# NTP-CERHR EXPERT PANEL REPORT

on Di- n -Butyl Phthalate

NTP-CERHR-DBP-00 

# October 2000 

# 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-89i The National Toxicology Program (NTP) es

tablished 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 repro

ductive and/or developmental health effects associated with exposure to environmental and/or occupational chemicals. The CERHR

is located at the National Institute of Envi

ronmental 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 includ

ing 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 tox

icity.

The CERHR convenes a scientific expert panel that meets in a public forum to review, discuss,

and evaluate the scientific literature on the se

lected chemical. Public comment is invited pri

or to and during the meeting. The expert panel produces a report on the chemical’s reproduc

tive 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 evalu

ate 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 in

cludes the NTP brief on the chemical evaluated, the expert panel report, and all public com

ments. The goal of the NTP brief is to provide the public, as well as government health, regu

latory, 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 mono

graph 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 contact

ing 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 contact

ing the NTP Office of Liaison and Scientific Re

view at the NIEHS: liaison@starbase.niehs.nih.gov [email] 919-541-0530 [phone]i i iii In 1999, the CERHR Core Committee, an ad

visory 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 occur

rence within the environment, (b) they have a high production volume, (c) there is substantial scientific literature addressing the reproductive and/or de

velopmental toxicities of these chemi

cals, 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 flex

ibility 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 sur

rounding environment. DEHP has the greatest production volume of the selected phthalates (approximately 260 million pounds [1994]), followed by DIDP (approximately 240 mil

lion pounds [1994]), and DINP (approximately 215 million pounds [1994]). The scientific lit

erature 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 repro

ductive 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 mono

graph 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’sreports 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 pres

ents the NTP’s interpretation of the potential for exposure to DBP to cause adverse repro

ductive 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, scien

tifically sound information on the potential for DBP exposures to result in adverse health ef

fects on development and reproduction. 

### Introductioni i iii While there are biological and practical rea

sons for considering developmental toxicity and reproductive toxicity as 2 separate is

sues, 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 de

velopment of the offspring, birth, post-natal development, sexual maturity, and, again, production of sperm and eggs.

In the past, toxic effects were often stud

ied 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. Re

productive toxicity was studied by exposing sexually mature adults to the chemical of in

terest and effects were detected as impaired capacity to reproduce. Over the years, toxi

cologists realized that exposure during one part of the life cycle could lead to adverse effects that might only be apparent at a dif

ferent part of the life cycle. For example, exposure of a sexually mature individual to an agent capable of inducing genetic dam

age 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 con

trast, the reproductive system begins devel

oping 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 pu

berty or sexual maturity.

Thus, in the case of genetic damage induced in eggs or sperm, what might be considered reproductive toxicity gives rise to develop

mental 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 re

productive toxicity. This issue is important in considering the phthalate evaluations because evidence of developmental toxic

ity 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**1NTP Brief **What is DBP?** DBP is a clear, oily liquid with the chemical

formula C 16 H 22 O 4 and the structure shown in Figure 1. It is one of a group of industri

ally important chemicals known as phthalates. Phthalates are used primarily as plasticizers to add flexibility to plastics. Unlike many phthal

ates, 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 plas

tics, and as a solvent for dyes.

DBP is produced by reacting n-butanol with phthalic anhydride. The most recent informa

tion 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 ex

posed to DBP. There are several ways that peo

ple 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 ad

hesives, 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-con

taining materials that are used to process and package food. Possible, sources of DBP found in food are latex adhesives used in food processingequipment 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. gener

al population is exposed to approximately 2-10 μg/kg bw/day (micrograms per kilogram body weight per day). This reflects a total daily expo

sure 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 DBP2NTP Brief

3NTP 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 appar

ently 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. Informa

tion 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 af

fects 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 con

clude that the results reported in laboratory animals indicate a potential for similar or other

adverse effects in humans. Scientific decisions concerning health risks aregenerally 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 con

clude 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 repro

ductive 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 labora

tory animals that rodents in prenatal and early postnatal stages of development are more sen

sitive 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 animals2NTP Brief

3NTP 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 ad

verse effects of DBP exposure. There is grow

ing 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 reproduc

tive 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 (An

derson 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 hu

man 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 un

1 Based on Kohn et al. (2000) estimated exposure of some women of reproductive

age to ~100 μg/kg bw/day (median 1.7; 95 th 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 adults 2 

### Developmental effects 2 

### 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 exposures 14NTP Brief

5NTP 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 imme

diate 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 Phthal

ates Expert Panel that there is minimal con

cern 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 ad

verse effects to human development, particu

larly development of the male reproductive
system.

This level of concern is greater than that ex

pressed 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 need

ed to determine if the higher DBP exposure levels reported for some women of reproduc

tive 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.4NTP Brief

5NTP Brief 

### References:

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. Ef

fects 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-1Appendix 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-2Appendix 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

NTP-CERHR-DBP-00 

# October 2000 

# Center For The Evaluation Of Risks **To Human Reproduction** National Toxicology Program U.S. Department of Health and Human ServicesAppendix 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

............................................................... ............................................................... ........................ 44Appendix IIAppendix 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 spe

cific chemicals or classes of chemicals, including descriptions of any uncertainties that would dimin

ish 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: 703-838-9440II-iiAppendix 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 EditorAppendix 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 C 16 H 22 O 4 Molecular Weight 278.35

Vapor Pressure 2.7 x 10 -5 mmHg at 25 o C

Melting Point -35 o C

Boiling Point 340 o C

Specific Gravity 1.042 Solubility in Water Slight: 11.2 mg/L

Log K

ow 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 en vironment can be deposited on or taken up by crops intended for consumption by humans or livestock and can thus enter the food supply.II-2Appendix IIAppendix II

II-3 **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

μ g/m 3 0.420 μ g/m 3 <1.0 μ g/L Various levels in a Canadian market basket survey. (See text)

Assumptions 22 m 3 inhaled/day; 64 kg bw; 4/24

hours outdoors 22 m 3 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

μ g/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 μ g/m 3 0.2 μ g/L 78

− 200 μ g/kg

Assumed Intake Rate 20 m 3 /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 ( μ g/kg bw/day) 0.0005

− 0.0009 0.005 0.007

− 0.02

(7)

Table 4: Health Canada DBP Exposure Estimates

ESTIMATED INTAKE DBP ( μ g/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)II-2Appendix IIAppendix 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 dif ferent 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 μ g/kg bw/day by IPCS (3) and at 1.9 μ g/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 μ g/kg bw/day in children aged 12

− 19 years to 5.0 μ g/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 μ g DBP/person/day and 31 μ g 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 μ g/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 μ g/kg bw/day at birth

and 1.4 μ g/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

μ g/kg bw/day was based on DBP levels of 78–200 μ g/kg that were reported for fish in studiesII-4Appendix IIAppendix II

II-5 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 μ g/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 μ g/kg bw/day in children aged 12–19 years to 0.11 μ g/kg bw/day in infants aged 0–6 months. Adult DBP exposure through drinking water was estimated by ATSDR (7) at 0.005 μ g/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 μ g/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 μ g/kg bw/day and exposures in children ranged from 0.68 μ g/kg bw/day in 0–6 month-old infants to 1.1 μ g/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 b y inhalation of vapors and dust. Phthalates are manufactured within closed systems, but workers can be exposed duringII-4Appendix IIAppendix 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/m 3 ) to 0.08 mg/m 3 (3). OSHA established a permissible exposure limit of 5 mg/m 3 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/m 3 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 m 3 /day inhalation rate and a 70 kg body weight. The resulting exposure estimate was 143 μ g/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) ( μ g/kg bw/day) Exposure in Adults ( μ g/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-6Appendix IIAppendix II

II-7 

### 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 LD 50 for DBP appears to be between 8,000 and 20,000 mg/kg bw in rats (3) and the 90-day

dermal LD 50 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.II-6Appendix IIAppendix 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-8Appendix IIAppendix II

II-9 chronic study.

The NTP also conducted a sub-chronic study in 6-week-old B6C3F 1 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 4,278mg/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/m 3 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/cm 2 / 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 ofII-8Appendix IIAppendix 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 radioacti vity 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-10Appendix IIAppendix II

II-11 Pregnant Rodents. Saillenfait et al. (21) studied metabolism and placental transfer of 14 C-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 14 C 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 refl ect 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 follo wing 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 effi cient 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).II-10Appendix IIAppendix 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 w eight 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-12Appendix IIAppendix II

II-13 

### 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/da y 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.II-12Appendix IIAppendix 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 14 C-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 F 0 dams. Developmental effects included a reduction in litter size in all dose groups and in live pup weight in the two highest doses

of F 1 rats. F 2 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-14Appendix IIAppendix II

II-15 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 F 1 rats and reductions in F 2 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 B6C3F 1 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 F 1 pups/group were fed a diet with a DBPII-14Appendix IIAppendix 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 F 1 pups died by pnd 1 in the 20,000 ppm group. Male pup bodyweights 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 B6C3F 1 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 F 1 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 F 1 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-16Appendix IIAppendix II

II-17 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 F 1 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) gavaged19–20 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 deliveredII-16Appendix IIAppendix 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 F 1 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 F 1 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-18Appendix IIAppendix II

II-19 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 of fspring. 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-exposedII-18Appendix IIAppendix 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 putati ve mediator in trans-abdominal migration of the testis.

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

II-21 

### 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 B6C3F 1 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/da y 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 , numberII-20Appendix IIAppendix 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 F 0 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 F 0 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 F 0 animals. Systemic effects in the F 0 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 F 1 litters following the continuous F 0 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 F 1 males and females within the same treatment group were housed in pairs for 1 week and then housed individually until delivery of an F 2 litter.

F 1 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 F 1 group with only 1 of 20 pairings resulting in a litter. While litter size was unaffected,

F 2 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 F 1 males from the high-dose group. There were no effects on the ovary

of F 1 females. Epididymal sperm count and testicular spermatid count was significantly reduced in the high-dose F 1 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 F 1 females and there was no effect on ante-mortem estrous cyclicity.

In Wine et al. (38), the F 1 high-dose group had a high rate of infertility, the middle dose had fewer (F 0 mating) and lighter pups (F 0 and F 1 matings), while the low-dose animals had fewer pups (F 0 mating) and lighter pups (F 1 mating). Thus, a NOAEL was not established. The LOAEL was 52–80 mg/kg bw/day based on reductions in F 0 litter size and F 2 pup weight. The Expert Panel’s confidence in the quality of the study is high, and our confi dence is also high that these dosesII-22Appendix IIAppendix II

II-23 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 (F 0 ) 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 F 0 were evaluated.

Treated rats were mated with untreated controls. When the F 1 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 F 0 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 F 0 males and females in the 500 mg/kg bw/day group. Infertility in F 0 males was apparently due to testicular atrophy and reduced sperm counts. F 0 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 F 1 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 F 1 pups were not treated with DBP after weaning. Four to eighteen pairs of F 1 pups from treated dams were selected for continuous mating within dose groups for 11 cycles and the remaining F 1 pups were necropsied. The F 2 pups born during the continuous breeding phase were counted and discarded. Fecundity was reduced in F 1 rats from treated dams and the number

of F 2 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 F 1 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 F 0II-22Appendix IIAppendix 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 uterineweight 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 14 C-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 signifi cantly 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 areII-24Appendix IIAppendix II

II-25 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.**II-24Appendix IIAppendix 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 μ g/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 μ g/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 refl ect 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-26Appendix IIAppendix II

II-27 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 LD 50 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 B6C3F 1 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.II-26Appendix IIAppendix 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 14 C-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 e valuation 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 ofII-28Appendix IIAppendix II

II-29 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 ossificationII-28Appendix IIAppendix 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 B6C3F 1 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-30Appendix IIAppendix II

II-31 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 authorsII-30Appendix IIAppendix 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 insuffi cient 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 ef fects. 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 F 1 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 confi dence.

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 F 0 generation there was clear indication that DBP, when administered in the diet, affected total number of pups per litter in all treated groups. The F 1 high-dose group had malformations of the reproductive tract and a high rateII-32Appendix IIAppendix II

II-33 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 F 0 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 F 1 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 signifi cant 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 treatedwith 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 unaf fected. 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 14 Clabelled 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, theII-32Appendix IIAppendix 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 ef fects due to gestational exposure is MBuP. This pattern of effects induced in rodents by late gestational exposure (gdII-34Appendix IIAppendix II

II-35 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).II-34Appendix IIAppendix 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 secondgeneration 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

↓ F 1 live litter size. ↓ F 2 pup weight. M: 509; F: 794

↓ Body weight gain in

F 0 females and

F 1 males and

females. ↑ Liver and kidney weight

in F 0 males and

females and F 1 males. ↑ Malformed reproductive organs in F 1 males. ↓ Mating, pregnancy, and fertility in F 1 . ↓ Reproductive organ weights in F 1 males. ↑ Testicular lesions in F 1 males. ↓ Sperm counts in F 1 . ↓ F 1 litter size. ↑ F 1 pup mortality. ↓ F 1 and F 2 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 F 0 females. ↓ Uterine weight in

F 0 . ↓ Live pups/litter.

No effects on sperm

in F 0 . 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.

F 1 rats were not treated following weaning.

(41) Reproductive:

None

Systemic: Not reported 250

Delayed puberty in F 0 males. ↓ Sperm production

in F 1 males (nonsignificant). ↓ Fecundity in F 1 . ↑ Malformations in F 1 reproductive organs. ↓ F 2 litter size. Not reported. Delayed puberty in F 0 males. ↓ Fertility in F 0 males

and females. ↑ Midterm abortion in F 0 females. ↑ Testicular lesions in F 0 males. ↓ Sperm production in

F 0 and F 1 males. ↓ Fecundity in F 1 . ↑ Malformations in F 1 reproductive organs. ↓ F 2 litter size.II-36Appendix IIAppendix II

II-37 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 ef fects 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 dif ferent 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 ef fects 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 reportingII-36Appendix IIAppendix 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 tar get 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 of fspring. 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 structuralII-38Appendix IIAppendix II

II-39 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 of fspring. 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 examinationsII-38Appendix IIAppendix II

II-39 

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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-44Appendix IIAppendix II

II-45 

### 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

............................................................... ............................................................... 62II-44Appendix IIAppendix 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-chronicstudy. Forty-two day-oldrats of both sexes wereexposed to DBP in thediet at concentrations of0, 400, 2,000, or 10,000ppm and then killed andnecropsied. 26 tissuescollected, histopathology ofcontrol and high dose liver,kidney, and testes examinedat all doses. Hematology,clinical chemistry, urinalysisat mid- and end of study.Neurobehavior assessed 3xduring study.1001027(M)/33(F)NENENENENE10141(M)/162(F)NENENENENENOAEL10688(M)/816(F)NE↑Li, Ki(F)↓Lipid inhepatocytes.No testiculareffects.Transient↓RBC, Hb,Hct (M).↑Glu, Alb(M).↓Trigl, T3.↑PCoANoneurologicaleffects.*Dose in mg/kg bw/day.NA=Not AnalyzedM = MaleTrigl = TriglyceridesGlu = GlucoseNE=No EffectF = FemaleAlb = Albumin↑= Statistically Significant IncreaseLi = LiverRBC = Red Blood Cell↓=Statistically Significant DecreaseKi = KidneyHb = HemoglobinPCoA=Palmitoyl-CoA OxidaseT3 = TriiodothyronineHct = HematocritII-46Appendix IIAppendix II

II-47Table 7.2 : DBP General Toxicity, RatsSpecies,Strain, andSourceExperimentalRegimenAnimal/Number/SexDose*BodyWeightGainOrgan/BodyWeight RatioHistopathologyHematologyChemistryOtherF344/N RatsMarsman 1995(2)Sub-chronic study(13 weeks), 5−6-week-old ratswere fed DBPand then killedand necropsied.Lowest dose was2,500 ppm, dosesthen doubleduntil highestdose of 40,000ppm achieved.Extensive tissueexam, hematology,clinical chemistry,semen, peroxisomeproliferationenzyme evaluationat term.10010176(M)/177(F)NENENENE↑Alb (M)NOAEL10359(M)/356(F)↓(M)↑Li and Ki (M)NE↓Hb (M)↓RBC (M)↑Pl (M)↑Alb (M)↓Trigl (M)↑Bile Ac (F)↑PCoA10720(M)/712(F)↓(M)↑Li and KiHepatic lesions.Testicular lesions.↓Hb, Hct (M)↓RBC (M)↑MCV↑Pl (M)↑Alb (M)↓Trigl (M)↑Bile Ac (F)↑AP (F)↑PCoA101,540(M)/1,413(F)↓↑Li and Ki↓TeHepatic lesions.Testicular lesions,marked hypospermia.↓Sperm motility andconcentration.↓Hb, Hct (M)↓RBC (M)↑MCV↑Pl (M)↑Alb (M)↓Chol, Trigl↑Bile Ac↑AP↓TP (F )↑PCoA↓Testic. Zn↓Testost.102,964(M)/2,943(F)↓a↑Li and Ki↓TeHepatic lesionsand peroxisomalproliferation.Testicular lesions andhypospermia.↓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.aFoodconsumption only 58% (M) and 83% (F) of control.M= MaleTe =TestesTrigl = TriglyceridesZn = ZincNE= No EffectF= FemaleTp = Total ProteinAP = Alkaline PhosphatasePl = Platelets↑= Statistically Significant IncreaseLi = LiverAlb = AlbuminBile Ac = Bile AcidsTestost = Testosterone↓=Statistically Significant DecreaseKi = KidneyChol = CholesterolPCoA = Palmitoyl-CoA OxidaseHCT= HematocritHb = HemoglobinRBC= Red Blood Cell CountMCV= Mixed Cell VolumeII-46Appendix IIAppendix II

II-47Table 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 or10,000 ppm DBP duringprenatal development until8 weeks of age. At 8 weeksof age, the rats were then fedDBP in the diet for 13 weeks,killed and necropsied.100**100↑a↑TeaNENE↓Testa↑PCoA atweaning.10138(M)/147(F)↑Ki(F)b, Li(F)a↑TeaNENE10279(M)/294(F)↑Ki(F)b(M)a↑Li (F)a(M)a,b↑TeaNENE↑Alb(F)a↑PCoA(M)a,bNo-effect levelfor liver andtestes.10571(M)/593(F)↓Fb, Mab↑Ki(F)b(M)a,b↑Lia,b↑TeaHepatic andtesticular lesions.↓Hct↓Hb↓RBC(M)b↑Pl(M)b↑Alba,b↓Trigl(M)a,b↑PCoAa,b101,262(M)/1,182(F)↓ab↑Kia,b↑Lia,b↓Tea,bHepatic andtesticular lesions.↓Sperm counts andhypospermia ofepididymis.NE↑Alba,b↓Chola,b↓Trigla,b↑APa,b↑Zn inserum(M)a,b↑PCoAa,b102,495(M)/2,445 (F)↓a,b,c↑Kia,b↑Lia,bTea,bHepatic lesions,peroxisomalproliferation, andtesticular lesions.↓Sperm counts andhypospermia ofepididymis.↓Hct↓Hb↓RBCa,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↓Testa↑Zn inserum(M)b↓TesticularZna,b↑PCoAa,b*Dose in mg/kg bw/day.** No prenatal exposureaSignificant compared to control with no perinatal DBP exposurebSignificant compared to control with 10,000 ppm DBP perinatal exposurecSignificant reduction in food consumption, rats emaciatedNA=Not AnalyzedM = MaleTe = TestesTrigl = TriglyceridesZn = ZincHb = HemoglobNE=No EffectF = FemaleTot Prot = Total ProteinAP = Alkaline PhosphatasePl = PlateetsRBC = Red Blood cells↑= Statistically Significant IncreaseLi= LiverAlb = AlbuminBile Ac = Bile AcidsTest = Testosterone↓=Statistically Significant DecreaseKi = KidneyChol = CholesterolHct = HematocritPCoA = Palmitoyl-CoA OxidaseII-48Appendix IIAppendix II

II-49**Table 7.4: DBP General Toxicity, Mice**Species,Strain, andSourceExperimentalRegimenAnimalNumber/SexDose*BodyWeightOrgan/BodyWeightRatioHistopathologyHematologyChemistryB6C3F1MiceMarsman1995(2)13 week sub-chronicstudy. 6 week-old micewere exposed to DBPin the diet at levels of0,1,250, 2,500, 5,000,10,000, or 20,000 ppmfor 13 weeks and thenkilled and necropsied.Organ weights,histological exam oftissues. Hematology,sperm morphology andvaginal cytology.10010163(M)/238(F)NE↑Ki(F)NENENA10353(M)/486(F)NE↑Ki(F)NENENA10812(M)/971(F)↓↑Li↑Ki(F)NENENA101,601(M)/2,137(F)↓↑Li↑Ki(F)Liver lesions (M).NENA103,689(M)/4,278(F)↓↑Li↑Ki(F)Liver lesions.No testicular lesionsor other adversereproductive effects↓Hct (F)NA*Dose in mg/kg bw/day.NA=Not AnalyzedM=MaleZn=ZincNE=No EffectF=FemaleHct=Hematocrit↑= Statistically Significant IncreaseLi=Liver↓=Statistically Significant DecreaseKi=KidneyII-48Appendix IIAppendix II

II-49**Table 7.5: DBP Developmental Toxicity, Rats**Species,Strain, andSourceExperimental RegimenNumberaDose*Maternal EffectsFetal EffectsICR-JCL MiceShiota et al.1980; Shiotaand Mishimura1982(3, 4)Prenatal developmental toxicity study. Micewere fed diets with 0, 0.05, 0.1, 0.2, 0.4, or1% DBP from gd 0−18. Body weights weremeasured on gd 0−18. Dams were sacrificedon gd 18. Corpora lutea were counted andpups were examined for skeletal and softtissue malformations.80780NELOAELbDelayed ossification.8180NEDelayed ossification.6350NEDelayed ossification.9660NOAEL↓Fetal weight.Delayed ossification.152,100↓Bodyweight gain.↑Resorptions (98.4 vs 5%).↓Fetal weight.Delayed ossification.↑Neural tube defects (2/3fetuses).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 effectsII-50Appendix IIAppendix II

II-51Table 7.6: DBP Developmental Toxicity, RatsSpecies, Strain,and SourceExperimental RegimenNumberaDose*Maternal EffectsFetal EffectsWistar RatsEma et al. 1993(5)Prenatal developmental toxicity study.Rats were gavaged with DBP from gd7−15. Body weights and food intake weremeasured daily. Dams were sacrificed ongd 20. Implantation sites were examined.Pups were sexed, weighed, and evaluatedfor external malformations. Two-thirdsof fetuses were examined for skeletalmalformations and 1/3 for visceralmalformations.11(11)011(11)500NOAELNOAEL12(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) in6/10 fetuses (2 litters) vs 0/118 fetusesin 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 DecreaseII-50Appendix IIAppendix II

II-51**Table 7.7: DBP Developmental Toxicity, Rats**Species,Strain, andSourceExperimental RegimenNumberaDose*Maternal EffectFetal EffectsWistar 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 foodintake were measured. Dams were sacrificedon gd 21. Implantation sites were examined.Pups were sexed, weighed, and evaluated forexternal malformations. Two-thirds of fetuseswere examined for skeletal malformations and1/3 for visceral malformations.11011331NOAELNOAEL.11555↓Corrected weight gain.b↓Food intake.↓Anogenital distance in males.↑Undescended testes (15% vs 0 in 7/11litters).11661↓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 4/11litters) and skeletal (fused sternebrae;55% vs 0 in 11/11 litters) malformations*Dose in mg/kg bw/day.aNumber of pregnant rats and litters evaluatedbBody weight excluding gravid uterus↑=Statistically Significant Increase↓=Statistically Significant DecreaseII-52Appendix IIAppendix II

II-53**Table 7.8: DBP Developmental Toxicity, Rats**Species,Strain,SourceExperimental RegimenAnimalNumberDose*Maternal EffectsF1Offspring EffectsF344/N RatMarsman etal. 1995(2)Pre- and post-natal exposure study.DBP administered in feed to damsthroughout gestation and lactation.Dams were weighed on gd 0 and 18,and weekly during lactation.Uteri of nulliparous rats in high-dosegroup were stained with ammoniumsulfide. Gestation indexa, litter size,and pup survival were examined. Pupswere weighed at birth and pnd 0, 4, andweekly thereafter.28b01592 (1,250 ppm)dNENE15184 (2,500 ppm)NOAEL↓Weight days 21−28.13368 (5,000 ppm)↓Gestation index(68 vs. 93%)a↓Gestation length.↓Weight days 1−28.14551 (7,500 ppm)NE↓Weight days 0-28.16736 (10,000 ppm)↓Weight gain during lactation.↓Weight days 0-28.↓Percent live pups/litter (89 vs. 96%).141,472 (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 wereadministered DBP in feed for 4 weeksat the same levels administered to theirmothers (1,250, 2,500, 5,000, 7,500,10,000 ppm).Body weights were measured weekly.Necropsies were conducted and organweights determined for all groups.Histopathology was evaluated incontrol and high dose rats. Testisevaluated in dose groups receiving2,500 ppm and higher.10c010133(F)/143(M)e↑Kidney (M) & liver to body weight ratio (M).↑Weight gain in females.10275(F)/284(M)↑Kidney (M) to body weight ratio (M).↑Liver to body weight ratio.10500(F)/579(M)Hypospermia in 4/10 males.↑Kidney & liver to body weight ratio .10836(F)/879(M)Hypospermia in 10/10 males.↓Weight gain in males.↑Kidney & liver to body weight ratios.101,104(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 femalebNumber of rats delivering litterscNumber of pups/sexdDoses estimated by CERHR, see Section 3 for explanation.eAuthor calculated doses for females and males, respectivelyNE=No effect↑=Statistically Significant Increase↓=Statistically Significant DecreaseII-52Appendix IIAppendix II

II-53**Table 7.9: DBP Developmental Toxicity, Mice**Species,Strain,SourceExperimental RegimenAnimalNumberDose*Maternal effectsOffspring effectsB6C3F1MouseMarsman etal. 1995(2)Pre- and post-natal exposure study. DBPadministered in feed to dams throughoutgestation and lactation. Dams wereweighed on gd 0 and 17, and weekly duringlactation. Uteri of nulliparous mice in highdose group were stained with ammoniumsulfide. Litter size, and pup survival wereexamined. Pups were weighed at birth andpd 0, 4, and weekly thereafter.11a010227 (1,250 ppm)cNENE12454 (2,500 ppm)↑Gestation length (2%).↓Litter size.9908 (5,000 ppm)↑Gestation length (3%).NE111,359 (7,500 ppm)↓Gestational weight gain (18%).↑Gestation length (5%).↓Litter size (28%).↓Live pups/litter (48%).51,816 (10,000 ppm)↓Gestational weight gain (34%).↑Gestation length (6%).↓Litter size (48%).↓Live pups/litter (89%).↓Pup birth weight (14%).03,632 (20,000 ppm)No live deliveries.After weaning, pups were administeredDBP in feed for 4 weeks at the same levelsadministered to their mothers (0, 1,250,2,500, 5,000, 7,500, 10,000 ppm). Bodyweights were measured weekly. Necropsieswere conducted and organ weightsdetermined for all groups. Histopathologywas evaluated in controls and the1,060−1,286 mg/kg bw/day group.10b010170(F)/199(M)d↑Liver to body weight ratio (M).↑Kidney to body weight ratio (F).10399(F)/437(M)↓Male body weights (7%).↑Liver to body weight ratio in males.↑Kidney to body weight ratio in females.10714(F)/750(M)↓Male body weights (11%).↑Liver to body weight ratio in males.↑Kidney to body weight ratio in females.101,060(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.13,804(M)*Dose in mg/kg bw/day.aNumber of mice delivering littersbNumber of pups/sexcDoses estimated by CERHR, see section 3 for explanation.dAuthor calculated doses for females and males respectively↑=Statistically Significant IncreaseNE=No Effect(M)=Male(F)=FemaleII-54Appendix IIAppendix II

II-55**Table 7.10: DBP Developmental Toxicity, Rats**Species,Strain, andSourceExperimental RegimenNumberaDose*Maternal effectsOffspring effectsCD RatMylchreest etal. 1998(7)Pre- and post-natal developmental toxicitystudy.Rats were gavaged with DBP from gd 3until the end of lactation.Body weights were measured daily andfood intake was measured weekly. Damswere killed and necropsied followingweaning of pups. Implantation sites wereexamined.Pups were sexed, weighed, and evaluatedfor sexual maturation. Pups were sacrificedon pnd 100−105. All males and up to 3females/litter were necropsied. Histologicalexams were conducted on malformed ratsand≤2 normal rats/litter. Sperm analysiswas conducted at necropsy.908250NE↑Hypospadia (1/32 pups), underdeveloped or absentepididymis (3/32 pups; 2 litters) and seminal vesicles(0 pups), and undescended testes (1/32 pups).7500↓Uterine weight.↓Anogenital distance in males (pnd 1).↑Hypospadia (7/34 pups; 4 litters), underdevelopedor absent epididymis (17/34 pups; 6 litters)and seminal vesicles (2/34 pups; 2 litters), andundescended testes (2/34 pups; 2 litters).↓Testis (24%) and seminal vesicle weight (16%).4750↓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), underdevelopedor absent epididymis (10/14 pups; 3 litters)and seminal vesicles (7/14 pups; 3 litters), andundescended 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 estrouscycles.*Dose in mg/kg bw/day.aThe number ofdams that litteredlittersevaluated.↑Statistically Significant Increase↓Statistically Significant DecreaseII-54Appendix IIAppendix II

II-55**Table 7.11: DBP Developmental Toxicity, Rats**Species,Strain,SourceExperimental RegimenNumber*Dose*Maternal effectsOffspring effectsCD RatMylchreestet al. 1999(8)Pre- and post-natal developmentaltoxicity study. Rats were gavaged withDBP from gd 12−21. Body weightswere measured daily during dosing andweekly at other times. Food intake wasmeasured weekly. Dams were killedand necropsied following weaningof pups. Implantation sites wereexamined. Pups were sexed, weighed,and evaluated for sexual maturation.Male pups were sacrificed on pnd100–105 and a histological examinationof sex organs was conducted. Femaleswere sacrificed on pnd 25–30 and theirreproductive tracts were evaluated forgross abnormalities.1009100NE↑Age of preputial separation (5%).10250NE↓Anogenital distance in males (9%).↑Retained thoracic nipples (35/62 pups; 5 litters).↑Absent or underdeveloped epididymis (6/62 pups;4 litters).9500Large weight loss(16%) and completelitter 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 underdevelopedepididymis (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 0/51pups 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 thoseinduced by the antiandrogenic drug,flutamide.5100 flutamide↓Body weight gain.↓Anogenital distance in males.↑Retained thoracic nipples.↑Hypospadias, underdeveloped or absent seminalvesicles, complete lack of prostate and epididymis,and vas deferens development.↑Testicular lesions.↑Suprainguinal testes.↓Absolute testes, epididymis, and seminal vesicalweight.*Dose in mg/kg bw/day.*Numbers of litters evaluatedNE=No Effect↑=Statistically Significant Increase↓=Statistically Significant DecreaseII-56Appendix IIAppendix II

II-57**Table 7.12: DBP Developmental Toxicity, Rats**Species,Strain, andSourceExperimental RegimenNumberaDose*Maternal effectsOffspring effectsSpragueDawley RatMylchreestet al. 2000(9)Pre- and post-natal developmental toxicitystudy.Rats were gavaged with DBP in corn oil fromgd 12−21.Dams delivered litters and pups were examinedand weighed at birth.After the pups were weaned, dams were killedand organ weights and implantation sites wereevaluated.Pups were weighed weekly and evaluated forsexual maturation until killed at puberty.Male and female pup organs were weighedand testes and epididymides were examinedhistologically.190200.5No effects observedat any dose level.NE195NE2050NOAEL20100↑Seminiferous tubule degeneration(3% of rats in 2/10 litters).↑Retained areolas or nipples in males (31% ofrats in 16/20 litters).11500↑Seminiferous tubule degeneration(56% of rats in 3/5 litters).↑Retained areolas or nipples in males (90% ofrats in 11/11 litters).↓Anogenital distance in males.↑Hypospadias (9% of rats in 4/11 litters).↑Agenesis of epididymis (36% of rats in 9/11litters), vas deferens (28% of rats in 9/11litters), and prostate (1/58 rats).↓Testis, epididymis, prostate, and levator animuscle weight.↑Interstitial cell hyperplasia (35% of rats in 3/5litters) and adenoma (1/23 rats).↑Intra-abdominal testes (4 rats/3 litters).No effect on vaginal opening or on femalereproductive organ weight or histology.*Dose in mg/kg bw/day.aNumber of litters evaluated.↑=Statistically Significant Increase↓=Statistically Significant DecreaseNE=No EffectII-56Appendix IIAppendix II

II-57**Table 7.13: DBP Developmental Toxicity, Rats**Species,Strain, andSourceExperimental RegimenNumberaDose*Maternal effectsOffspring effectsSpragueDawley RatPre- and post-natal developmentaltoxicity study.DBP administered in oil by gavage fromgd 14 to lactation day 3.Male pups were examined for sexualmaturation.At 5 months of age, male offspring werekilled and necropsied. Organ weightswere measured and a histologicalexamination was conducted onreproductive organs.908500Not reported↓Anogenital distance(2.79 vs 3.70mm).↑Percentage of areolas (55 vs 0%) andnumbers of areolas/nipples at birth(n=2.7 vs 0) and adulthood (2.2 vs 0).↑% Hypospadias (6.2 vs 0%) andtesticular and epididymal atrophy oragenesis (46 vs 0%).↓Seminal vesicle, prostate, epididymis,testes, levator ani, and penis weight.LE HoodedRatsGray et al.1999(10)LE Hooded Rats were gavaged withDBP from gd 16−19.All other details are as described abovefor longer exposure in Sprague-Dawleyrats.604500Not reported↓Anogenital distance(2.83 vs 3.21 mm).↑Percentage of areolas (87 vs 0%) andnumbers of areolas/nipples at birth andadulthood (1.9 vs 0).↓Seminal vesicle, prostate, and levatorani muscle weight.*Dose in mg/kg bw/day.aNumber of pregnant rats.↑Statistically Significant Increase↓Statistically Significant DecreaseII-58Appendix IIAppendix II

II-59**Table 7.14: DBP Reproductive Toxicity, Rats**Species,Strain, andSourceExperimental RegimenNumberbDosec(mg/kg bw/day)EffectsCD RatsWine et al.1997(11)aFertility assessment through a continuous breedingstudy.DBP administered in feed at 1,000, 5,000 or 10,000ppm.Breeding pairs housed together for 112 days; femalebody weight was measured on days of litteringand both sexes at necropsy; clinical signs, andfood intake were recorded; litters were counted,sexed, weighed, and removed following birth. Ina crossover breeding study, high-dose F0malesand females were mated with control animals for1week. At the end of the study Necropsy and ahistopathological examination were conducted.4002052(M)/80(F)↓Live pups/litter.19256(M)/385(F)↓Live pups/litter.↓Pup weight.20509(M)/794(F)↓Live pups/litter.↓Pup weight.↑Liver and kidney to body weight ratio.↓Pup weight from treated females in crossover.Final F1litters from continuous breeding studywere weaned and mated within dose groups for1week. Rats continued to receive the same DBPconcentrations as their parents.2002052(M)/80(F)↓F2Pup weight.20256(M)/385(F)↑Kidney to body weight ratio (M).↓F2Pup weight.↑Degeneration of seminiferous tubules.20509(M)/794(F)30% mating, 5% pregnancy, 17% fertility indices↓Sperm count (49%).↑Degeneration of seminiferous tubules, interstitial cellhyperplasia, underdeveloped epididymis, and malformedpenises 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.↓F2Pup weightaThis study is also addressed in Marsman et al. 1995bNumber of male and female pairscAuthor-calculated male and female doses, respectively.↑=Statistically Significant Increase(F)=Female↓=Statistically Significant Decrease(M)=MaleII-58Appendix IIAppendix II

II-59**Table 7.15: DBP Reproductive Toxicity, Rats Species,** **Strain, and** **Source Experimental Regimen** **Animal Number** **Dose*****Effects**LE HoodedRatsGray et al.1999(10)Multigeneration reproductive study. Maleand female rats (F0) were gavaged with DBPfrom puberty through adulthood, mating, andlactation. Sexual maturation and estrous cycleswere evaluated. Treated rats were mated withuntreated controls. Following weaning of F1pups, F0rats were killed. At necropsy, serumhormone levels, organ weights, testicularhistology, and implantation sites wereexamined.24a010250↑Age of F0preputial separation (42.6 vs 39.5 days).↑Malformed F1pups (14.5 vs 0.7%) and litters with malformedpups (50 vs 5.5%).4500↑Age of F0preputial separation (43.4 vs 39.5 days).↑Malformed F1pups (33 vs 0.7%) and litters with malformedpups (100 vs 5.5%).↓Fertility in F0males and females.↑Testicular atrophy in F0males.↓Sperm production in F0males.↑Midterm abortions in F0females.8−12b(malesonly)1,000↑Age of F0preputial separation (44.4 vs 39.5 days).↑Testicular atrophy in F0males.↓Sperm production in F0males.The F1rats were not exposed to DBP followingweaning.Some F1pups from treated dams were matedwithin dose groups for 11 cycles, and theremainder were necropsied.F2pups were counted and discarded.18c0d18250↓Fecundity in F1.↓Number of F2pups born.↓Caudal sperm levels in F1(non-significant; 19%).4500↓Fecundity in F1.↓Number of F2pups born.↓Caudal sperm levels in F1(34%).*Dose in mg/kg bw/day.aNumber of litters evaluatedbNumber of males, only males treated with highest dosecNumber of breeding pairsdMaternal (F0) exposure levels↑=Statistically Significant Increase↓=Statistically Significant DecreaseII-60Appendix IIAppendix II

II-61**Table 7.16: DBP Reproductive Toxicity, Mice Species,** **Strain, and Source** **Experimental Regimen Number** **b Dose** **c (mg/kg bw/day)** **Effects**CD-1 MiceReel et al.1984; Lamb1987(12, 13)aFertility assessment through acontinuous breeding study. DBPadministered in feed. Breeding pairshoused together for 98 days; bodyweight was measured on 7days,clinical signs, and food intake wererecorded; litters were counted, sexed,weighed, and removed following birth.In a crossover breeding study, highdose males and females were matedwith control mice. Breeding pairswere housed together for 7days oruntil a copulatory plug was observed.Necropsy and a histopathologicalexamination were conducted.3902052.5NE18525NOAEL201,750↓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 F0females.↓Body weight in F0males.↑Liver to body weight ratio in F0males 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 DecreaseII-60Appendix IIAppendix II

II-61Table 7.17: MBuP Developmental Toxicity, RatsSpecies,Strain, andSourceExperimental RegimenDose*MaternaleffectsOffspring effectsWistar-KingA rats.Imajima etal. 1997(14)Pre and postnatal developmental toxicitystudy with prenatal exposure.Rats were gavage dosed with 0 or 300mg/day MBuP in sesame oil from gd 15-18.Testicular descent was evaluated in maleoffspring on gd 20 or pnd 30-40.19 / 15015 / 261,000Not reported.↑Testicular ascent on gd 20.↑Chryptorchidism in 22/26 malepups on pnd 30–40 with 87% of theundescended testes in abdominalcavity 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 increaseII-62Appendix II 

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# NTP-CERHR EXPERT PANEL REPORT

on Di- n -Butyl Phthalate

NTP-CERHR-DBP-00 

# October 2000 

# 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**