11). Approximately 10−12% of the dose was excreted in urine per day with approximately 

60% of the dose excreted within 1 week. Thirty-three percent of the dose was present at the 

application site 1 week following treatment.

Rodents: Oral. The extent of intestinal absorption of phthalate esters has been estimated 

by monitoring urinary excretion of the parent compounds or their metabolites after orally 

administering a known amount of the radiolabeled compound. Greater than 90% of radioactivity 

following an oral dose of DBP in rats is recovered in the urine within 2 days, indicating nearly 

complete intestinal absorption of this compound over a range of administered doses (17). This is 

consistent with the general observation that dialkyl phthalate esters are well absorbed following oral 

dosing. It is generally accepted that orally-ingested phthalate diesters are quantitatively hydrolyzed 

by gut lipases and absorbed almost entirely as the corresponding monoester. 

Biotransformation

Humans. In a study comparing the relative rates of monohydrolysis of DBP by rat, baboon, and 

human gut preparations, Lake et al. (18) demonstrated that these species possess similar intrinsic 

lipase activity. Rates observed in human intestinal preparations were similar enough to the other 

species to expect that human intestinal metabolism of DBP would result in absorption of the 

monoester similarly to rats. The activity of pancreatic lipase was not assessed, so the quantitative 

relationships of this study to in vivo exposure cannot be accurately determined (18).

Rodents. Dialkyl phthalates including DBP were found to be metabolized to the monoesters by 

enzymes present in many tissues. It is generally accepted that orally-ingested phthalate diesters are 

quantitatively hydrolyzed by lipases in the wall of the small intestine and pancreatic lipases and not 

by gut flora. Absorption occurs almost entirely as the corresponding monoester (19).

Metabolites of DBP include monobutylphthalate, monobutylphthalate glucuronide, o-phthalic acid 

and oxidized monobutylphthalate glucuronide metabolites (17).

Distribution

Humans. No human data were located for Expert Panel review.

Rodents. DBP is rapidly cleared following oral or intravenous (IV) administration. There is little 

or no bioaccumulation observed. Radioactivity associated with DBP administration can be found 

in the GI tract and excretory organs of the liver and kidney, and in fat. Liver, kidney, and the GI 

tract probably accumulate the phthalate esters as a mechanism of excretion and not as depots (20). 

One week following dermal treatment of Fischer 344 rats with 30−40 mg/kg radiolabeled DBP, no 

tissues examined contained more than 2% of the administered dose (11). 
II-10

Appendix II

Pregnant Rodents. Saillenfait et al. (21) studied metabolism and placental transfer of 14C-DBP, 

administered by gavage on gestation day (gd) 14 at 500 or 1,500 mg/kg to Sprague-Dawley rats. 

Radioactivity peaked followed by a rapid decline in all tissues within 1–2 hours of administration. 

Maternal plasma had the highest peak concentration; all tissue levels were less than 7% of peak 

concentrations by 24 hours. Fifty-five percent and 29% of a 500 mg/kg 14C dose were detected in 

urine and feces respectively in 24-hour samples; there was a slight increase to about 60% in urine 

at 48 hours, whereas fecal values did not change. Radioactivity in placenta, embryo, and amniotic 

fluid were 0.3, 0.15, and 0.2% of the administered dose, respectively. Concentrations in placenta 

and embryo did not exceed 30 and 21% of maternal plasma levels. The 1,500 mg/kg dose indicated 

slower absorption from the gastrointestinal tract; total fecal radioactivity was not affected, although 

there was lower excretion in urine over 48 hours. In maternal plasma, placental, and embryonic 

tissues, monobutyl phthalate (MBuP) and its glucuronide represented most of the DBP-derived 

activity. MBuP Levels ranged from 50 to 95%, dependent upon the time after administration when 

samples were taken. In contrast, unchanged DBP accounted for less than 1%. The authors speculate 

that the lower levels of MBuP glucuronide in embryonic tissues compared to those in maternal 

plasma could reflect limited placental transfer or limited ability to conjugate this substrate. Levels 

of radioactivity in placenta and embryos associated with DBP administration were approximately 

65% of the levels found in maternal serum and there was no bioaccumulation of radioactivity 

observed in the embryonic tissues. DBP, MBuP, and MBuP-glucuronide were present in embryonic 

tissues at levels lower than were found in maternal plasma. MBuP accounted for most of the 

radioactivity recovered in maternal plasma, placenta, and embryos, which is consistent with the 

hypothesis that MBuP is the ultimate teratogenic species in vivo. 

Distribution following IV exposure produces a different distribution pattern than that observed 

following oral administration. Since DBP is not in direct contact with gut esterases, metabolism to 

the monoester is slowed. This produces more DBP-associated radioactivity to distribute to lungs 

and blood in addition to liver and kidney. Radioactivity was detectable in adipose tissue 7 days 

after IV exposure (22). The difference between the oral and IV distribution probably reflects a 

higher concentration of parent DBP reaching adipose tissue following IV exposure, which would be 

expected to distribute to lipophilic tissues such as adipose tissue. 

Excretion

Humans. No human data were located for Expert Panel review.

Rodents. The primary route of MBuP, the major DBP metabolite, elimination in rodents and 

humans is urinary excretion. The monobutylphthalate glucuronide appears to be the primary 

metabolite identified in rat urine (23). MBuP is excreted into the bile (about 45%), but only about 

5% is eliminated in the feces, indicating that efficient enterohepatic recirculation occurs (17). 

Biliary metabolites of DBP include monobutylphthalate, monobutylphthalate glucuronide, and 

oxidized monobutylphthalate glucuronide metabolites (17). Following dermal exposure of rats to 

DBP, urine was the primary route of excretion with the excretion rate remaining nearly constant at 

10–12% of the dose excreted per day (11). 
Appendix II

II-11

Mice are known to excrete higher amounts of glucuronidated phthalate ester metabolites than rats 

and primates excrete higher levels of glucuronidated phthalate ester metabolites than mice (24). 

There appears to be little retention of DBP or MBuP in tissues of rats treated with DBP for 12 

weeks (20).

Models

A physiologically-based pharmacokinetic (PBPK) model of the tissue distribution of DBP and 

its monoester metabolite, MBuP, in rats administered DBP by various routes has been developed 

by Keys et al. (25). The model is based on an earlier model developed by the same group for 

DEHP and its metabolite, MEHP (26). It includes a combined perfusion-limited and pH trapping 

mechanism for uptake of MBuP into tissues, and it provides a valuable tool for extrapolations of 

tissue doses among various routes and rates of exposure. With modification, the model can be used 

to extrapolate doses to target tissues among various species and ages and between genders and 

gravid vs non-gravid females. The model allows estimation of the internal dose to specific target 

tissues for the evaluation of risk, rather than using total exposure or total internal dose as a risk 

estimate.

Side Chain-associated Toxicokinetics (butanol)

Butanol, a metabolite of DBP, is a primary alcohol that is easily oxidized to butyric acid (n-butanoic 

acid) by alcohol dehydrogenase and aldehyde dehydrogenase. Further metabolism by oxidation 

pathways converts butyric acid into acetyl-CoA conjugates in intermediary metabolism pathways 

with no toxicological importance (27).

2.3 Genetic Toxicity

DBP has tested negative or marginally positive in gene mutation and chromosomal aberration 

studies. The ASTDR (7) concluded that DBP may be weakly mutagenic. The significance of these 

findings is not known because in vivo genotoxicity studies have not been conducted. The Woodward 

et al. (28) review concluded that the evidence indicates that DBP is not directly genotoxic, but 

noted it does cause increases in sister chromatid exchanges and small increases in the incidence of 

gaps and breaks. However, the effect does not appear to be dose-related (29). IPCS (3) reviewed a 

number of mutagenic and related endpoints for DBP and concluded that the weight of the evidence 

indicated that DBP is not genotoxic. DBP was positive in the L5178Y mouse lymphoma assay in 

the presence, but not in the absence, of an Aroclor-induced rat liver activation system (S9) (30). 

The authors conclude that the positive activity was likely the result of in vitro metabolism of the 

DBP to an aldehyde, and therefore, that the results may not represent any real potential for in vivo 

genotoxicity. DBP is not mutagenic in the Salmonella/mammalian microsome mutagenicity assay 

(31), and was negative in the Balb/3T3 cell transformation assay (30).

The summary for Section 2, including general toxicity, toxicokinetics, and genetic toxicity, is 

located in Section 5.1.2.
II-12

Appendix II

3.0 DEVELOPMENTAL TOXICITY DATA

3.1 Human Data

There were no human data located for Expert Panel review.

3.2 Experimental Animal Toxicity

A number of studies have evaluated DBP for both prenatal and postnatal developmental toxicity; 

the vast majority of studies have been performed in the rat using the oral route of exposure. In most 

cases, the doses were high (> 0.5% in diet; > 500 mg/kg bw/day), and the number of animals per 

dose group was small (10–15).

3.2.1 Prenatal Development

DBP

Results from a set of investigations in mice have been reported by Shiota et al. (32) and Shiota and 

Nishimura (33) (Table 5). They evaluated the effects of oral exposure to DBP in concentrations 

of 0, 0.05, 0.1, 0.2, 0.4, and 1.0% in the diet. On the day a cervical plug was observed (gd 0), 

female ICR-JCL mice commenced eating the DBP diet until they were killed on gd 18. Using food 

consumption data, the authors calculated mean daily intake of DBP to be 0, 80, 180, 350, 660, 

and 2,100 mg/kg bw/day. Six-to-nine litters were examined per dose group, except that 15 litters 

were examined from the highest dose group. Food intake levels were not affected in pregnant 

dams. Maternal weight gain was significantly reduced at the high dose (2,100 mg/kg bw/day), but 

the effect may have been secondary to increased fetal loss. Resorptions (prenatal mortality) were 

significantly increased (98.4%) in the high-dose group. At this dose, malformations in 2/3 surviving 

fetuses (increase not statistically significant) were limited to neural tube defects (exencephaly 

and spina bifida, to which murine species are predisposed). Delayed ossification was observed at 

all dose levels as indicated by a reduction in the number of ossified coccygia in treated fetuses 

(n=9.4, 5.1, 4.5, 6.0, and 2.6 in the control to 660 mg/kg bw/day groups). Reduced fetal body 

weight was observed at the two highest doses. Because ossification was delayed at all dose levels, 

a developmental NOAEL could not be identified for this study and, therefore, a LOAEL of 80 mg/

kg bw/day was selected by the Expert Panel. However, the authors stated that “the maximum nonembryotoxic dose” was 370 mg/kg bw/day. The maternal NOAEL and LOAEL were identified as 

660 and 2,100 mg/kg bw/day, respectively. 

Ema et al. (34-36) used Wistar rats to evaluate the developmental toxicity of DBP by exposure 

through gavage and feed. In all studies, dams were sacrificed on gd 20–21 and examined for 

implantation sites. Fetuses were weighed and examined for external, skeletal, and visceral 

malformations. In one Ema (34) study, 12 dams/group were gavaged with 0, 500, 630, 750, or 1,000 

mg/kg bw/day (0, 1.80, 2.27, 2.70, or 3.60 mmol/kg bw/day) on gd 7–15 (Table 7-6). Gestational 

weight gain was reduced in dams of the 630 mg/kg bw/day group and adjusted weight gain (dam 

weight not including gravid uterus) was reduced in dams exposed to 750 mg/kg bw/day and higher. 
Appendix II

II-13

Complete resorptions occurred in 2/12, 10/12, and 9/9 litters of the 630, 750, and 1,000 mg/kg 

bw/day dose groups, respectively, thus resulting in decreased live fetuses/litter. Fetal weight was 

reduced in groups exposed to 630 mg/kg bw/day and higher. External malformations, consisting 

entirely of cleft palate, were increased in the 750 mg/kg bw/day group. Maternal and developmental 

NOAELs and LOAELs of 500 and 630 mg/kg bw/day, respectively, were identified.

Another study conducted by Ema et al. (36) is of particular interest because it examines additional 

endpoints including anogenital distance and testicular descent (Table 7-7). In this study, 11 dams/

group were fed diets containing 0, 0.5, 1.0, or 2.0% DBP on gd 11–21. Authors estimated daily 

intake rates of 0, 331, 555, and 661 mg/kg bw/day for the control to high-dose groups, respectively. 

Maternal gestational and corrected weight gain were reduced in dams exposed to 555 mg/kg bw/

day and higher and were accompanied by a reduction in food intake. Fetal weight was reduced and 

the incidence of external malformations (cleft palate) and skeletal malformations (fused sternebrae) 

were increased in the 661 mg/kg bw/day dose group. Reduced anogenital distance and increased 

incidence of undescended testes were observed in male fetuses exposed to 555 and 661 mg/kg 

bw/day. The maternal and developmental NOAEL and LOAEL of 331 and 555 mg/kg bw/day, 

respectively, were identified for this study. 

The two remaining studies by Ema et al. (35, 37) focused on the time- and dose-dependency of 

DBP developmental toxicity. In the studies, groups of 10–13 pregnant rats were gavaged with 0, 

750, 1,000, 1,250, or 1,500 mg/kg bw/day on gd 7–9, 10–12, or 13–15. Resorptions were increased 

in all dose groups at all time points. All dams treated with 1,500 mg/kg bw/day experienced 

complete litter resorptions. However, the types and frequencies of malformations varied according 

to the exposure time course. Treatment on gd 10–12 did not result in an increased malformation 

rate. Treatment with doses of 750 mg/kg bw/day and higher on gd 7–9 resulted in increased skeletal 

malformations (fusion or absence of vertebral arches and ribs). Administration of 750 mg/kg bw/

day and higher on gd 13–15 resulted in the greatest incidence of teratogenicity, including increased 

external malformations (cleft palate) and skeletal malformations (fusion of sternebrae). 

Saillenfait et al. (21) exposed Sprague-Dawley rats (27 per group) to a single administration 

of DBP by gavage on gd 14 at 0, 500, 1,000, 1,500, or 2,000 mg/kg body weight. Increased 

resorptions at 1,500 and 2,000 mg/kg and reduced fetal body weights at 2,000 mg/kg were 

observed. Skeletal variations were also increased at these doses. Key aspects of the paper were 

studies on metabolism and placental transfer of 14C-DBP, administered by gavage on gd 14 at 500 

or 1,500 mg/kg. The toxicokinetic data are presented in Section 2.2. The authors concluded that 

their data support the view that MBuP may be the proximate toxicant.

Developmental effects were also noted in reproductive toxicity studies, which are discussed in 

detail under Section 4. In a continuous-breeding study, two generations of Sprague Dawley rats 

were exposed to 0, 80, 385, or 794 mg/kg bw/day through diet during a 98-day mating period (38). 

Maternal effects were only observed in the high-dose group and included a decrease in body weight 

for both generations and increased liver and kidney weights in F0 dams. Developmental effects 

included a reduction in litter size in all dose groups and in live pup weight in the two highest doses 

of F1 rats. F2 pups in all treatment groups experienced a reduction in body weight. A developmental 

LOAEL of 80 mg/kg bw/day and a maternal NOAEL of 385 mg/kg bw/day were identified.
II-14

Appendix II

A similar continuous-breeding study was conducted in one generation of CD-1 mice treated with 

0, 53, 525, and 1,750 mg/kg bw/day in diet (39, 40). Fetal effects that were observed only at the 

highest dose included reductions in litter size, live pups/litter, and pup weight. The developmental 

NOAEL was identified as 525 mg/kg bw/day, but a maternal NOAEL could not be identified 

because necropsies were only conducted in the high-dose group. In a multigeneration reproductive 

study, Long Evans Hooded rats were treated with 0, 250, or 500 mg/kg bw/day DBP by gavage 

from the time they were weanlings through the time that they nursed their own litters (41). Maternal 

toxicity was not reported. Developmental effects included malformations in reproductive organs, 

kidneys, and eyes in F1 rats and reductions in F2 litter size in all dose groups. The developmental 

LOAEL was identified as 250 mg/kg bw/day. 

MBuP

The prenatal developmental effects of administering MBuP by gavage in the Wistar rat were 

reported (42, 43). The Expert Panel noted that some of the doses used in these studies were 

equimolar equivalents to doses used in earlier studies with DBP (described above). Ema et al. (42) 

studied doses of 0, 250, 500, and 625 mg/kg bw/day (0, 1.13, 2.25, or 2.80 mmol/kg bw/day) on gd 

7–15. They observed maternal toxicity at the two highest doses expressed as reduced weight gain 

and feed consumption. Also, at these doses there were significant increases in post-implantation 

loss/litter and decreases in live fetuses/litter and fetal body weight/litter. Fetal malformations were 

increased, with cleft palate, deformed vertebral column, and dilated renal pelves the predominant 

findings. A maternal and developmental NOAEL and LOAEL of 250 and 500 mg MBuP/kg bw/day, 

respectively, were identified for this study. 

Ema et al. (43) then followed up with evaluation of stage specificity studies by administering MBuP 

at doses of 0, 500, 625, or 750 mg/kg bw/day on gd 7–9, 10–12, or 13–15. Embryolethality was 

increased at all doses for all dosing intervals. No teratogenicity was observed from the gd 10–12 

dosing interval. Increased incidences of fetal external malformations were present at the 500 and 

750 mg/kg bw/day doses on gd 7–9 and 13–15. Increased skeletal malformations were observed 

at 500, 625, and 750 mg/kg bw/day on gd 7–9 and at 625 and 750 mg/kg bw/day on gd 13–15 

(deformed cervical vertebrae were predominant on gd 7–9). Cleft palate and fused sternebrae were 

observed on gd 13–15. These results are consistent with the findings for DBP and imply that MBuP 

(and/or subsequent metabolites) may account for the developmental toxicity (embryolethality and 

malformations) for DBP.

3.2.2 Postnatal Development

DBP

Marsman et al. (14) exposed F344/N rats and B6C3F1 mice to high dietary concentrations of DBP 

during gestation and lactation. Both species were exposed to 0, 1,250, 2,500, 5,000, 7,500, 10,000, 

and 20,000 ppm. Dosages in mg/kg bw/day were estimated by using average values from 2 NTP 

studies that included a food intake rate of 14.8 g/day and a body weight of 203.71 g for rats and a 

food intake rate of 7.18 g/day and body weight of 39.63 g for mice (44-46). The dosages are listed 

in Tables 8 and 9. After weaning on pnd 21, up to 10 F1 pups/group were fed a diet with a DBP 
Appendix II

II-15

concentration identical to that fed to their dams fed for an additional 4 weeks. Author-calculated 

doses for pups were: 143, 284, 579, 879, and 1,165 mg/kg bw/day for male rats; 133, 275, 500, 

836, and 1,104 mg/kg bw/day for female rats; 199, 437, 750, 1,286, and 3,804 mg/kg bw/day for 

male mice; and 170, 399, 714, and 1,060 mg/kg bw/day for female mice. Complete necropsies were 

performed on one rat and one mouse pup of each sex per litter at weaning and on all pups at the 

end of the 4-week post-weaning dietary exposure. Organ weights were obtained on major organs, 

including testis. Histopathological examination was performed on a broad array of tissues from 

all animals in the control and highest exposure group. In addition, the epididymis of rats from the 

2,500, 5,000 and 7,500 ppm groups were studied. 

For the rats (Table 7-8), gestational index was reduced (fewer live litters) at 5,000 and 20,000 ppm, 

and gestational length was reduced at 5,000 ppm. Litter size and postnatal survival were reduced 

at 10,000 and 20,000 ppm. All F1 pups died by pnd 1 in the 20,000 ppm group. Male pup body 

weights were reduced during lactation in dose groups receiving 7,500 ppm and higher. In the postweaning period, relative liver and kidney weights were increased in female offspring exposed to 

≥2,500 and ≥5,000 ppm (275 and 500 mg/kg bw/day), respectively. Increased liver and kidney to 

body weight ratios were observed in males of all dose groups. Reduced relative testis weights were 

observed at the highest dose. Mild-to-marked hypospermia was seen in all males at the 879 and 

1,165 mg/kg bw/day doses and in 4/10 males of the 579 mg/kg bw/day dose group. There were no 

histopathological lesions observed in liver or kidney. Acquisition of vaginal patency and preputial 

separation were not assessed. Based on increased liver and kidney to body weight ratios in all 

treated males, no NOAEL was identified.

For B6C3F1 mice (Table 7-9), length of gestation was increased at 2,500 ppm and higher with 75 

and 95% of litters lost at 10,000 and 20,000 ppm. Decreases were observed in litter size and pup 

body weights at 2,500, 7,500, and 10,000 ppm. In the F1 post-weanling phase, males exhibited 

increased relative liver weights (one surviving male pup at 10,000 ppm exhibited hepatic lesions), 

and females exhibited increased relative kidney weights at 1,250 ppm (170–199 mg/kg bw/day) 

and higher. Except for liver lesions in the male at 10,000 ppm, no histpathological changes were 

observed, including in the testis. No NOAEL was identified.

Taking note of the Wine et al. (38) continuous-breeding study results (see Section 4), Mylchreest et 

al. (47) followed up the study using comparable dose levels (Table 7-10). However, three important 

changes in experimental design were introduced: 1) shortening the exposure period to include 

only gestation and lactation; 2) using gavage (with corn oil) to control exposure more closely; and 

3) including more sensitive endpoints of reproductive development, such as markers of sexual 

maturation. Thus, pregnant CD rats (10 per group) were administered DBP by gavage at 0, 250, 

500, or 750 mg/kg bw/day from gd 3 until pnd 20. At birth, pups were counted, sexed, weighed, and 

examined for signs of toxicity. Sexual maturity was assessed by observing age of vaginal opening 

and preputial separation in females and males, respectively. Estrous cycles were assessed in females 

for 2 weeks. The F1 rats were sacrificed at 100–105 days of age. Necropsies were conducted on all 

males and up to three females per litter. A histological examination of sex organs was conducted 

on all rats with lesions and up to two unaffected rats per litter. Testes were preserved in Bouin’s 

fixative. 
II-16

Appendix II

Maternal body weight gain was comparable to controls throughout the dosing period. At 750 mg/kg 

bw/day, the number of live pups per litter at birth was decreased and maternal effects on pregnancy 

and postimplantation loss are likely to have occurred. Anogenital distance was decreased at birth 

in the male offspring at 500 and 750 mg/kg bw/day. The epididymis was absent or underdeveloped 

in 0, 9, 50, and 71% of adult offspring (100 days old) at 0, 250, 500, and 750 mg/kg bw/day, 

respectively, and was associated with testicular atrophy and widespread testicular germ cell loss. 

Hypospadias occurred in 0, 3, 21, and 43% of males, and ectopic or absent testes in 0, 3, 6, and 

29% of males at 0, 250, 500, and 750 mg/kg bw/day, respectively. Absence of prostate gland and 

seminal vesicles as well as small testes and seminal vesicles were noted at low incidence in the 500 

and 750 mg/kg bw/day dose groups. Dilated renal pelves, frequently involving the right kidney, 

were observed in all DBP dose groups. Vaginal opening and estrous cyclicity were not affected 

in the female offspring, although low incidences of reproductive tract malformations, mainly 

involving development of the uterus, were observed in 2 rats and 1 rat at the 500 and 750 mg/kg 

bw/day doses, respectively. 

In the Mylchreest et al. 1998 study (47), all exposed groups showed adverse effects on male 

reproductive tract structure and indices of puberty. Based on this, the LOAEL in this study is 250 

mg/kg bw/day/day. Based on the relationship between testis weight/histopathology and sperm 

production, the relationships between sperm numbers and fertility (48), and the number of major 

malformations of the reproductive tract, it is expected that at least the high- and mid-dose animals 

would be sub-fertile. The Panel’s confidence in the quality of the study is high. 

In a subsequent study, Mylchreest (49) reduced DBP exposure to just late gestation (gd 12–21) 

and compared the effects of DBP to the pharmacological androgen receptor antagonist, flutamide 

(Table 7-11). Pregnant CD rats received DBP at 0, 100, 250, or 500 mg/kg bw/day by gavage 

with corn oil (n =10) or flutamide at 100 mg/kg bw/day (n =5) on gd 12−21. Males were killed at 

approximately 100 days of age and females at 25–30 days of age. In F1 males, DBP (500 mg/kg 

bw/day) and flutamide caused hypospadias, cryptorchidism, agenesis of the prostate, epididymis, 

and vas deferens, degeneration of the seminiferous epithelium, and interstitial cell hyperplasia 

of the testis. Agenesis of the epididymis was also observed at 250 mg/kg bw/day. Flutamide and 

DBP (250 and 500 mg/kg bw/day) also caused retained thoracic nipples and decreased anogenital 

distance. Interstitial cell adenoma occurred at 500 mg/kg bw/day in two males from the same litter. 

The only effect seen at 100 mg/kg bw/day was delayed preputial separation. The low incidence of 

DBP-induced intra-abdominal testes contrasted with the high incidence of inguinal testes seen with 

flutamide. Thus, the prenatal period is sensitive for the reproductive toxicity of DBP. Uterine and 

vaginal development in female offspring was not affected by DBP treatment. There were no signs 

of maternal toxicity with the exception of a 16% body weight loss at the time of birth and complete 

fetal mortality in 1 dam of the 500 mg/kg bw/day group. In addition, testicular focal interstitial cell 

hyperplasia and an adenoma (in 1 male) were observed in males at 500 mg/kg bw/day at 3 months 

of age. A LOAEL of 100 mg/kg bw/day was established in this study, based on delay in preputial 

separation at all dose levels. A NOAEL was not established.

To identify a NOAEL for DBP-induced developmental toxicity, Mylchreest et al. (50) gavaged 

19–20 Sprague-Dawley CD rats/group with 0, 0.5, 5, 50, or 100 mg/kg bw/day and 11 SpragueDawley CD rats with 500 mg/kg bw/day in corn oil on gd 12–21 (Table 7-12). Dams delivered 
Appendix II

II-17

and pups were weighed and examined at birth. After the pups were weaned, dams were killed and 

implantation sites and organ weights were evaluated. Pups were weighed weekly and examined 

for sexual maturation. When pups reached puberty they were killed and organ weights were 

determined. The testes and epididymides were preserved in Bouin’s solution and examined 

histologically. 

There was no evidence of maternal toxicity at any dose. In male pups, the incidence of retained 

aereolas or nipples was increased at the 100 and 500 mg/kg doses (31% of rats in 16/20 litters 

and 90% of rats in 11/11 litters, respectively). Malformations observed in the highest dose group 

included: hypospadias (9% of rats in 4/11 litters); and agenesis of the epididymis (36% of rats 

in 9/11 litters), vas deferens (28% of rats in 9/11 litters), and prostate (1/58 rats). Reduced testis, 

epididymis, prostate, and levator muscle weight and reduced anogenital distance in males were 

also observed at the high dose. Histological effects in high-dose males included interstitial cell 

hyperplasia (35% of rats in 3/5 litters), adenoma (1/23 rats), and seminiferous tubule degeneration 

(56% of rats in 3/5 litters). The single case of seminiferous tubule degeneration in the 100 mg/kg 

bw/day group was considered equivocal because the lesion does occur spontaneously in a small 

number of Sprague-Dawley rats. In female offspring, the age of vaginal opening and reproductive 

organ weight and histology were unaffected. A developmental NOAEL and LOAEL of 50 and 100

mg/kg bw/day, respectively, and a maternal NOAEL of 500 mg/kg bw/day, were identified for this 

study. 

The qualitative findings of Mylchreest et al. (47, 49, 50) were confirmed by Gray et al. (41) who 

gavaged 8–10 Sprague-Dawley rats/group from gd 14 to lactation day 3 with corn oil vehicle or 

DBP at 500 mg/kg bw/day, and groups of 4−6 Long Evans Hooded rats with 0 or 500 mg/kg bw/

day on gd 16–19.

Gray et al. (41) also compared the effects of DBP at 500 mg/kg bw/day and an equimolar 

concentration of 750 mg/kg bw/day DEHP administered by gavage to 8–10 Sprague-Dawley 

rats/group from gd 14 to lactation day 3 (Table 13). The male F1 pups were evaluated for sexual 

maturation and were then killed and necropsied at 5 months of age. Organ weights were measured 

and a histological examination of reproductive organs (preserved in Bouin’s) was conducted. The 

presence or absence of maternal toxicity was not described. Effects in F1 males are summarized 

in Table 6 and included reduced anogenital distance, and increases in percent areolas and nipples 

at birth, numbers of areolas and nipples at birth and adulthood, hypospadias, and testicular and 

epididymal atrophy or agenesis. A decrease in weight for prostates, epididymides, testes, penis, and 

the levator ani muscle was also observed in the treated rats. None of the control pups were found 

to have nipple development, malformations, or testicular degeneration. DEHP and DBP exposure 

resulted in effects that were qualitatively similar. Several males from DEHP-treated dams also 

had hemorrhagic testes. The authors stated that DEHP was considerably more toxic to the male 

reproductive system than DBP.
II-18

Appendix II

Table 6: Comparison of Reproductive Effects following in Utero Exposure to Equimolar 

Concentrations of DEHP (750 mg/kg bw) and DBP (500 mg/kg bw) in Sprague Dawley Rats

Effect Control

Chemical

DEHP DBP 

Anogenital distance (mm) 3.7±0.09 2.45±0.11* 2.79±0.09*

Areolas at birth (%) 0 88±12 55±14

Number of areolas at birth 2.7±0.75 8.4±15 2.7±0.75

Retained nipples at birth 0 8.1±1.4* 2.2±0.8*

Number of nipples at necropsy 0 8.1±1.4* 2.2±0.8*

Hypospadias (%) 0 67±14 6.2±6.2

Vaginal pouch (%) 0 45±17 0

Ventral prostate agenesis (%) 0 14±14 0

Testicular & epididymal 

atrophy or agenesis (%)

0 90±10 45.8±12

*Statistically significant.

(41)

In an abstract, DBP was reported to have been evaluated for developmental toxicity in amphibian 

and non-rodent mammalian test systems (51). Xenopus laeris (African clawed toad) tadpoles 

were exposed to 0 (n=14) or 10 (n=52) ppm DBP beginning at 2 weeks of age (stage 52) through 

complete metamorphosis (stage 66), with mortality and time to complete metamorphosis 

monitored weekly. Mortality at 10 ppm was 85% in week 1 (0% in controls) and 92% in week 

16 (28% in controls). Seventy-five percent of the controls were metamorphosed by week 12 

with 100% by week 14; none of the treated tadpoles completed metamorphosis until week 16. 

The authors concluded that DBP or its metabolite(s) may disrupt thyroid hormone cascade, since 

metamorphosis, a thyroid hormone-dependent event, is affected at 10 ppm. The same group 

administered DBP in corn syrup at 0 or 400 ppm/kg body weight to pregnant Dutch belted rabbits, 

6 does/group, on gd 15−30. Does were allowed to litter and male pups were monitored until 12 

weeks of age. At 12 weeks of age, body, testes, and epididymides weights were unaffected, but 

accessory gland weights and anogenital distance were lower in treated male offspring. In addition, 

analogously to male rats effects, one treated rabbit had undescended testes, ambiguous external 

genitalia, hypospadias, and was missing (agenesis of) the prostate and bulbourethral glands. The 

authors concluded that DBP disrupts androgen-dependent developmental events and is consistent 

with anti-androgenic effects of DBP observed in rodents after perinatal exposure.

MBuP 

Imajima et al. (52) gavaged pregnant Wistar-King A (WKA) rats with MBuP in sesame oil at 0 or 

300 mg/day on gd 15–18 (equivalent to approximately 1,000 mg/kg bw/day based on actual rat 

body weights) (Table 7-17). Male offspring were evaluated on gd 20 and on pnd 30–40 to determine 

the position of the testes. In control males, all the testes were located in the lower abdomen on gd 

20 (19 pups, 3 litters) and had descended into the scrotum on pnd 30–40 (15 pups, 3 litters). In stark 

contrast, in males exposed in utero to MBuP, all testes were located high in the abdominal cavity 

(15 pups, 3 litters) with significantly higher testes ascent on gd 20. On pnd 30–40, MBuP-exposed 
Appendix II

II-19

males exhibited cryptorchidism (22 of 26 pups, 5 litters) with uni- or bi-lateral undescended testes; 

87% of the undescended testes were in the abdominal cavity, the remaining 13% were located at the 

external inguinal ring. Testis descent is under androgenic control; the authors suggest that phthalate 

esters may interfere with FSH stimulation of cAMP accumulation in Sertoli cells, resulting in 

the reduced secretion of Mullerian inhibiting substance, a putative mediator in trans-abdominal 

migration of the testis.

The summary for Section 3 is located in Section 5.1.3.
II-20

Appendix II

4.0 REPRODUCTIVE TOXICITY 

4.1 Human Data

The relationship between either human sperm density or total number of sperm and DBP

concentration in the cellular fraction of ejaculates was studied in a group of unselected college 

students (53). A negative correlation between DBP concentration and the studied sperm indices 

was found. The authors point out that there was no reason to believe that any of the students 

examined had been exposed to phthalate esters other than at ambient levels in the environment. 

However, the use of this study to support a causal relationship to DBP exposure is limited because 

subjects’ characteristics and other potential risk factors that could confound or modify the observed 

association were not taken into account by the authors.

4.2 Experimental Animal Toxicity

Approximately 20 studies were reviewed in the evaluation of the reproductive toxicity of DBP. 

Collectively, these studies predominantly used rodents, and built on the original observation 

that DBP produced testicular atrophy in a sub-acute toxicity study (54). The literature contains 

numerous redundant studies, usually at high doses (e.g., 2 g/kg, usually in rats), all of which show 

similar effects on the testis. For example, Gray et al. (55) reported on the testicular effects of DBP 

in the adult rat, mouse, guinea pig, and hamster. In these studies, DBP was administered by gavage 

for 7 or 9 days at doses of 2,000 or 3,000 mg/kg bw/day. Severe effects were seen on testis weight 

with histopathological damage (reduction in spermatids and spermatogonia) affecting almost all 

tubules. Mouse testis was less severely affected and no effects were observed in hamsters. The 

monoester of DBP was also essentially without effect in the hamster. As discussed in Section 2.1.2 

of this monograph, sub-chronic oral exposure of adult F344 rats resulted in testicular lesions at 

doses of 720 mg/kg bw/day and higher (14). A second study (14) demonstrated that exposure to 

DBP during gestation and lactation did not increase sensitivity in rats exposed to DBP for 3 months 

during adulthood. Sub-chronic studies in B6C3F1 mice at doses up to 3,689 mg/kg bw/day did not 

cause histological or organ weight changes in the testes.

A number of more specific studies in the rat have attempted to investigate the mode of action of 

DBP using in vivo and in vitro protocols. The papers summarized here illustrate important facets of 

DBP-induced reproductive effects.

The key study for the quantitative assessment of the reproductive toxicity of DBP is reported by 

Wine et al. (38) (Table 7-14). CD Sprague Dawley rats, 10 weeks old at the start of exposure, were 

used for continuous-breeding phase and cross-over mating studies. There were 20 breeding pairs 

in each treated dose group, and 40 pairs in the control group. DBP was mixed with feed to levels 

of 0, 0.1, 0.5 and 1.0% (w/w); this yielded calculated doses of 0, 52, 256, and 509 mg/kg bw/day 

for males and 0, 80, 385, and 794 mg/kg bw/day for females. Following a 7-day premating period, 

the rats were housed as breeding pairs for 14 weeks. Litters were removed immediately after birth. 

Endpoints in-life included clinical signs, parental body weight and food consumption, fertility 

(numbers of pairs producing a litter/total number of breeding pairs), number of litters/pair, number 
Appendix II

II-21

of live pups/litter, proportion of pups born alive, sex ratio, and pup body weights within 24 hours of 

birth. 

In the F0 generation there was no effect on the overall fertility of the breeding pairs (i.e., the ability 

to produce litters with at least one live pup); all produced approximately five litters. There was clear 

indication that DBP, when administered in the diet, affected total number of live pups per litter in 

all treated groups (reduced by ~ 8–17%) and live pup weights in the 256−385 and 509−794 mg/kg 

bw/day groups by 6–12 %. 

A cross-over mating study was conducted between the high-dose treatment group and the controls. 

The percent of pairs mating, becoming pregnant, and delivering a litter was unaffected, as was litter 

size, although adjusted live pup weight was reduced in litters from treated females. At F0 necropsy, 

there were no gross or histopathologic effects in the reproductive tracts of treated animals. 

Epididymal sperm count, testicular spermatid number, and estrous cycle length were not affected by 

DBP treatment in the F0 animals. Systemic effects in the F0 rats included decreased body weight in 

females and increased liver and kidney to body weight ratios in both sexes of the high-dose group.

The final F1 litters following the continuous F0 breeding phase were weaned and raised to sexual 

maturity (pnd 88) and received the same dose in feed as their parents. Upon reaching sexual 

maturity, 20 non-sibling F1 males and females within the same treatment group were housed in 

pairs for 1 week and then housed individually until delivery of an F2 litter. 

F1 pup weight was significantly reduced in the high-dose group on pnd 0, 14, and 21. During 

rearing, three high-dose males were found to have small and malformed prepuces and/or penises 

and were without palpable testes. Mating, pregnancy, and fertility were significantly lower in the 

high-dose F1 group with only 1 of 20 pairings resulting in a litter. While litter size was unaffected, 

F2 pup weight was reduced in all treatment groups. All dose groups were killed and necropsied, 

at which point the body weights of the high-dose animals were 8–14 % lower than controls, but 

unchanged at other dose levels. For males only, kidney to body weight ratio increased at the 

256−509 mg/kg bw/day levels and liver to body weight ratio was increased at the highest level. The 

relative weights of the ventral prostate and seminal vesicles and the absolute weight of the right 

testis were decreased in the F1 males from the high-dose group. There were no effects on the ovary 

of F1 females. Epididymal sperm count and testicular spermatid count was significantly reduced in 

the high-dose F1 males. Histologic analysis was only performed on selected males (n=10) from the 

control, mid-, and high-dose groups (the solution used to preserve testes is not clear). Widespread 

seminiferous tubular degeneration was noted in 1/10 controls, 3/10 in the mid-dose group, and 8/10 

in the high-dose group. The high-dose group also exhibited interstitial cell hyperplasia. Five of ten 

high-dose males also had underdeveloped or defective epididymides. No ovarian or uterine lesions 

were noted in F1 females and there was no effect on ante-mortem estrous cyclicity.

In Wine et al. (38), the F1 high-dose group had a high rate of infertility, the middle dose had fewer 

(F0 mating) and lighter pups (F0 and F1 matings), while the low-dose animals had fewer pups 

(F0 mating) and lighter pups (F1 mating). Thus, a NOAEL was not established. The LOAEL was 

52–80 mg/kg bw/day based on reductions in F0 litter size and F2 pup weight. The Expert Panel’s 

confidence in the quality of the study is high, and our confidence is also high that these doses 
II-22

Appendix II

correctly represent the LOAEL.

A multigeneration reproductive study was conducted to assess effects of DBP exposure in Long 

Evans Hooded rats (41) (Table 15). Weanling male and female rats of the parental (F0) generation 

(10–12/sex/group) were gavaged daily with DBP in corn oil through puberty, adulthood, mating, 

gestation, and lactation. Females received 0, 250, or 500 mg/kg bw/day; male rats received 0, 

250, 500, or 1,000 mg/kg bw/day. Sexual maturation and estrous cycles of the F0 were evaluated. 

Treated rats were mated with untreated controls. When the F1 litters were weaned, the parental rats 

were killed and necropsied. Implantation sites, serum hormone levels, organ weights, and testicular 

histology were evaluated.

A delay in puberty was observed in all treated F0 males based on the age of preputial separation 

(42.6, 43.4, and 44.4 days from low to high-dose group vs 39.6 days in control group). Fertility 

was reduced in F0 males and females in the 500 mg/kg bw/day group. Infertility in F0 males was 

apparently due to testicular atrophy and reduced sperm counts. F0 females in the 500 mg/kg bw/day 

group cycled and mated successfully, but experienced an increased incidence of mid-term abortion. 

Malformations were significantly increased in F1 pups from the 250 and 500 mg/kg bw/day groups. 

Types of malformations included low numbers of hypospadias, abdominal testes, anophthalmia, 

uterus unicornous, and renal agenesis. 

The F1 pups were not treated with DBP after weaning. Four to eighteen pairs of F1 pups from 

treated dams were selected for continuous mating within dose groups for 11 cycles and the 

remaining F1 pups were necropsied. The F2 pups born during the continuous breeding phase 

were counted and discarded. Fecundity was reduced in F1 rats from treated dams and the number 

of F2 pups born was reduced in breeding pairs from the 250 and 500 mg/kg bw/day groups. At 

necropsy, a non-significant reduction in caudal sperm counts (19%) and a significant reduction in 

caudal sperm levels (34%) were noted in F1 males from the 250 and 500 mg/kg bw/day groups, 

respectively. 

The study by Gray et al. (41) is somewhat limited because many endpoints and details of their 

experimental methods were not reported.

In Lamb et al. (39) and Reel et al. (40) (Table 7-16), DBP was one of four phthalate esters 

compared using the Continuous Breeding protocol in CD-1 mice; the same basic protocol as 

reported in Wine et al. (38). Male and female CD-1 mice, 20 pairs/treatment group and 40 pairs 

in control, were fed a diet with DBP at 0 300, 3,000, or 10,000 ppm (doses of 53, 525, and 1,750

mg/kg bw/day as reported by Reel et al. (40)) for 7 days prior to and during a 98-day cohabitation 

period. Litters were removed immediately after birth. Reproductive function was evaluated during 

the cohabitation period by measuring the numbers of litters per pair and of live pups per litter, pup 

weight, and offspring survival. Testes were fixed in Bouin’s solution for histological evaluation. 

DBP exposure reduced litter size, numbers of litters per pair, number of fertile pairs, live pups per 

litter, and proportion of pups born alive in the high-dose group. These effect were not seen at lower 

dose levels. A crossover mating trial demonstrated that female, but not male, mice were affected by 

DBP, as shown by significant decreases in the percentage of fertile pairs, the number of live pups 

per litter, the proportion of pups born alive, and live pup weight. Only the control and high-dose F0
Appendix II

II-23

DBP groups were necropsied. There were no effects on sperm parameters in the males, although 

body weight was significantly decreased (8%) and liver to body weight ratio significantly increased 

(11%). For females, liver to body weight ratio was significantly increased (19%) and relative uterine 

weight significantly decreased (28%), but there was no effect on estrous cycles. No treatmentrelated gross or histological lesions were noted. A second generation was not evaluated.

In Lamb et al. (39), the high-dose group was subfertile and the middle-dose and the low-dose 

groups were functionally unaffected. Thus, the NOAEL was calculated at 525 mg/kg bw/day, based 

on reductions in litter size and in proportions of pairs having litters. The mid- and low-dose groups 

were not necropsied or evaluated for reproductive development or performance. For these reasons, 

the Expert Panel has moderate-to-low confidence that these doses correctly represent the LOAEL 

and NOAEL. Confidence in the quality of the data reported is high. 

Mode of Action

The Expert Panel believes that data from studies with DEHP are relevant to a consideration of 

the mechanism by which DBP causes adverse effects. It is well understood that DEHP produces 

a range of hepatic effects in rats (induction of peroxisomes; increased Cyp4A1; PCoA) including 

hepatic tumors. The induction of these effects in rats is believed due to activation of PPAR-alpha. 

In PPAR-knockout mice, administration of DEHP does not result in the induction of hepatic effects 

or tumors unlike the wild-type control animals. In humans, PPAR-alpha is activated upstream of 

different enzymes from those noted in the rat. Recently, an IARC review of the cancer issue led 

them to conclude that DEHP rat tumor data was of limited relevance to human risk. 

In studies with DEHP, a genetically-modified strain of mouse (the PPAR-alpha knockout mouse) 

cannot activate PPAR-alpha, but is susceptible to phthalate-induced developmental toxicity and 

testicular toxicity. This mouse does express PPAR-gamma in the testis which can be activated by 

MEHP (56). PPAR-gamma may conceivably play a role in the reproductive toxicity of phthalates. 

PPAR-gamma has been found in human testis, ovary, placenta, and embryo. Other members of 

the PPAR family (beta and gamma) have not been extensively studied with regard to activation by 

phthalates.

Finally, the guinea pig, a non-responding species to the peroxisomal proliferating effects of DBP, is 

susceptible to the testicular effects of this phthalate. 

Gray et al. (55) investigated the reason for the lack of testicular lesions in hamsters orally 

administered DBP and MBuP at doses exceeding those that produced testicular lesions in rats. 

Using 14C-labelled DBP and monobutyl ester (MBuP), it was determined that intestinal esterase 

activities were similar in the two species and that the principal metabolite in the rat and hamster 

was MBuP glucuronide (23) However, the levels of unconjugated MBuP in urine were 3–4 fold 

higher in the rat. Finding that the activity of testicular beta-glucuronidase was significantly higher 

in the rat than the hamster, the authors speculated that the testicular damage might be associated 

with greater concentrations of unconjugated MBuP, the putative toxicant.

All phthalates that cause testicular toxicity produce a common lesion characterized by alterations 

in Sertoli cell ultrastructure and function (57-59). It is known that some Sertoli cell functions are 
II-24

Appendix II

mediated by follicle stimulating hormone (FSH) interaction with membrane bound receptors. Lloyd 

and Foster (60) demonstrated that MEHP disturbs FSH interaction with the FSH receptor. Further 

studies with MEHP using primary rat Sertoli cell cultures revealed that the monoester of DEHP 

inhibited FSH-stimulated cAMP accumulation. The MEHP-induced inhibition was specific for FSH 

(61).

Factors affecting increased sensitivity to phthalate-induced testicular toxicity in young animals 

were studied for DBP, DEHP, di-n-hexyl phthalate (DnHP), and dipentyl phthalate. The monoester 

derivatives of DBP and DEHP have been shown to cause similar testicular effects. Sjoberg et al. 

(62) demonstrated that gavage treatment with DEHP resulted in greater absorption of MEHP, and 

hence, a greater systemic dose to young versus mature rats. Further, in vitro studies did not find that 

FSH-stimulated cAMP accumulation and lactate secretion were age related (63). Lloyd and Foster 

(60) noted that initiation of spermatogenesis was dependent on FSH interaction with the Sertoli 

cell in young rats, but was not necessary for maintenance of spermatogenesis in adults. Their 

experiment in Sertoli cell cultures demonstrated that MEHP interferes with FSH interaction at the 

receptor level and provided a hypothesis for increased sensitivity to testicular toxicity in young 

animals.

The Panel was not able to reach agreement that interfering with FSH signaling function was the 

accepted mode or mechanism of action.

Several studies have examined the ability of selected phthalate esters to compete with labeled 

estradiol (E2) for binding to the estrogen receptor (ER). Sources of ER protein included rat uterine 

(64), rainbow trout hepatic cytosol (65), recombinant human ERs (rhER) overexpressed in SF9 

insect cells using the baculovirus system (66, 67) and rainbow trout ERs expressed in yeast . Triated 

E2 was used in the tissue cytosol binding assays while a high affinity fluorescent E2 derivative was 

used in the rhER binding assays. DBP exhibited no or weak activity in in vitro assays that measured 

binding of phthalates to estrogen receptors (64, 65, 68). The assays did not include the addition of 

esterases or lipases to metabolize DBP to its monoester. 

Selected phthalate esters have been examined in a number of in vitro gene expression assays 

systems. The assays have used stably transfected cells (64), transiently transfected cells (64, 65), 

yeast based assays (64, 68-70) and vitellogenin induction in rainbow trout hepatocyte cultures (68). 

DBP was weakly active in an assay of estrogen-induced gene expression, but its metabolite MBuP 

was inactive (70). There was no synergism in estrogenic response with DBP and other phthalates 

(70, 71). 

In vivo assays demonstrated that DBP does not increase uterine wet weight or vaginal epithelial cell 

cornification in immature or mature ovariectomized rats (64) and prepubertal mice (69). Uterine 

permeability was not affected following the subcutaneous injection of DBP (71). Malformations 

in reproductive organs and effects on androgen-related endpoints of male rats exposed to DBP or 

MBuP during prenatal development suggest antiandrogenic activity by DBP and MBuP (41, 49, 50, 

52).

The summary for Section 4 is located in Section 5.1.4.
Appendix II

II-25

5.0 DATA SUMMARY & INTEGRATION 

5.1 Summary

5.1.1 Human Exposure

The major use of DBP is as a coalescing aid in latex adhesive. It is also used as a plasticizer for 

cellulose plastics and as a solvent for dyes. DBP is not used as a plasticizer for PVC plastics (1). 

Several authoritative estimates of human exposure, described in Section 1, have been published 

since 1990. All estimates place total DBP exposure in the general population at less than 10 μg/kg 

bw/day and were consistent in identifying food as the major exposure source. In addition to food, 

general human exposure occurs primarily through indoor air followed by drinking water, soil, and 

ambient air. Infants and young children may have higher exposures than adults, primarily because 

of dietary differences and possible mouthing of DBP-containing household articles (not limited to 

toys). Using reasonable assumptions and data from surveillance and food surveys, Health Canada 

(6) estimated total exposures of 2.4, 5.0, 4.3, 2.3, and 1.9 μg/kg bw/day for humans aged 0−0.5, 

0.5−4, 5−11, 12−19, and 20−70 years, respectively. Discrepancies in food exposure estimates may 

be due to inherent variability of food eaten by individuals based on age, sex, ethnicity, time of 

sampling, and geographical locations. 

DBP was found in infant formula, but amounts vary internationally and seem to be falling (5, 9). 

The most recent estimate of DBP exposure from infant formula to a newborn in the UK is 2.4 

μg/kg bw/day (5) and is the same as the Health Canada total exposure estimate. DBP has been 

found in 1 of 17 European children’s toys at a very low level (0.01% by weight) (10). Use of DBP 

in plastic nasogastric tubing has also been reported (3). Occupational exposure during phthalate 

manufacture is estimated at 143 μg/kg bw/day. Exposures in other occupational settings have not 

been estimated.

5.1.1.1 Utility of Data to the CERHR Evaluation

DBP exposures resulting from contact with various media (e.g., food, drinking water, and air) have 

been estimated by several authoritative sources. Limitations in the dataset include the fact that 

most of the data used in calculations were 15–20 years old and may not reflect current exposure. 

Further, the majority of data was collected in Europe and Canada and may not accurately reflect 

US patterns. Data from Health Canada were selected for use since they provide age-based exposure 

estimates. 
II-26

Appendix II

5.1.2 General Biological and Toxicological Data

Toxicity. The Expert Panel had to rely on animal toxicity data in its evaluation of general biology 

and toxicity. DBP is not acutely toxic to rodents with the oral LD50 given in grams per kilogram 

(g/kg) quantities. There are sufficient data to establish that DBP in the diet is toxic to adult rats 

and mice at repeated daily doses of ∼350 mg/kg bw/day and higher (14). The liver and testes 

are consistently found to be target organs with the hematopoetic system also affected in some 

strains of rats and at higher doses in mice. Testicular lesions were observed at doses of 720 mg/

kg bw/day and higher in adult rats (14). DBP increases liver to body weight and kidney to body 

weight ratios. These effects are consistent with effects seen with other phthalates. Indications of 

peroxisome proliferation, such as elevated levels of PCoA oxidation, were consistently observed. 

The lowest repeated dose NOAEL in rats was observed in males exposed through diet to 142 mg/

kg bw/day (13). The corresponding NOAEL in male mice was 353 mg/kg bw/day (14). Chronic 

carcinogenicity studies were not available for review.

Table 7: Summaries of NOAELs and LOAELs and Major Effects in General Toxicity Studies

Protocol and DBP Doses (mg/kg 

bw/day) 

NOAEL

(mg/kg bw/day)

LOAEL (mg/kg bw/day) 

and Effects

Major Effects at Higher Doses

3-month repeat dose dietary study 

in Wistar rats. 

6 weeks old at start of study, 

10 rats/sex/group. 

Doses – M: 0, 27, 142, 688;

F: 0, 33, 162, 816.

(13)

M: 142

F: 162

M: 688; F: 816

↑Liver and kidney weight 

(F).

Peroxisomal proliferation.

Histological liver changes.

↓Thyroid hormone.

Anemia (M).

No testicular lesions.

No higher doses in study.

13-week repeat-dose dietary study 

in F344 rats.

5−6 weeks old at start of study, 10 

rats/sex/group.

Doses – M: 0, 176, 359, 720, 

1,540, 2,964

F: 0, 177, 356, 712, 1,413, 2,943.

(14)

M: 176

F: 177

M: 359; F: 356

↑ Liver and kidney 

weights (M).

Peroxisomal proliferation.

Anemia (M).

↑ Liver and kidney weights.

Hepatic lesions.

Changes in liver enzyme activity.

Peroxisomal proliferation.

Testicular lesions.

Hypospermia.

↓Testes weight.

↓ Testicular testosterone levels. 

Anemia (M).

13-week repeat-dose dietary study 

in B6C3F1 mice.

6 weeks-old at start of study, 10 

mice/sex/group.

Doses – M: 0, 163, 353, 812, 

1,601, 3,689

F: 0, 238, 486, 971, 2,137, 4,278.

(14)

M: 353

F: None

M: 812

F: 238

↑ Kidney weight (F)

(No dose response or 

histological changes).

↑ Liver weight (M).

↓Body weight gain (M).

↑ Kidney weight (F)

(No dose response or histological 

changes).

↑ Liver weight.

↓Body weight gain.

Mild histological liver effects.

No testicular lesions.
Appendix II

II-27

Toxicokinetics. There are no conclusive in vivo toxicokinetic data in humans. Orally-administered 

DBP in rodents is rapidly hydrolyzed to the monoester, MBuP, by pancreatic lipases secreted 

into the small intestine. The monoester is rapidly absorbed from the gut, widely distributed in 

tissues, and is rapidly excreted in urine, mainly as a glucuronide. No studies are available on the 

absorption of orally-administered DBP in primates. Thus, it is not known whether DBP is more 

poorly hydrolyzed and absorbed in the gut of primates compared to rats, as has been observed 

with other phthalates. Rodent studies indicate there is no bioaccumulation of absorbed DBP or its 

metabolites (including testes and prostate tissue). In vitro human and rat skin were compared for 

their absorption of DBP; and human skin was found to be much less permeable than rat skin (12). 

In rats, dermal absorption of DBP as identified by urinary excretion of metabolites is 10–12% of 

the 30–40 mg/kg dose per day (11). 

Rats treated with 14C-DBP on gd 14 showed concentrations of radioactivity in placenta and fetuses 

that were approximately 65% of the levels in maternal serum. MBuP was the major metabolite 

found in both maternal and embryonic tissues (21). 

A PBPK model of the tissue distribution of DBP and MBuP in rats has been developed by Keys 

et al. (25); the model includes diffusion limitations and pH trapping as mechanisms of uptake 

of MBuP into tissue. A model has been derived to extrapolate rodent data to predicted values in 

humans. The model does not contain parameters for estimating fetal or pediatric values.

Genetic Toxicity. IPCS (3) reviewed a number of mutagenicity and related endpoints for DBP and 

concluded that the weight of the evidence indicated that DBP is not genotoxic. 

5.1.2.1 Utility of Data to the CERHR Evaluation

The oral subchronic studies in rats and mice are adequate for the evaluation of general toxicity 

induced by DBP. Some studies were conducted according to GLP standards and relevant exposure 

routes were utilized. Small group numbers, used in some studies, are of limited concern considering 

the reproducibility of effects between studies. Adult rodents were tested for DBP-induced testicular 

lesions. Sections 3 and 4 of this document address studies where the male rodent reproductive tract 

was exposed to DBP during prenatal and postnatal development. The examination of hepatic effects 

was adequate and included an evaluation of peroxisome proliferation in rodents. 

There are acceptable toxicokinetic data for DBP, consisting of absorption, distribution, metabolism, 

and excretion, following oral and dermal exposure in the rat. The human data available are of very 

limited utility. In vitro comparisons of DBP metabolism suggest that effects observed in rodents are 

relevant to humans. 

5.1.3 Developmental Toxicity

The are no data on the developmental toxicity of DBP in humans. The most complete description of 

effects characterizing key aspects of the developmental toxicity of DBP is contained in a series of 
II-28

Appendix II

publications by Ema et al. (34-36) and Mylchreest et al. (47, 49, 50). Ema et al. (42) characterized 

the prenatal developmental toxicity of DBP in Wistar rats and subsequently demonstrated that the 

metabolite MBuP caused developmental toxicity similar to DBP. These effects were produced 

at approximately equimolar concentrations. For example, a maternal and development NOAEL

and LOAEL of 500 and 630 mg/kg bw/day (1.80 and 2.27 mmol/kg bw/day), respectively, were 

identified for DBP following gavage of Wistar rats on gd 7–15 (34). Using a similar experimental 

design, a maternal and developmental NOAEL and LOAEL for MBuP of 250 and 500 mg/kg bw/

day (1.13 and 2.25 mmol/kg bw/day), respectively, were identified (42). Similar fetal effects in 

these studies included increased prenatal mortality, decreased fetal weight, and cleft palate. Dose 

and time dependency studies with DBP and MBuP resulted in similar findings and are described in 

Section 3.2.1. 

The most complete prenatal exposure study by Ema et al. from the perspective of group size and 

development of the male reproductive tract established a maternal and fetal NOAEL and LOAEL 

of 331 and 555 mg/kg bw/day, respectively, in Wistar rats fed DBP-dosed diets on gd 11–21 (36). 

Developmental effects at higher doses (≥555 mg/kg bw/day) included decreased fetal weight, cleft 

palates, fused sternebrae, reduced anogenital distance in males, and cryptorchidism. 

A group of studies from Mylchreest et al. looked at postnatal effects following in utero exposure 

to DBP (47, 49, 50). CD rats were gavaged with DBP from gd 3 to pnd 20 or gd 12–21. Delayed 

preputial separation and retained nipples were observed at doses as low as 100 mg/kg bw/day. 

Effects noted at doses of 250 mg/kg bw/day or higher were consistent between studies and included 

hypospadias, agenesis of epididymides or seminal vesicles, cryptorchidism, decreased anogenital 

distance in males, and/or a low incidence of interstitial adenomas. A NOAEL of 50 mg/kg bw/day 

was identified. The three Mylchreest studies (47, 49, 50) exposed animals during the appropriate 

window of development, analyzed the tissues appropriately, and combined them with other indices 

of puberty and reproductive development. The concordance in dose-response to the Wine et al. (38) 

study is good. 

The role of the monoester metabolite of DBP in developmental toxicity was elucidated by 

Saillenfait et al. (21), who gavaged Sprague-Dawley rats with 500 or 1,500 mg/kg of radiolabeled 

DBP/kg bw/day on gd 14. They demonstrated radioactivity in placentas and embryos at levels of 

21–30% of those measured in maternal plasma. The majority of the radioactivity was associated 

with MBuP and its glucuronide. Postnatal effects following in utero exposure to the DBP 

metabolite MBuP were studied in WKA rats that were gavaged with 300 mg MBuP/day (∼1,000 

mg/kg bw/day) on gd 15–18 (52). Testes descent was reduced on both gd 20 and pnd 30–40. 

Although only one dose was administered, the findings are consistent with those observed in DBP 

developmental toxicity studies conducted by Ema et al. (36) and Mylchreest et al. (47, 49, 50), thus 

supporting the hypothesis that MBuP is responsible for effects associated with DBP exposure. 

The hallmark of developmental toxicity in the mouse following oral exposure to DBP appears to be 

primarily systemic toxicity and death. In a study with ICR mice exposed to diet containing DBP on 

gd 0–18, Shiota et al. (32, 33) reported a 98% incidence of fetal mortality at 2,100 mg/kg bw/day. 

Fetal body weight was reduced at 660 mg/kg bw/day. The authors stated that the maximum nonembryotoxic dose was 370 mg/kg bw/day. However, the Expert Panel noted that delayed ossification 
Appendix II

II-29

occurred at all dose levels, and selected the lowest dose, 80 mg/kg bw/day, as a LOAEL. These 

data are from groups with small sample size and have not been replicated. In a continuous breeding 

protocol with CD-1 mice, Lamb et al. (39) observed a decrease in the number of pups, live pups 

per litter, and pup weight in dams that consumed a dose of 1,750 mg/kg bw/day in the diet. The 

developmental NOAEL was identified as 525 mg/kg bw/day. Effects of in utero and lactational 

exposure to DBP were studied in B6C3F1 mice where Marsman et al. (14) reported that length of 

gestation was increased at 2,500 ppm (454 mg/kg bw/day) and higher. Seventy-five and ninety-five 

percent of litters were lost at 10,000 (1,816 mg/kg bw/day) and 20,000 (3,632 mg/kg bw/day) ppm. 

Decreases were observed in litter size and pup body weights at 2,500, 7,500, and 10,000 ppm. The 

Expert Panel is not confident that these three studies fully assessed DBP developmental toxicity, 

including reproductive function, due to limitations in study design that include small group size, 

failure to perform necropsies in critical dose groups, and failure to assess appropriate landmarks of 

maturation. 

NOAELs and LOAELs for the key developmental toxicity studies for DBP are listed in Table 

8. The Ema et al. (36) study examined the most sensitive prenatal endpoints and allows for a 

comparison between maternal and developmental toxicity. The Ema et al. (34) study of DBP was 

also included to allow comparison with the study of its metabolite, MBuP (42) that was evaluated 

according to the same protocol. The Mylchreest et al. (50) study is considered key because it 

examined the most sensitive endpoints at the lowest doses.
II-30

Appendix II

Table 8: Summaries of NOAELs and LOAELs and Major Effects in Key Developmental Toxicity Studies

Protocol and Study NOAEL

(mg/kg bw/day)

LOAEL

(mg/kg bw/day) Developmental Effects 

Observed at Higher Dose 

Levels

Maternal Developmental

Prenatal studies in 

Wistar rats.

11–12/group received 

DBP (0, 500, 630, 

750, or 1,000 mg/kg 

bw/day) or MBuP 

(0, 250, 500, or 625 

mg/kg bw/day) on gd 

7–15 by gavage. 

In a third study rats 

were treated by diet 

with

 0, 331, 555, or 661 

mg/kg bw/day on gd 

11–21. 

Fetuses were evaluated 

late in gestation.

(34, 36, 42)

DBP Gavage:

Maternal: 500

Fetal: 500

(1.80 mmol/kg 

bw/day)

MBuP Gavage:

Maternal: 250

Fetal: 250

(1.13 mmol/kg 

bw/day)

DBP Diet:

Maternal: 331

Fetal: 331

DBP Gavage:

630

(2.27 mmol/kg 

bw/day)

↓ Weight gain.

MBuP Gavage:

500

(2.25 mmol/kg 

bw/day)

↓ Weight gain.

DBP Diet:

555

↓ Weight gain.

DBP Gavage:

630

(2.27 mmol/kg bw/day)

↑ Prenatal mortality.

↓ Fetal weight.

MBuP Gavage:

500

(2.25 mmol/kg bw/day)

↑ Prenatal mortality.

↓ Fetal weight.

↑ External and skeletal 

malformations.

↑ Visceral variations.

DBP Diet:

555

↓ Anogenital distance in 

males.

↑ Fetuses with 

undescended testes.

DBP Gavage:

↑ Prenatal mortality.

↓ Fetal weight.

↑ External malformations

MBuP Gavage:

↑ Prenatal mortality.

↓ Fetal weight.

↑ External and skeletal 

malformations.

↑ Visceral variations.

DBP Diet:

↓ Fetal weight.

↑ External and skeletal 

malformations.

↓ Anogenital distance in 

males.

↑ Fetuses with undescended 

testes.

Prenatal gavage 

study with postnatal 

evaluation in

CD rats. 

11−22 per group 

received 0, 0.5, 5, 50, 

100 or 500 mg/kg bw/

day on gd 12–21.

Pups were evaluated 

until puberty.

(50)

Maternal: 500

Developmental:

50

None 100

Retained aereolas and 

nipples in males.

Retained aereolas and nipples 

in males.

Testicular lesions and 

adenoma.

Malformations of 

reproductive organ.

↓ Reproductive organ 

weights.

↓ Anogenital distance in 

males.

Prenatal dietary study 

in ICR-JCL mice.

6−15 dams per treated 

group received 0, 80, 

180, 350, 660, and 

2,100 mg/kg bw/day 

on gd 0–18. 

Dams and pups 

examined late in 

gestation.

(32, 33)

Maternal: 660

Developmental:

None

2,100

↓ Body weight 

gain.

80*

Delayed ossification 

(number of ossified 

coccygia from control 

to 660 mg/kg bw/day 

group: 9.4, 5.1, 4.5, 6.0, 

2.6). 

Delayed ossification.

↑ Prenatal mortality.

↓ Fetal weight.

↑ Neural tube defects.

* Effect level selected by Expert Panel differs from that of the study authors
Appendix II

II-31

5.1.3.1 Utility of Data to the CERHR Evaluation

The data in rats are adequate for an assessment of developmental toxicity. Studies examined effects 

following dosing of dams through portions of or the entire period of pregnancy. Fetuses were 

evaluated for prenatal malformations and postnatal effects. Evaluations included an examination of 

reproductive organs and androgen-regulated endpoints, which are thought to be the most sensitive 

indicators of phthalate-induced toxicity. Prenatal effects following prenatal exposure to MBuP 

were also examined. A second rodent species, the mouse, was examined in a prenatal exposure and 

effect study. Based on the limited parameters examined in the mouse it is not possible to compare 

sensitivity in rats and mice. 

5.1.4 Reproductive Toxicity

Human Data

The relationship of human sperm density and total number of sperm to DBP concentration was 

studied in a group of unselected college students (53). A negative correlation was found between 

sperm density and DBP concentration in the cellular fraction of ejaculate of non-occupationally 

exposed subjects, but the data are of little value to the Expert Panel due to the insufficient evidence 

for a causal relationship of sperm characteristics to DBP levels.

Experimental Animal Studies

Reproductive studies have been performed primarily in the rat and, to a lesser extent, the mouse. 

There are single reports of studies in guinea pigs and hamsters (55). Collectively, the data are 

sufficient to show that oral exposure to DBP can cause reproductive toxicity in male rats, mice, and 

guinea pigs. In contrast, the hamster failed to show testicular effects. Data that characterize effects 

on female reproduction are not as complete and detailed interpretation is therefore less certain. The 

data do indicate a decrease in female fertility in mice and rats. 

Females. The Lamb et al. (39) data from a continuous breeding study in mice clearly show 

adult female functional effects at 1,750 mg/kg bw/day. The limited examination of the lower 

dose groups (necropsies were not performed) precludes the setting of a reliable NOAEL. The 

continuous-breeding study by Wine et al. (38) in F344 rats did not show specific deficits in female 

parameters; however, the data do not rule out that decreases in litter size at all doses may have a 

female component. In contrast, Gray et al. (41) reported that fertility in female Long Evans rats was 

reduced following treatment with 500 mg/kg bw/day from weaning through puberty, gestation, and 

lactation. The effect was apparently due to an increase in mid-term abortions. The F1 female pups in 

this study were also mated and experienced a reduction in fecundity at doses of 250 mg/kg bw/day 

and higher. Thus, clear effects on female reproduction are seen in rats at doses of 250 mg/kg bw/

day (LOAEL) and in mice at higher doses. NOAELs can not be established with any confidence.

Males. Data from the Wine et al. (38) continuous breeding study clearly show functional and 

structural reproductive effects in male Sprague-Dawley rats. In the F0 generation there was clear 

indication that DBP, when administered in the diet, affected total number of pups per litter in all 

treated groups. The F1 high-dose group had malformations of the reproductive tract and a high rate 
II-32

Appendix II

of infertility. Dose related increases in seminiferous tubular degeneration were seen at the 256 and 

509 mg/kg bw/day doses. The LOAEL was 52–80 mg/kg bw/day based on reductions in F0 litter 

size. Thus, a reproductive NOAEL was not established.

A delay in preputial separation was observed in Long Evans rats exposed at the lowest dose of 250 

mg/kg bw/day by gavage from the time they were weaned until the litters they sired were weaned 

in the Gray et al. study (41). At higher doses (500–1,000 mg/kg bw/day), reductions in sperm 

counts and fertility and testicular lesions were also observed. The F1 offspring exposed to DBP only 

during gestation and lactation experienced a reduction in sperm counts. 

Three studies by Mylchreest et al. (47, 49, 50), presented in Sections 3.2.2 and 5.1.3 of this 

document, indicated that the range of male structural abnormalities in the Wine et al. (38) study 

could be reproduced with a much shorter dosing regime. Mylchreest et al. (49, 50) also detected 

a significant increase in testicular Leydig cell hyperplasia and a low incidence of Leydig cell 

adenomas in ~3-month-old animals following only a late gestational exposure (gd 12–21) of 500 

mg/kg bw/day. Wine et al. (38) dosed for 14 weeks with DBP in the diet, whereas Mylchreest et al. 

(50) exposed pregnant rats by gavage during gd 12–21. A NOAEL was established at 50 mg/kg bw/

day in the Mylchreest et al. (50) study. 

The existing data show consistent effects (testicular pathology, reduced sperm numbers, effects 

on reproductive tract development), and are sufficient to conclude that DBP is a reproductive 

and developmental toxicant in male rats at doses of 100 mg/kg bw/day and higher. Treatment of 

rat weanlings with 250 mg/kg bw/day resulted in delayed puberty and doses of 500 mg/kg bw/

day induced testicular lesions. In general toxicity studies (Section 2.1.2), testicular lesions were 

observed in adult rats (6 weeks old) treated with 720 mg/kg bw/day, but not in adult mice treated 

with up to 3,689 mg/kg bw/day for 3 months (14). Histological changes in testes of 4–6 weekold mice and guinea pigs of a similar nature have also been observed following administration of 

a single high dose (2,000 mg/kg bw/day) for 7–9 days, but hamsters were unaffected. The overall 

effects on the testes indicate an age sensitivity with fetal sensitivity >pubertal sensitivity> adult 

sensitivity in male rats to the action of DBP. 

The responses that occur at the lowest doses appear to involve the development of the reproductive 

system. These responses were seen with some consistency in the studies by Mylchreest et al. (47, 

49) and Wine et al. (38). The report by Reel et al. (40) and the paper by Lamb et al. (39) did not 

report on measures of reproductive system development. However, they are consistent with the 

Mylchreest et al. and Wine et al. papers in that they show reproductive toxicity under oral (dietary) 

exposure, and do so in a second species, the mouse.

Mode of Action

Gray et al. (55) investigated the reason for the lack of testicular lesions in hamsters administered 

DBP and MBuP orally at doses exceeding those that produced testicular lesions in rats. Using 14Clabelled DBP and MBuP, it was determined that intestinal esterase activities were similar in the 

two species and that the principal metabolite in the rat and hamster was MBuP glucuronide (23). 

However, the levels of unconjugated MBuP in urine were 3–to 4-fold higher in the rat. Finding that 

the activity of testicular beta-glucuronidase was significantly higher in the rat than the hamster, the 
Appendix II

II-33

authors speculated that the testicular damage might be associated with greater concentrations of 

MBuP, the putative toxicant.

All phthalates that cause testicular toxicity produce a common lesion characterized by alterations 

in Sertoli cell ultrastructure and function (57-59). It is known that some Sertoli cell functions are 

mediated by FSH interaction with membrane bound receptors. Lloyd and Foster (60) demonstrated 

that MEHP disturbs FSH interaction with the FSH receptor. Further studies with MEHP using 

primary rat Sertoli cell cultures revealed that the monoester of DEHP inhibited FSH-stimulated 

cAMP accumulation. The MEHP-induced inhibition was specific for FSH (61).

Factors affecting increased sensitivity to phthalate-induced testicular toxicity in young animals 

were studied for DBP, DEHP, DnHP, and dipentyl phthalate. The monoester derivatives of DBP 

and DEHP have been shown to cause similar testicular effects. Sjoberg et al. (62) demonstrated that 

gavage treatment with DEHP resulted in greater absorption of MEHP, and hence, a greater systemic 

dose to young versus mature rats. Further, in vitro studies did not find that FSH-stimulated cAMP 

accumulation and lactate secretion were age related (63). Lloyd and Foster (60) noted that initiation 

of spermatogenesis was dependent on FSH interaction with the Sertoli cell in young rats but 

was not necessary for maintenance of spermatogenesis in adults. Their experiment in Sertoli cell 

cultures demonstrated that MEHP interferes with FSH interaction at the receptor level and provided 

a hypothesis for increased sensitivity to testicular toxicity in young animals.

The Panel was not able to reach agreement that interfering with the FSH-signaling function was the 

accepted mode or mechanism of action.

The Expert Panel believes that data from studies with DEHP are relevant to a consideration of 

mechanism for DBP-induced toxicity. It is well understood that DEHP produces a range of hepatic 

effects in rats (induction of peroxisomes; increased Cyp4A1; PCoA) including hepatic tumors. 

The induction of these effects in rats is believed due to activation of PPAR-alpha. In genetically 

altered mice who do not express PPAR, administration of DEHP does not result in the induction of 

hepatic effects or tumors unlike the wild-type control animals. In humans, PPAR-alpha is activated 

upstream of different enzymes from those noted in the rat. Recently, an IARC review of the cancer 

issue led them to conclude that DEHP rat tumor data was of limited relevance to human risk. 

In studies with DEHP, a genetically modified strain of mouse (the PPAR-alpha knockout mouse) 

cannot activate PPAR-alpha, but is susceptible to phthalate-induced developmental toxicity and 

testicular toxicity. This mouse does express PPAR-gamma in the testis which can be activated by 

MEHP (56). PPAR-gamma may conceivably play a role in the reproductive toxicity of phthalates. 

PPAR-gamma has been found in human testis, ovary, placenta, and embryo. Other members of 

the PPAR family (beta and gamma) have not been extensively studied with regard to activation by 

phthalates.

Finally, the guinea pig, a non-responding species to the peroxisomal-proliferating effects of DBP, is 

susceptible to the testicular effects of this phthalate. 

Imajima et al. (52) suggests that the active metabolite for reproductive effects due to gestational 

exposure is MBuP. This pattern of effects induced in rodents by late gestational exposure (gd 
II-34

Appendix II

12–21) is ‘anti-androgenic,’ in that flutamide mimics these effects (49); however, DBP/MBuP does 

not bind to the androgen receptor (72). In pubertal and adult rodents, the Sertoli cell is the likely 

cellular target for testicular injury mediated by the monoester (63, 73). 

DBP exhibited no or weak activity in in vitro assays that assess estrogenicity (64, 65, 68, 70). The 

assays did not include the addition of esterases or lipases to metabolize DBP to its monoester. 

However, the DBP metabolite MBuP was determined to be inactive in one assay (70). There was 

no synergism in estrogenic response with DBP and other phthalates (70, 71). DBP was inactive 

in rodent in vivo assays that measure endpoints such as increases in uterine wet weight, vaginal 

epithelial cell cornification, or uterine permeability (64, 69, 71). Malformations in reproductive 

organs and effects on androgen-mediated endpoints in male rats exposed to DBP or MBuP during 

prenatal development suggest antiandrogenic activity by DBP and MBuP (41, 49, 50, 52).
Appendix II

II-35

Table 9: Summaries of NOAELs and LOAELs and Major Effects in Reproductive Toxicity Studies

Protocol & Study

NOAEL

(mg/kg bw/day)

LOAEL

(mg/kg bw/day)

and Effects

Reproductive Effects 

Observed at Higher Dose 

Levels

Reproductive Systemic

Dietary continuous 

breeding protocol with 

crossover breeding and 

evaluation of second 

generation in SpragueDawley rats. 

20 pairs per group were 

treated at doses of M: 0, 

52, 256, or 509 mg/kg bw/

day; F: 0, 80, 385, or 794 

mg/kg bw/day during a 14 

week mating period.

(38)

Reproductive: None

Systemic: 

256 (M); 385 (F)

M: 52; 

F: 80

↓ F1 live litter size.

↓ F2 pup weight.

M: 509; F: 794

↓Body

weight gain in 

F0 females and 

F1 males and 

females.

↑ Liver and 

kidney weight 

in F0 males and 

females and F1

males.

↑ Malformed 

reproductive organs in F1

males.

↓ Mating, pregnancy, 

and fertility in F1.

↓ Reproductive organ 

weights in F1 males.

↑ Testicular lesions in F1

males.

↓ Sperm counts in F1.

↓ F1 litter size.

↑ F1 pup mortality.

↓ F1 and F2 pup weight.

Dietary continuousbreeding protocol with 

crossover mating in CD-1 

mice. 

20 pairs per group were 

treated with 0, 53, 525, and 

1,750 mg/kg bw/day during 

a 14-week mating period.

(39, 40)

Reproductive: 

(M): 525

(F): 525

Systemic: Not 

known because only 

high-dose group was 

necropsied. 

M: Can’t determine

F: 1,750

↓ Fertility in F0

females.

↓ Uterine weight in 

F0.

↓ Live pups/litter.

No effects on sperm 

in F0.

1,750

↓ Bodyweight in 

males.

↑ Liver weight.

No higher doses.

Multigenerationreproductive study in 

Long Evans Hooded rats. 

10–12 pairs per group 

were treated by gavage 

from weaning throughout 

puberty, adulthood, mating, 

and lactation with 0, 250 or 

500 mg/kg bw/day. Males 

were also dosed with 1,000 

mg/kg bw/day.

F1 rats were not treated 

following weaning.

(41)

Reproductive:

None

Systemic: Not 

reported

250

Delayed puberty in F0

males.

↓ Sperm production 

in F1 males (nonsignificant).

↓ Fecundity in F1.

↑ Malformations in F1

reproductive organs.

↓ F2 litter size.

Not reported. Delayed puberty in F0

males.

↓ Fertility in F0 males 

and females.

↑ Midterm abortion in F0

females.

↑ Testicular lesions in F0

males.

↓ Sperm production in 

F0 and F1 males.

↓ Fecundity in F1.

↑ Malformations in F1

reproductive organs.

↓ F2 litter size.
II-36

Appendix II

5.1.4.1 Utility of Data to the CERHR Evaluation

The data in rats are adequate for an assessment of reproductive toxicity as several studies are 

available that evaluate both structure and reproductive function. Transgenerational effects were 

examined in many of the studies. Animals were treated during gestational development, during 

lactation, and at weaning, thus ensuring that the most sensitive age for reproductive effects was 

assessed. The evaluation included androgen-regulated endpoints that are believed to be the most 

sensitive indicators of DBP effects. Reproductive organs were preserved in Bouin’s fixative, a 

method that reduces histological artifacts. Although studies in other species are not as detailed, they 

do allow for limited comparisons of interspecies sensitivity.

5.2 Integrated Evaluation

DBP is used as a coalescing aid in latex adhesives, as a plasticizer in cellulose plastics, and as a 

solvent for dyes. General human exposure occurs primarily through food. All estimates place total 

DBP exposure in the general population at less than 10 μg/kg bw/day. Although infants and young 

children may have higher exposures than adults, primarily because of different dietary patterns, 

estimates of their exposure remain at approximately 10 μg/kg bw/day, with the possible exception 

of non-dietary intake through mouthing of phthalate-containing objects. Workplace exposure at 

phthalate production facilities is estimated to be below 143 μg/kg bw/day. Exposure levels during 

incorporation of DBP in plastics are not known.

Following oral exposure to rodents and humans, DBP is quickly metabolized in the small intestine 

to mono-n-butyl phthalate, MbuP, and n-butyl alcohol. Several investigators have postulated that 

it is the monoester that is of toxicological interest. The Panel finds logic and data to support this 

view. Absorption of the monoester into blood occurs in both rats and humans. Although data for 

DBP is not available for humans or primates, it is reasonable to assume that MBuP would be 

rapidly glucuronidated and excreted in the urine in a manner analogous to DEHP in humans. The 

toxicokinetic data indicate that no tissue bioaccumulation would be expected via the oral or dermal 

route.

There are no data on the developmental or reproductive toxicity of DBP in humans. There 

are data in rats and mice that show oral exposure to DBP causes developmental toxicity. The 

developing male reproductive system is most sensitive to the formation of structural and functional 

abnormalities, with effects seen in rats whose mothers were exposed to 100 mg/kg bw/day during 

pregnancy. The NOAEL for male reproductive system developmental effects in rats is 50 mg/

kg bw/day. Breeding studies provide a good indication of the potential for adverse functional 

reproductive effects from DBP exposure. Moreover, it is apparent that DBP testicular exposure 

late in gestation can induce Leydig cell hyperplasia and a low incidence of Leydig cell adenoma. 

Traditional teratogenicity protocols that evaluate fetuses just prior to birth were not effective in 

detecting these effects on the developing male reproductive system. While a series of three recent 

studies have replicated and characterized the male reproductive system effects in rats, studies of 

similar design have not been performed in other species. The Panel is confident that these studies in 

rats correctly characterize the effects based on replication, good dose response, and full reporting 
Appendix II

II-37

of study results. As a default assumption, these data in rats are assumed relevant to a prediction of 

hazard to humans. 

The Expert Panel notes that the male reproductive system is a sensitive target organ for effects 

in rodent studies where exposure is confined to the adult phase of life. Data in several species, 

including rat, mouse, and guinea pig, show such effects. The Panel also notes that studies in the 

hamster, although limited, do not show effects on the testes. 

There are indications that oral exposure of females during the adult phase of life impairs functional 

reproductive performance in rats at doses of 250 mg/kg bw/day and higher. There is also a report 

that exposure to similar doses during gestation and nursing may impair fertility in female offspring. 

However, the data are not of the scope and quality for the Expert Panel to confidently characterize 

these effects.

Data indicate that the monoester of DBP, MBuP, is the principal toxicant. Studies suggest that an 

antiandrogenic mechanism appears to be responsible for the most sensitive endpoints observed in 

developing males rats (e.g., anogenital distance, nipple retention, preputial separation). It is not 

currently known whether the target for DBP is similar or different for gestational versus postnatal 

exposures.

The Panel is aware of studies performed at CDC using urine from human subjects. Results of these 

studies were given in an oral presentation in Copenhagen, Denmark, in May, 2000. MBuP values in 

the urine of women of child-bearing age were among the higher values. Such data, when published, 

should serve to improve our ability to assess phthalate exposure in the general populations.

5.3 Expert Panel Conclusions

DBP is used as a coalescing aid in latex adhesives, as a plasticizer in cellulose plastics, and as 

a solvent for dyes. The best estimate for exposures from all sources to the general population is 

2–10 μg/kg bw/day. There is significant uncertainty in the exposure database based on the age 

of many of the values/studies. The Expert Panel has high confidence in the available studies to 

characterize reproductive and developmental toxicity based upon a strong database containing 

studies in multiple species using conventional and investigative study designs. When administered 

via the oral route, DBP elicits malformations of the male rat reproductive tract via a disturbance 

of the androgen status: a mode of action relevant for human reproductive development. This 

antiandrogenic mechanism occurs via effects on testosterone biosynthesis and not androgen 

receptor antagonism. DBP is a testicular toxicant in three species of young adult laboratory 

animals in high dose (>1,000 mg/kg bw/day), sub-acute oral exposure studies. In the rat, there is 

a life-stage sensitivity for testicular toxicity with the fetus most sensitive, pubertal less sensitive, 

and adult least sensitive. Adult female functional reproductive toxicity (decreases in fertility) 

has been noted in rats; however, the data do not permit confident characterization of dose-effects 

below 250 mg/kg bw/day. The Expert Panel has negligible concern for adult reproductive toxicity.

DBP is developmentally toxic to both rats and mice by the oral routes; it induces structural 
II-38

Appendix II

malformations. A confident NOAEL of 50 mg/kg bw/day by the oral route has been established in 

the rat. Data from which to confidently establish a LOAEL/NOAEL in the mouse are uncertain. 

The Expert Panel has minimal concern about effects to human development and development 

of the reproductive system from current estimated exposure to DBP. A modified dietary 

multigeneration study is available, but did not establish a NOAEL. The LOAEL (M:52 F:80 mg/

kg bw/day) is based on decreases in litter size and pup weight.

5.4 Critical Data Needs 

The multigeneration study for DBP in rodents, with support from other studies that incorporated 

more modern endpoints including developmental landmarks, indicates no immediate data gaps. The 

potential effects of DBP on female rats warrant further investigation. 

Although there are no critical data needs, studies in the following areas would increase 

understanding about reproductive and developmental effects that occur following DBP exposures.

There is a need to extend the current PBPK model for DBP to include parameters for pregnant 

women and their fetuses.

There is a need to find out how broad or narrow the window of prenatal exposure is that results in 

postnatal male effects. The known current window in rats, 12−20 days, is still quite wide from a 

rodent ontogenesis perspective. Greater precision as to size of the window of sensitivity may be 

relevant to estimating the temporal bounds of human sensitivity.

Much of the recent focus on reproductive toxicity of phthalate esters has focused on the ability 

of certain esters to induce effects on reproductive development. Significant primate data exist 

to support the view that the high blood levels of monoester necessary to achieve adult testicular 

toxicity in rodents will not occur in humans. Appropriate exposure to monoester in blood from 

diester exposure could be achieved such that experiments could be conducted in primates to 

elucidate species sensitivity for equivalent exposures. This would require exposure of pregnant 

animals during the critical window of development of the reproductive system for the species 

studied, followed by an examination of reproductive development in the resulting offspring. Such a 

study would indicate if there is a species sensitivity. In the absence of such a study, the rodent data 

must be considered relevant and critical for human risk examinations
Appendix II

II-39

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benzyl phthalate and di-n-butyl phthalate in rats. Arch Environ Contam Toxicol 28:223-

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38. Wine R, Li L-H, Barnes LH, Gulati DK, Chapin RE. Reproductive toxicity of di-n-butyl 

phthalate in a continuous breeding protocol in Sprague-Dawley rats. Environ Health 

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39. Lamb JC, IV. Reproductive effects of four phthalic acid esters in the mouse. Toxicol Appl 

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40. Reel JR, Lawton AD, Feldman DB, Lamb JC. Di(N-Butyl) Phthalate: Reproduction and 

fertility assessment in CD-1 mice when administered in the feed. NTP-84-411: National 

Toxicology Program, National Institute of Environmental Health Sciences, 1984.

41. Gray LE, Jr, Wolf C, Lambright C, Mann P, Price M, Cooper RL, Ostby J. Administration 

of potentially antiandrogenic pesticides (procymidone, linuron, iprodione, chlozolinate, 

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PCB 169, and ethane dimethane sulphonate) during sexual differentiation produces diverse 

profiles of reproductive malformations in the male rat. Toxicol Ind Health 15:94-118(1999).

42. Ema M, Kurosaka R, Amano H, Ogawa Y. Developmental toxicity evaluation of mono-nbutyl phthalate in rats. Toxicol Lett 78:101-106(1995).

43. Ema M, Kurosaka R, Harazono A, Amano H, Ogawa Y. Phase specificity of developmental 

toxicity after oral administration of mono-n-butyl phthalate in rats. Arch Environ Contam 

Toxicol 31:170-176(1996).

44. Tyl RW, Price CJ, Marr MC, Kimmel CA. Developmental toxicity evaluation of dietary 

di(2-ethylhexyl)phthalate in Fischer 344 rats and CD-1 mice. Fundam Appl Toxicol 10:395-

412(1988).

45. Price CJ, Tyl RW, Marr MC, Myers CB, Sadler BM, Kimmel CA. Reproduction and fertility 

evaluation of diethylhexyl phthalate (CAS No. 117-81-7) in CD-1 mice exposed during 
II-42

Appendix II

gestation. Research Triangle Park, NC: National Toxicology Program, 1988.

46. Price CJ, Tyl RW, Marr MC, Sadler BM, Kimmel CA. Reproduction and fertility evaluation 

of diethylhexyl phthalate (CAS No. 117-81-7) in Fischer 344 rats exposed during gestation 

NTP 86-309. Research Triangle Park, NC: National Toxicology Program, 1986.

47. Mylchreest E, Cattley RC, Foster PM. Male reproductive tract malformations in rats 

following gestational and lactational exposure to di(n-butyl) phthalate: An antiandrogenic 

mechanism? Toxicol Sci 43:47-60(1998).

48. Chapin R, Sloane R, Haseman J. Reproductive endpoints in general toxicity studies: are 

they predictive? Reprod Toxicol Jul-Aug; 12(4):489-94(1998).

49. Mylchreest E, Sar M, Cattley RC, Foster PMD. Disruption of androgen-regulated male 

reproductive development by di(n-butyl) phthalate during late gestation in rats is different 

from flutamide. Toxicol Appl Pharmacol 156:81-95(1999).

50. Mylchreest E, Wallace DG, Cattley RC, Foster P. Dose-dependent alterations in androgenregulated male reproductive development in rats exposed to di(n-butyl)phthalate during late 

gestation. Toxicol Sci(2000).

51. Higuchi TT, Kane CM, Palmer JS, Veeramachaneni DNR. Developmental effects of 

digbutyl phthalate in frogs and rabbits. Abstract 193 at the Society of the Study of 

Reproduction meeting, 1999. (1999).

52. Imajima T, Shono T, Zakaria O, Suita S. Prenatal phthalate causes cryptorchidism 

postnatally by inducing transabdominal ascent of the testis in fetal rats. J Pediatr Surg 32:

18-21(1997).

53. Murature DA, Tang SY, Steinhardt G, Dougherty RC. Phthalate esters and semen quality 

parameters. Biomed Environ Mass Spectrom 14:473-477(1987).

54. Cater BR, Cook MW, Gangolli SD, Grasso P. Studies on dibutyl phthalate-induced 

testicular atrophy in the rat: Effect on zinc metabolism. Toxicol Appl Pharmacol 41:609-

618(1977).

55. Gray T, Rowland I, Foster P, Gangolli S. Species differences in the testicular toxicity of 

phthalate esters. Toxicol Lett 11:141-147(1982).

56. Maloney EK, Waxman DJ. Trans-activation of PPARα and PPARγ by structurally diverse 

environmental chemicals. Toxicol Appl Pharmacol 161:209-218(1999).

57. Gray TJ, Butterworth KR. Testicular atrophy produced by phthalate esters. Arch Toxicol 4:

452-455(1980).

58. Creasy DM, Beech LM, Gray TJB, Butler WH. The ultrastructural effects of di-n-pentyl 

phthalate on the testis of the mature rat. Exp Mol Pathol 46(1987).

59. Creasy DM, Foster JR, Foster PMD. The morphological development of di-n-pentyl 

phthalate induced testicular atrophy in the rat. J Pathol 139:309-321(1993).

60. Lloyd SC, Foster PM. Effect of mono- (2-ethylhexyl) phthalate on follicle-stimulating 
Appendix II

II-43

hormone responsiveness of cultured rat Sertoli cells. Toxicol Appl Pharmacol 95:484-

489(1988).

61. Heindel JJ, Chapin RE. Inhibition of FSH-stimulated cAMP accumulation by mono(2-

ethylhexyl) phthalate in primary rat Sertoli cell cultures. Toxicol Appl Pharmacol 97:377-

385(1989).

62. Sjoberg P, Bondesson U, Kjellen L, Linquist NG, Montin G, Ploen L. Kinetics of di(2-

ethylhexyl) phthalate in immature and mature rats and effect on testis. Acta Pharmacol 

Toxicol 56:30-37(1985).

63. Heindel JJ, Powell CJ. Phthalate ester effects on rat Sertoli cell function in vitro: Effects of 

phthalate side chain and age of animal. Toxicol Appl Pharmacol 115:116-123(1992).

64. Zacharewski TR, Meek MD, Clemons JH, Wu ZF, Fielden MR, Matthews JB. Examination 

of the in vitro and in vivo estrogenic activities of eight commercial phthalate esters. Toxicol 

Sci 46:282-293(1998).

65. Jobling S, Reynolds T, White R, Parker MG, Sumpter JP. A variety of environmentally 

persistent chemicals, including some phthalate plasticizers, are weakly estrogenic. Environ 

Health Perspect 103:582-587(1995).

66. Bolger R, Wiese TE, Ervin K, Nestich S, Checovich W. Rapid screening of environmental 

chemicals for estrogen receptor binding capacity. Environ Health Perspect 106:551-7(1998).

67. Nakai M, Tabira Y, Asa D, Yakabe Y, Shimyozu T, Noguchi M, Takatsuki M, Shimohigashi 

Y. Binding characteristics of dialkyl phthalates for the estrogen receptor. Biochemical and 

Biophysical Research Communications 254:311-314(1999).

68. Petit F, Le Goff P, Cravedi J-P, Valotaire Y, Pakdel F. Two complementary bioassays for 

screening the estrogenic potency of xenobiotics: Recombinant yeast for trout estrogen 

receptor and troup hepatocyte cultures. Journal of Molecular Endocrinology 19:321-

335(1997).

69. Coldham NG, Dave M, Sivapathasundaram S, McDonnell DP, Connor C, Sauer MJ. 

Evaluation of a recombinant yeast cell estrogen screening assay. Environ Health Perspect 

105:734-742(1997).

70. Harris CA, Henttu P, Parker MG, Sumpter JP. The estrogenic activity of phthalate esters in 

vitro. Environ Health Perspect 1997 105:802-811(1997).

71. Mulligan SR, Balasubramanian AV, Kalita JC. Relative potency of xenobiotic estrogens in 

an acute in vivo mammalian assay. Environ Health Perspect 106:23-26(1998).

72. Foster P. Personal communication with J. Moore, 2000.

73. Gray TJ, Beamand JA. Effects of some pthalate esters and other testicular toxins on primary 

cultures of testicular cells. Food Chem Toxicol 22:123-131(1984).
II-44

Appendix II

7.0 TABLES

7.1 DBP General Toxicity, Rats ...................................................................................................45

7.2 DBP General Toxicity, Rats ...................................................................................................46

7.3 DBP General Toxicity, Rats ...................................................................................................47

7.4 DBP General Toxicity, Mice ..................................................................................................48

7.5 DBP Developmental Toxicity, Rats........................................................................................49

7.6 DBP Developmental Toxicity, Rats........................................................................................50

7.7 DBP Developmental Toxicity, Rats........................................................................................51

7.8 DBP Developmental Toxicity, Rats........................................................................................52

7.9 DBP Developmental Toxicity, Mice.......................................................................................53

7.10 DBP Developmental Toxicity, Rats........................................................................................54

7.11 DBP Developmental Toxicity, Rats........................................................................................55

7.12 DBP Developmental Toxicity, Rats........................................................................................56

7.13 DBP Developmental Toxicity, Rats........................................................................................57

7.14 DBP Reproductive Toxicity, Rats...........................................................................................58

7.15 DBP Reproductive Toxicity, Rats...........................................................................................59

7.16 DBP Reproductive Toxicity, Mice .........................................................................................60

7.17 DBP MBuP Developmental Toxicity, Rats ............................................................................61

7.16 References..............................................................................................................................62
Appendix II

II-45

Table 7.1: DBP General Toxicity, Rats

Species, 

Strain, and 

Source

Experimental Regimen

Animal 

Number

/Sex

Dose* Body Weight Organ/Body Weight Ratio Histopathology Hematology Chemistry Other

Wistar RatsBASF 1992(1)

3-month sub-chronic study. Forty-two day-old rats of both sexes were exposed to DBP in the diet at concentrations of 0, 400, 2,000, or 10,000 ppm and then killed and necropsied. 26 tissues collected, histopathology of control and high dose liver, kidney, and testes examined at all doses. Hematology, clinical chemistry, urinalysis at mid- and end of study. Neurobehavior assessed 3x during study. 

10 010 27(M)/

33(F)

NE NE NE NE NE

10 141(M)/

162(F)

NE NE NE NE NE NOAEL

10 688(M)/

816(F)

NE ↑Li, Ki(F) ↓Lipid in 

hepatocytes.No testicular effects.

Transient ↓RBC, Hb, Hct (M).

↑Glu, Alb (M).↓Trigl, T3.

↑ PCoANo neurological effects.

*Dose in mg/kg bw/day.NA=Not Analyzed M = Male Trigl = Triglycerides Glu = GlucoseNE=No Effect F = Female Alb = Albumin↑= Statistically Significant Increase Li = Liver RBC = Red Blood Cell↓=Statistically Significant Decrease Ki = Kidney Hb = HemoglobinPCoA=Palmitoyl-CoA Oxidase T3 = Triiodothyronine Hct = Hematocrit
II-46

Appendix II

Table 7.2 : DBP General Toxicity, Rats

Species, 

Strain, and 

Source

Experimental 

Regimen

Animal/

Number/

Sex

Dose*

Body 

Weight

Gain

Organ/Body 

Weight Ratio Histopathology Hematology Chemistry Other

F344/N RatsMarsman 1995(2)

Sub-chronic study (13 weeks), 5−6-week-old rats were fed DBP and then killed and necropsied. Lowest dose was 2,500 ppm, doses then doubled until highest dose of 40,000 ppm achieved. Extensive tissue exam, hematology, clinical chemistry, semen, peroxisome proliferation enzyme evaluation at term.

10 010 176(M)/

177(F)

NE NE NE NE ↑Alb (M) NOAEL

10 359(M)/

356(F)

↓(M) ↑Li and Ki (M) NE ↓Hb (M)

↓RBC (M)↑Pl (M)

↑Alb (M)↓Trigl (M)↑Bile Ac (F)

↑PCoA

10 720(M)/

712(F)

↓(M) ↑Li and Ki Hepatic lesions.

Testicular lesions.

↓Hb, Hct (M)↓RBC (M)↑MCV↑Pl (M)

↑Alb (M)↓Trigl (M)↑Bile Ac (F)↑AP (F)

↑PCoA

10 1,540(M)/

1,413(F)

↓ ↑Li and Ki

↓Te

Hepatic lesions.Testicular lesions, marked hypospermia.↓ Sperm motility and concentration.

↓Hb, Hct (M)↓RBC (M)↑MCV↑Pl (M)

↑Alb (M)↓Chol, Trigl↑Bile Ac↑AP↓TP (F )

↑PCoA ↓Testic. Zn↓Testost.

10 2,964(M)/

2,943(F)

↓a ↑Li and Ki

↓Te

Hepatic lesions and peroxisomal proliferation.Testicular lesions and hypospermia.

↓Hb, Hct (M)↓ RBC (M)↑ MCV (M)↑Pl (M)

↑Alb(M)↓TP↓Chol, Trigl↑Bile Ac↑AP

↑PCoA↓Testic. Zn & serum Zn↓Testost.

*Dose in mg/kg bw/day.aFood consumption only 58% (M) and 83% (F) of control.

M= Male Te =Testes Trigl = TriglyceridesZn = Zinc

NE= No Effect F= Female Tp = Total Protein AP = Alkaline Phosphatase Pl = Platelets↑= Statistically Significant Increase Li = Liver Alb = Albumin Bile Ac = Bile Acids Testost = Testosterone↓=Statistically Significant Decrease Ki = Kidney Chol = Cholesterol PCoA = Palmitoyl-CoA Oxidase HCT= HematocritHb = Hemoglobin RBC= Red Blood Cell Count MCV= Mixed Cell Volume
Appendix II

II-47

Table 7.3: DBP General Toxicity, Rats 

Species, 

Strain, and 

Source

Experimental Regimen

Animal 

Number

/Sex

Dose*

Body 

Weight

Gain

Organ/Body 

Weight Ratio Histopathology Hematology Chemistry Other

F344/N RatsMarsman 1995(2)

Rats were exposed to 0 or 10,000 ppm DBP during prenatal development until 8 weeks of age. At 8 weeks of age, the rats were then fed DBP in the diet for 13 weeks, killed and necropsied.

10 0**10 0 ↑a ↑Tea NE NE ↓Test a ↑PCoA at 

weaning.

10 138(M)/

147(F)

↑Ki(F)b, Li(F)a↑Tea

NE NE

10 279(M)/

294(F)

↑Ki(F)b(M)a↑Li (F)a(M)a,b↑Tea

NE NE ↑Alb(F)a ↑PCoA(M) a,b

No-effect level for liver and testes.

10 571(M)/

593(F)

↓Fb, Mab ↑Ki(F)b(M)a,b

↑Li a,b↑Tea

Hepatic and testicular lesions.

↓ Hct ↓Hb ↓RBC(M)b↑Pl(M)b

↑Alba,b↓Trigl(M)a,b

↑PCoA a,b

10 1,262(M)/

1,182(F)

↓ab ↑Kia,b

↑Lia,b↓Tea,b

Hepatic and testicular lesions.↓Sperm counts and hypospermia of epididymis.

NE ↑Alba,b

↓ Chola,b↓Trigl a,b↑ AP a,b

↑Zn in serum(M)a,b↑PCoA a,b

10 2,495(M)/

2,445 (F)

↓a,b,c ↑Kia,b

↑Lia,bTea,b

Hepatic lesions, peroxisomal proliferation, and testicular lesions.↓Sperm counts and hypospermia of epididymis.

↓Hct ↓Hb ↓RBC a,b↑Pl(M)a,b

↓Tot Prota,b↑Alb(M)a,b↓ Chola,b↓Trigla,b↑ APa,b↑Bile Ac (F)b, (M)a,b↓Test a

↑Zn in serum(M)b↓Testicular Zna,b↑PCoA a,b

*Dose in mg/kg bw/day. ** No prenatal exposure aSignificant compared to control with no perinatal DBP exposurebSignificant compared to control with 10,000 ppm DBP perinatal exposure c Significant reduction in food consumption, rats emaciatedNA=Not Analyzed M = Male Te = Testes Trigl = Triglycerides Zn = Zinc Hb = HemoglobNE=No Effect F = Female Tot Prot = Total Protein AP = Alkaline Phosphatase Pl = Plateets RBC = Red Blood cells↑= Statistically Significant Increase Li= Liver Alb = Albumin Bile Ac = Bile Acids Test = Testosterone↓=Statistically Significant Decrease Ki = Kidney Chol = Cholesterol Hct = Hematocrit PCoA = Palmitoyl-CoA Oxidase
II-48

Appendix II

Table 7.4: DBP General Toxicity, Mice

Species, 

Strain, and 

Source

Experimental 

Regimen

Animal 

Number/

Sex

Dose* Body Weight Organ/Body Weight Ratio Histopathology Hematology Chemistry

B6C3F1MiceMarsman 1995(2)

13 week sub-chronic study. 6 week-old mice were exposed to DBP in the diet at levels of 0,1,250, 2,500, 5,000, 10,000, or 20,000 ppm for 13 weeks and then killed and necropsied. Organ weights, histological exam of tissues. Hematology, sperm morphology and vaginal cytology.

10 010 163(M)/ 

238(F)

NE ↑Ki(F) NE NE NA

10 353(M)/

486(F)

NE ↑Ki(F) NE NE NA

10 812(M)/

971(F)

↓ ↑Li

↑Ki(F)

NE NE NA

10 1,601(M)/

2,137(F)

↓ ↑Li

↑Ki(F)

Liver lesions (M). NE NA

10 3,689(M)/

4,278(F)

↓ ↑Li

↑Ki(F)

Liver lesions.No testicular lesions or other adverse reproductive effects

↓ Hct (F) NA

*Dose in mg/kg bw/day.NA=Not Analyzed M=Male Zn=ZincNE=No Effect F=Female Hct=Hematocrit↑= Statistically Significant Increase Li=Liver↓=Statistically Significant Decrease Ki=Kidney
Appendix II

II-49

Table 7.5: DBP Developmental Toxicity, Rats

Species, 

Strain, and 

Source

Experimental Regimen Numbera Dose* Maternal Effects Fetal Effects

ICR-JCL MiceShiota et al. 1980; Shiota and Mishimura 1982(3, 4)

Prenatal developmental toxicity study. Mice were fed diets with 0, 0.05, 0.1, 0.2, 0.4, or 1% DBP from gd 0−18. Body weights were measured on gd 0−18. Dams were sacrificed on gd 18. Corpora lutea were counted and pups were examined for skeletal and soft tissue malformations.

8 07 80 NE LOAELb

Delayed ossification.

8 180 NE Delayed ossification.6 350 NE Delayed ossification.9 660 NOAEL ↓ Fetal weight.

Delayed ossification.

15 2,100 ↓ Bodyweight gain. ↑ Resorptions (98.4 vs 5%).

↓ Fetal weight.Delayed ossification.↑ Neural tube defects (2/3 fetuses).c

*Dose in mg/kg bw/day.aNumber of pregnant females at sacrifice. bDiffers from author’s selection of effect level. cEffect not statistically significant.↑=Statistically Significant Increase↓=Statistically Significant DecreaseNE=No effects
II-50

Appendix II

Table 7.6: DBP Developmental Toxicity, Rats 

Species, Strain, 

and Source Experimental Regimen Numbera Dose* Maternal Effects Fetal Effects

Wistar RatsEma et al. 1993(5)

Prenatal developmental toxicity study.Rats were gavaged with DBP from gd 7−15. Body weights and food intake were measured daily. Dams were sacrificed on gd 20. Implantation sites were examined. Pups were sexed, weighed, and evaluated for external malformations. Two-thirds of fetuses were examined for skeletal malformations and 1/3 for visceral malformations.

11(11) 011(11) 500 NOAEL NOAEL

12(12) 630 ↓Weight gain. Complete resorption in 2/12 litters.

↓Live fetuses/litter (43%).↓ Fetal weight (9−10%).

12(12) 750 ↓Adjusted weight gain 

(38%).

Complete resorption in 10/12 litters.↓ Live fetuses/litter (93%).↓ Fetal weight (14−18%).↑ External malformations (cleft palate) in 6/10 fetuses (2 litters) vs 0/118 fetuses in control.

11(9) 1,000 ↓Adjusted weight gain

(71%).

Complete resorption in 9/9 litters.

*Dose in mg/kg bw/day.aNumber of pregnant rats (Number of litters evaluated)↑=Statistically Significant Increase↓=Statistically Significant Decrease
Appendix II

II-51

Table 7.7: DBP Developmental Toxicity, Rats

Species, 

Strain, and 

Source

Experimental Regimen Numbera Dose* Maternal Effect Fetal Effects

Wistar RatsEma et al. 1998(6)

Prenatal developmental toxicity study.Rats were fed diets with 0, 0.5, 1.0, or 2.0% DBP on gd 11−21. Body weights and food intake were measured. Dams were sacrificed on gd 21. Implantation sites were examined. Pups were sexed, weighed, and evaluated for external malformations. Two-thirds of fetuses were examined for skeletal malformations and 1/3 for visceral malformations.

11 011 331 NOAEL NOAEL.11 555 ↓Corrected weight gain.b

↓Food intake.

↓Anogenital distance in males.↑Undescended testes (15% vs 0 in 7/11 litters).

11 661 ↓Corrected weight gain.b

↓Food intake.

↓Fetal weight (22%).↓Anogenital distance in males.↑Undescended testes (53% vs 0 in 11/11 litters).↑External (cleft palate; 4% vs 0 in 4/11 litters) and skeletal (fused sternebrae; 55% vs 0 in 11/11 litters) malformations

*Dose in mg/kg bw/day.aNumber of pregnant rats and litters evaluated bBody weight excluding gravid uterus↑=Statistically Significant Increase↓=Statistically Significant Decrease
II-52

Appendix II

Table 7.8: DBP Developmental Toxicity, Rats

Species, 

Strain, 

Source

Experimental Regimen Animal Number Dose* Maternal Effects F1 Offspring Effects

F344/N RatMarsman et al. 1995(2)

Pre- and post-natal exposure study.DBP administered in feed to dams throughout gestation and lactation. Dams were weighed on gd 0 and 18, and weekly during lactation. Uteri of nulliparous rats in high-dose group were stained with ammonium sulfide. Gestation indexa, litter size, and pup survival were examined. Pups were weighed at birth and pnd 0, 4, and weekly thereafter.

28b 0

15 92 (1,250 ppm)d NE NE15 184 (2,500 ppm) NOAEL ↓ Weight days 21−28.13 368 (5,000 ppm) ↓ Gestation index(68 vs. 93%)a

↓ Gestation length.

↓ Weight days 1−28.

14 551 (7,500 ppm) NE ↓ Weight days 0-28.16 736 (10,000 ppm) ↓ Weight gain during lactation. ↓ Weight days 0-28.

↓ Percent live pups/litter (89 vs. 96%).

14 1,472 (20,000 ppm) ↓ Gestation index (21 vs. 93%)a

↓ Gestational weight gain.

↓ Pup weight day 0.↓ Litter size (72%) and % live pups/litter (29 vs 99)Complete pup mortality by pnd 1.

After weaning on day 28, pups were administered DBP in feed for 4 weeks at the same levels administered to their mothers (1,250, 2,500, 5,000, 7,500, 10,000 ppm).Body weights were measured weekly.Necropsies were conducted and organ weights determined for all groups. Histopathology was evaluated in control and high dose rats. Testis evaluated in dose groups receiving 2,500 ppm and higher.

10c 0

10 133(F)/143(M)e ↑ Kidney (M) & liver to body weight ratio (M).

↑ Weight gain in females.

10 275(F)/284(M) ↑ Kidney (M) to body weight ratio (M).

↑ Liver to body weight ratio.

10 500(F)/579(M) Hypospermia in 4/10 males.

↑ Kidney & liver to body weight ratio .

10 836(F)/879(M) Hypospermia in 10/10 males.

↓ Weight gain in males.↑ Kidney & liver to body weight ratios.

10 1,104(F)-1165(M) Hypospermia in 10/10 males.

↓Testis to body weight ratio (11%).↓ Weight gain in males.↑ Kidney & liver to body weight ratios .

*Dose in mg/kg bw/day.aDelivery of ≥ 1 live pup per sperm positive female bNumber of rats delivering litters cNumber of pups/sexdDoses estimated by CERHR, see Section 3 for explanation. eAuthor calculated doses for females and males, respectively NE=No effect↑=Statistically Significant Increase ↓=Statistically Significant Decrease
Appendix II

II-53

Table 7.9: DBP Developmental Toxicity, Mice

Species, 

Strain, 

Source

Experimental Regimen Animal Number Dose* Maternal effects Offspring effects

B6C3F1MouseMarsman et al. 1995(2)

Pre- and post-natal exposure study. DBP administered in feed to dams throughout gestation and lactation. Dams were weighed on gd 0 and 17, and weekly during lactation. Uteri of nulliparous mice in highdose group were stained with ammonium sulfide. Litter size, and pup survival were examined. Pups were weighed at birth and pd 0, 4, and weekly thereafter.

11a 0

10 227 (1,250 ppm)c NE NE12 454 (2,500 ppm) ↑ Gestation length (2%). ↓ Litter size.

9 908 (5,000 ppm) ↑ Gestation length (3%). NE

11 1,359 (7,500 ppm) ↓ Gestational weight gain (18%).

↑ Gestation length (5%).

↓ Litter size (28%). ↓ Live pups/litter (48%).

5 1,816 (10,000 ppm) ↓ Gestational weight gain (34%).

↑ Gestation length (6%).

↓ Litter size (48%). ↓ Live pups/litter (89%).↓ Pup birth weight (14%).

0 3,632 (20,000 ppm) No live deliveries.

After weaning, pups were administered DBP in feed for 4 weeks at the same levels administered to their mothers (0, 1,250, 2,500, 5,000, 7,500, 10,000 ppm). Body weights were measured weekly. Necropsies were conducted and organ weights determined for all groups. Histopathology was evaluated in controls and the 1,060−1,286 mg/kg bw/day group.

10b 0

10 170(F)/199(M)d ↑ Liver to body weight ratio (M).

↑ Kidney to body weight ratio (F).

10 399(F)/437(M) ↓Male body weights (7%).

↑ Liver to body weight ratio in males.↑ Kidney to body weight ratio in females.

10 714(F)/750(M) ↓Male body weights (11%).

↑ Liver to body weight ratio in males.↑ Kidney to body weight ratio in females.

10 1,060(F) /1,286(M) ↓Male body weights (12%).

↓ Female body weight (11%).↑ Liver to body weight ratio in males.↑ Kidney to body weight ratio in females.

1 3,804(M)

*Dose in mg/kg bw/day.aNumber of mice delivering litters bNumber of pups/sex cDoses estimated by CERHR, see section 3 for explanation.dAuthor calculated doses for females and males respectively ↑=Statistically Significant IncreaseNE=No Effect (M)=Male (F)=Female
II-54

Appendix II

Table 7.10: DBP Developmental Toxicity, Rats

Species, 

Strain, and 

Source

Experimental Regimen Numbera Dose* Maternal effects Offspring effects

CD RatMylchreest et al. 1998(7)

Pre- and post-natal developmental toxicity study. Rats were gavaged with DBP from gd 3 until the end of lactation.Body weights were measured daily and food intake was measured weekly. Dams were killed and necropsied following weaning of pups. Implantation sites were examined.Pups were sexed, weighed, and evaluated for sexual maturation. Pups were sacrificed on pnd 100−105. All males and up to 3 females/litter were necropsied. Histological exams were conducted on malformed rats and ≤ 2 normal rats/litter. Sperm analysis was conducted at necropsy.

9 08 250 NE ↑Hypospadia (1/32 pups), underdeveloped or absent 

epididymis (3/32 pups; 2 litters) and seminal vesicles (0 pups), and undescended testes (1/32 pups).

7 500 ↓Uterine weight. ↓Anogenital distance in males (pnd 1).

↑Hypospadia (7/34 pups; 4 litters), underdeveloped or absent epididymis (17/34 pups; 6 litters) and seminal vesicles (2/34 pups; 2 litters), and undescended testes (2/34 pups; 2 litters).↓Testis (24%) and seminal vesicle weight (16%).

4 750 ↓Uterine weight 

(non-significant).

↓Live pups/litter (27%).↓Pup survival during lactation (85 vs 96%).↓Anogenital distance in males (pnd 1).↑Hypospadia (6/14 pups; 2 litters), underdeveloped or absent epididymis (10/14 pups; 3 litters) and seminal vesicles (7/14 pups; 3 litters), and undescended testes (4/14 pups; 2 litters).↓Testis (33%), seminal vesicle (32%), epididymis (34%), and prostate weight (27%) .↓Kidney weight.No effects on female sexual development or estrous cycles.

*Dose in mg/kg bw/day.aThe number of dams that litteredlitters evaluated. ↑ Statistically Significant Increase ↓ Statistically Significant Decrease
Appendix II

II-55

Table 7.11: DBP Developmental Toxicity, Rats

Species, 

Strain, 

Source

Experimental Regimen Number* Dose* Maternal effects Offspring effects

CD RatMylchreest et al. 1999(8)

Pre- and post-natal developmental toxicity study. Rats were gavaged with DBP from gd 12−21. Body weights were measured daily during dosing and weekly at other times. Food intake was measured weekly. Dams were killed and necropsied following weaning of pups. Implantation sites were examined. Pups were sexed, weighed, and evaluated for sexual maturation. Male pups were sacrificed on pnd 100–105 and a histological examination of sex organs was conducted. Females were sacrificed on pnd 25–30 and their reproductive tracts were evaluated for gross abnormalities.

10 0

 9 100 NE ↑Age of preputial separation (5%).

10 250 NE ↓Anogenital distance in males (9%).

↑Retained thoracic nipples (35/62 pups; 5 litters).↑Absent or underdeveloped epididymis (6/62 pups; 4 litters).

 9 500 Large weight loss 

(16%) and complete litter death in one dam.

↓Anogenital distance in males (24%).↑ Retained thoracic nipples (47/54 pups; 8 litters).↑Age of preputial separation (9%).↑Hypospadias (21/52 pups; 4 litters), absent prostate (3/52 pups; 1 litter), absent or underdeveloped epididymis (26/52 pups; 8 litters) and vas deferens (14/52 pups; 4 litters).↑Testicular and epididymal lesions.↑Interstitial adenoma (2/45 in 1 litter versus 0/51 pups in control).↑Intra-abdominal testes (5/52 pups; 3 litters).↓Absolute testes (16%), epididymis (26%), and seminal vesical (21%) weight.↓Absolute kidney weight.

Results were compared to those induced by the antiandrogenic drug, flutamide.

 5 100 flutamide ↓Body weight gain. ↓Anogenital distance in males.

↑ Retained thoracic nipples.↑Hypospadias, underdeveloped or absent seminal vesicles, complete lack of prostate and epididymis, and vas deferens development.↑Testicular lesions. ↑Suprainguinal testes.↓Absolute testes, epididymis, and seminal vesical weight.

*Dose in mg/kg bw/day.*Numbers of litters evaluated NE=No Effect ↑=Statistically Significant Increase ↓=Statistically Significant Decrease
II-56

Appendix II

Table 7.12: DBP Developmental Toxicity, Rats

Species, 

Strain, and 

Source

Experimental Regimen Numbera Dose* Maternal effects Offspring effects

SpragueDawley RatMylchreest et al. 2000(9)

Pre- and post-natal developmental toxicity study.Rats were gavaged with DBP in corn oil from gd 12−21. Dams delivered litters and pups were examined and weighed at birth. After the pups were weaned, dams were killed and organ weights and implantation sites were evaluated.Pups were weighed weekly and evaluated for sexual maturation until killed at puberty. Male and female pup organs were weighed and testes and epididymides were examined histologically.

19 020 0.5 No effects observed 

at any dose level.

NE

19 5 NE20 50 NOAEL20 100 ↑Seminiferous tubule degeneration 

(3% of rats in 2/10 litters).↑Retained areolas or nipples in males (31% of rats in 16/20 litters).

11 500 ↑Seminiferous tubule degeneration 

(56% of rats in 3/5 litters).↑Retained areolas or nipples in males (90% of rats in 11/11 litters).↓Anogenital distance in males.↑Hypospadias (9% of rats in 4/11 litters).↑Agenesis of epididymis (36% of rats in 9/11 litters), vas deferens (28% of rats in 9/11 litters), and prostate (1/58 rats).↓Testis, epididymis, prostate, and levator ani muscle weight.↑Interstitial cell hyperplasia (35% of rats in 3/5 litters) and adenoma (1/23 rats).↑Intra-abdominal testes (4 rats/3 litters).No effect on vaginal opening or on female reproductive organ weight or histology.

*Dose in mg/kg bw/day.aNumber of litters evaluated. ↑=Statistically Significant Increase ↓=Statistically Significant Decrease NE=No Effect
Appendix II

II-57

Table 7.13: DBP Developmental Toxicity, Rats

Species, 

Strain, and 

Source

Experimental Regimen Numbera Dose* Maternal effects Offspring effects

Sprague Dawley Rat

Pre- and post-natal developmental toxicity study.DBP administered in oil by gavage from gd 14 to lactation day 3. Male pups were examined for sexual maturation. At 5 months of age, male offspring were killed and necropsied. Organ weights were measured and a histological examination was conducted on reproductive organs.

9 08 500 Not reported ↓Anogenital distance 

(2.79 vs 3.70mm).↑Percentage of areolas (55 vs 0%) and numbers of areolas/nipples at birth (n=2.7 vs 0) and adulthood (2.2 vs 0).

↑% Hypospadias (6.2 vs 0%) and 

testicular and epididymal atrophy or agenesis (46 vs 0%).↓Seminal vesicle, prostate, epididymis, testes, levator ani, and penis weight.

LE Hooded RatsGray et al. 1999(10)

LE Hooded Rats were gavaged with DBP from gd 16−19. All other details are as described above for longer exposure in Sprague-Dawley rats.

6 04 500 Not reported ↓Anogenital distance 

(2.83 vs 3.21 mm).↑Percentage of areolas (87 vs 0%) and numbers of areolas/nipples at birth and adulthood (1.9 vs 0).↓Seminal vesicle, prostate, and levator ani muscle weight.

*Dose in mg/kg bw/day.aNumber of pregnant rats. ↑ Statistically Significant Increase ↓ Statistically Significant Decrease
II-58

Appendix II

Table 7.14: DBP Reproductive Toxicity, Rats

Species, 

Strain, and 

Source

Experimental Regimen Numberb Dosec (mg/kg bw/day) Effects

CD RatsWine et al. 1997(11)a

Fertility assessment through a continuous breeding study.DBP administered in feed at 1,000, 5,000 or 10,000 ppm.Breeding pairs housed together for 112 days; female body weight was measured on days of littering and both sexes at necropsy; clinical signs, and food intake were recorded; litters were counted, sexed, weighed, and removed following birth. In a crossover breeding study, high-dose F0 males and females were mated with control animals for 1 week. At the end of the study Necropsy and a histopathological examination were conducted.

40 020 52(M)/80(F) ↓ Live pups/litter.19 256(M)/385(F) ↓ Live pups/litter.

↓ Pup weight.

20 509(M)/794(F) ↓ Live pups/litter.

↓ Pup weight.↑ Liver and kidney to body weight ratio.↓ Pup weight from treated females in crossover.

Final F1 litters from continuous breeding study were weaned and mated within dose groups for 1 week. Rats continued to receive the same DBP concentrations as their parents.

20 020 52(M)/80(F) ↓ F2 Pup weight.20 256(M)/385(F) ↑ Kidney to body weight ratio (M).

↓ F2 Pup weight.↑ Degeneration of seminiferous tubules.

20 509(M)/794(F) 30% mating, 5% pregnancy, 17% fertility indices 

↓ Sperm count (49%).↑ Degeneration of seminiferous tubules, interstitial cell hyperplasia, underdeveloped epididymis, and malformed penises and prepuces.↓ Prostate and seminal vesicle to body weight ratio.↓ Testis weight.↓ Body weight in males and females.↑ Liver and kidney to body weight ratio inmales.↓ F2 Pup weight

aThis study is also addressed in Marsman et al. 1995 bNumber of male and female pairs cAuthor-calculated male and female doses, respectively. ↑=Statistically Significant Increase (F)=Female↓=Statistically Significant Decrease (M)=Male
Appendix II

II-59

Table 7.15: DBP Reproductive Toxicity, Rats

Species, 

Strain, 

and 

Source

Experimental Regimen

Animal

Number Dose* Effects

LE Hooded RatsGray et al. 1999(10)

Multigeneration reproductive study. Male and female rats (F0) were gavaged with DBP from puberty through adulthood, mating, and lactation. Sexual maturation and estrous cycles were evaluated. Treated rats were mated with untreated controls. Following weaning of F1pups, F0 rats were killed. At necropsy, serum hormone levels, organ weights, testicular histology, and implantation sites were examined.

24a 0

10 250 ↑Age of F0 preputial separation (42.6 vs 39.5 days).

↑Malformed F1 pups (14.5 vs 0.7%) and litters with malformed pups (50 vs 5.5%).

4 500 ↑Age of F0 preputial separation (43.4 vs 39.5 days).

↑Malformed F1 pups (33 vs 0.7%) and litters with malformed pups (100 vs 5.5%).↓Fertility in F0 males and females.↑Testicular atrophy in F0 males.↓Sperm production in F0 males.↑Midterm abortions in F0 females.

8−12b

(males 

only)

1,000 ↑Age of F0 preputial separation (44.4 vs 39.5 days).

↑Testicular atrophy in F0 males.↓Sperm production in F0 males.

The F1 rats were not exposed to DBP following weaning.Some F1 pups from treated dams were mated within dose groups for 11 cycles, and the remainder were necropsied. F2 pups were counted and discarded.

18c 0d

18 250 ↓Fecundity in F1.

↓Number of F2 pups born.↓Caudal sperm levels in F1 (non-significant; 19%).

4 500 ↓Fecundity in F1.

↓Number of F2 pups born.↓Caudal sperm levels in F1 (34%).

*Dose in mg/kg bw/day.aNumber of litters evaluated bNumber of males, only males treated with highest dose cNumber of breeding pairs dMaternal (F0) exposure levels ↑=Statistically Significant Increase ↓=Statistically Significant Decrease
II-60

Appendix II

Table 7.16: DBP Reproductive Toxicity, Mice

Species, 

Strain, and 

Source

Experimental Regimen Numberb Dosec (mg/kg bw/day) Effects

CD-1 MiceReel et al. 1984; Lamb 1987(12, 13)a

Fertility assessment through a continuous breeding study. DBP administered in feed. Breeding pairs housed together for 98 days; body weight was measured on 7 days, clinical signs, and food intake were recorded; litters were counted, sexed, weighed, and removed following birth. In a crossover breeding study, high dose males and females were mated with control mice. Breeding pairs were housed together for 7 days or until a copulatory plug was observed. Necropsy and a histopathological examination were conducted.

39 020 52.5 NE18 525 NOAEL20 1,750 ↓ Number of fertile pairs.

↓ Number of litters delivered/pair.↓ Litter size.↓ Live pups.↑ Percentage of male pups.↓ Pup weight.↓ Uterus to body weight ratio in F0 females.↓ Body weight in F0 males.↑ Liver to body weight ratio in F0 males and females.No effects on estrous cycles, sperm morphology, or sex organs in F0mice.

aThis study is also addressed in Marsman et al. 1995bNumber of male and female pairscAuthor-calculated dosesNE=No Effect↑=Statistically Significant Increase↓=Statistically Significant Decrease
Appendix II

II-61

 Table 7.17: MBuP Developmental Toxicity, Rats

Species, 

Strain, and 

Source

Experimental Regimen Dose* Maternal effects Offspring effects

Wistar-King A rats.Imajima et al. 1997(14)

Pre and postnatal developmental toxicity study with prenatal exposure.Rats were gavage dosed with 0 or 300 mg/day MBuP in sesame oil from gd 15-18.Testicular descent was evaluated in male offspring on gd 20 or pnd 30-40.

19 / 15 015 / 26 1,000 Not reported. ↑ Testicular ascent on gd 20.

↑ Chryptorchidism in 22/26 male pups on pnd 30–40 with 87% of the undescended testes in abdominal cavity and 13% in the inguinal ring.

*Dose in mg/kg bw/day.aNumber of male fetuses evaluated on gd 20 / pnd 30-40.↑ Statistically significant increase
II-62

Appendix II

References:

1. BASF. Study on the oral toxicity of dibutyl phthalate in Wistar rats. Administration via the 

diet over 3 months. 31S0449//89020: Eastman Kodak Company, 1992.

2. Marsman DS. NTP technical report on toxicity studies of dibutyl phthalate (CAS No. 84-74-

2) administered in feed to F344 rats and B6C3F1 mice NIH Publication 95-3353. Research 

Triangle Park: National Toxicology Program, 1995.

3. Shiota K, Chou MJ, Nishimura H. Embryotoxic effects of di-2-ethylhexyl phthalate (DEHP) 

and di-n-butyl phthalate (DBP) in mice. Environ Res 22:245-253(1980).

4. Shiota K, Nishimura H. Teratogenicity of di (2-ethylhexyl) phthalate (DEHP) and di-n-butyl 

phthalate (DBP) in mice. Environ Health Perspect 45:65-70(1982).

5. Ema M, Amano H, Itami T, Kawasaki H. Teratogenic evaluation of di-n-butyl phthalate in 

rats. Toxicol Lett 69:197-203(1993).

6. Ema M, Miyawaki E, Kawashima K. Further evaluation of developmental toxicity of din-butyl phthalate following administration during late pregnancy in rats. Toxicol Lett:87-

93(1998).

7. Mylchreest E, Cattley RC, Foster PM. Male reproductive tract malformations in rats 

following gestational and lactational exposure to di(n-butyl) phthalate: An antiandrogenic 

mechanism? Toxicol Sci 43:47-60(1998).

8. Mylchreest E, Sar M, Cattley RC, Foster PMD. Disruption of androgen-regulated male 

reproductive development by di(n-butyl) phthalate during late gestation in rats is different 

from flutamide. Toxicol Appl Pharmacol 156:81-95(1999).

9. Mylchreest E, Wallace DG, Cattley RC, Foster P. Dose-dependent alterations in androgenregulated male reproductive development in rats exposed to di(n-butyl)phthalate during late 

gestation. Toxicol Sci(2000).

10. Gray LE, Jr, Wolf C, Lambright C, Mann P, Price M, Cooper RL, Ostby J. Administration 

of potentially antiandrogenic pesticides (procymidone, linuron, iprodione, chlozolinate, 

p,p’-DDE, and ketoconazole) and toxic substances (dibutyl- and diethylhexyl phthalate, 

PCB 169, and ethane dimethane sulphonate) during sexual differentiation produces diverse 

profiles of reproductive malformations in the male rat. Toxicol Ind Health 15:94-118(1999).

11. Wine R, Li L-H, Barnes LH, Gulati DK, Chapin RE. Reproductive toxicity of di-n-butyl 

phthalate in a continuous breeding protocol in Sprague-Dawley rats. Environ Health 

Perspect 105:102-107(1997).

12. Reel JR, Lawton AD, Feldman DB, Lamb JC. Di(N-Butyl) Phthalate: Reproduction and 

fertility assessment in CD-1 mice when administered in the feed. NTP-84-411: National 

Toxicology Program, National Institute of Environmental Health Sciences, 1984.

13. Lamb JC, IV. Reproductive effects of four phthalic acid esters in the mouse. Toxicol Appl 

Pharmacol 88:255-269(1987).

14. Imajima T, Shono T, Zakaria O, Suita S. Prenatal phthalate causes cryptorchidism 

postnatally by inducing transabdominal ascent of the testis in fetal rats. J Pediatr Surg 32:

18-21(1997).
Appendix III

NTP-CERHR EXPERT PANEL REPORT 

on

Di-n-Butyl Phthalate

October 2000 NTP-CERHR-DBP-00

Center For The Evaluation Of Risks 

To Human Reproduction

National Toxicology Program

U.S. Department of Health and Human Services

PUBLIC COMMENTS

ON THE PHTHALATES

EXPERT PANEL REPORTS
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