NTP-CERHR Monograph on the 

Potential Human Reproductive and 

Developmental Effects 

of Di-isononyl Phthalate (DINP)

March 2003 NIH Publication No. 03-4484

�����������������������������������

���������������������

���������������������������

��������������������������������������������
Table of Contents

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

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

NTP Brief on Di-isononyl Phthalate (DINP)...................................................................................1

References........................................................................................................................................4

Appendix I. NTP-CERHR Phthalates Expert Panel 

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

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

Appendix II. Phthalates Expert Panel Report on DINP

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

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

General Toxicological and Biological Parameters .........................................................II-7

Developmental Toxicity Data........................................................................................II-15

Reproductive Toxicity ...................................................................................................II-19

Data Summary & Integration........................................................................................II-21

References.....................................................................................................................II-34

Tables............................................................................................................................II-37

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-88
i

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

of Risks to Human Reproduction (CERHR) 

in 1998. The CERHR is a publicly accessible 

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

associated with exposure to environmental 

and/or occupational chemicals. The CERHR 

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

National Institutes of Health and Dr. Michael 

Shelby is the director.1

The CERHR broadly solicits nominations of 

chemicals for evaluation from the public and 

private sectors. The CERHR follows a formal 

process for review and evaluation of nominated 

chemicals that includes multiple opportunities 

for public comment. Chemicals are selected for 

evaluation based upon several factors including 

the following: 

• potential for human exposure from use 

and occurrence in the environment.

• extent of public concern.

• production volume.

• availability of scientific evidence for 

reproductive and/or developmental toxicity. 

The CERHR convenes a scientific expert panel 

that meets in a public forum to review, discuss, 

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

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

its opinion of the degree to which exposure 

to the chemical is hazardous to humans. The 

panel also identifies areas of uncertainty and 

where additional data are needed. The CERHR 

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

expert panel reports. Expert panel reports are 

made public and comments are solicited. 

Next, the CERHR prepares the NTP-CERHR 

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

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

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

interpretation of the potential for the chemical 

to adversely affect human reproductive health 

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

on the CERHR web site and in hard copy or 

CD-ROM from the CERHR.

Preface

1 Information about the CERHR is available on the 

web at http://cerhr.niehs.nih.gov or by contacting 

the director:

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

Research Triangle Park, NC 27709 

919-541-3455 [phone] 

919-316-4511 [fax]

shelby@niehs.nih.gov [email] 

Information about the NTP is available on the web 

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

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

919-541-0530 [phone]
ii 

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

from NTP member agencies, recommended 

seven phthalates for expert panel review. 

These chemicals were selected because: 

(a) there is the potential for human exposure 

from their widespread use and occurrence within the environment, 

(b) they have a high production volume, 

(c) there is substantial scientific literature 

addressing the reproductive and/or 

developmental toxicities of these chemicals, and 

(d) they are of concern to the public. 

These seven phthalates are as follows:

• di(2-ethylhexyl)phthalate (DEHP) 

• di-isononyl phthalate (DINP) 

• di-isodecyl phthalate (DIDP)

• di-n-butyl phthalate (DBP) 

• butyl benzyl phthalate (BBP) 

• di-n-octyl phthalate (DnOP)

• di-n-hexyl phthalate (DnHP)

Phthalates are a group of similar chemicals 

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

shower curtains, medical devices, upholstery, 

raincoats, and soft squeeze toys. They are not 

bound to the plastics and can leach into the surrounding environment. The scientific literature 

on the reproductive and developmental toxicities of several phthalates is extensive. In addition, there is widespread public concern about 

the safety of phthalates. 

As part of the evaluation of phthalates, the 

CERHR convened a panel of scientific experts 

(Appendix I) to review, discuss, and evaluate 

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

phthalate. There were three public meetings 

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

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

received numerous public comments on the 

phthalates throughout the evaluation process. 

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

includes the NTP brief on DINP, a list of the 

expert panel members (Appendix I), the expert 

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

public comments received on the expert panel’s 

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

single, collective source of information on the 

potential for DINP to adversely affect human 

reproduction or development. Those interested 

in reading this report may include individuals, 

members of public interest groups, and staff of 

health and regulatory agencies. 

The NTP brief included within this report 

presents the NTP’s interpretation of the potential for exposure to DINP to cause adverse 

reproductive or developmental effects in people. It is based upon information about DINP 

provided in the expert panel report, the public 

comments, and additional scientific information available since the expert panel meetings. 

The NTP brief is intended to provide clear, 

balanced, scientifically sound information on 

the potential for DINP exposures to result in 

adverse health effects on development and 

reproduction. 

Introduction
iii

While there are biological and practical reasons for considering developmental toxicity 

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

in mammals, including humans, is a cycle. 

In brief, the cycle includes the production 

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

development, sexual maturity, and, again, 

production of sperm and eggs. 

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

concerns for developmental toxicity were 

addressed by exposing pregnant mothers 

and looking for adverse effects in fetuses. 

Developmental toxicity was detected as 

death, structural malformations, or reduced 

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

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

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

part of the life cycle could lead to adverse 

effects that might only be apparent at a different part of the life cycle. For example, exposure of a sexually mature individual to an 

agent capable of inducing genetic damage 

in eggs or sperm might have no apparent 

effect on the exposed individual. However, 

if a genetically damaged egg or sperm from 

that individual is involved in fertilization, 

the induced genetic damage might lead to 

death or a genetic disorder in the offspring. 

In this example, chemical-induced damage 

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

sexual maturity is attained. Thus, exposure 

of sexually immature animals, either before 

or following birth, to agents or conditions 

that adversely affect development of the 

reproductive system can result in structural 

or functional reproductive disorders. These 

effects may only become apparent after the 

exposed individual reaches the age of puberty or sexual maturity.

Thus, in the case of genetic damage induced 

in eggs or sperm, what might be considered 

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

of adverse effects on development of the 

reproductive tract, developmental toxicity 

results in reproductive disorders. In both 

these examples it is difficult to make a clear 

distinction between developmental and reproductive toxicity. This issue is important 

in considering the phthalate evaluations 

because evidence of developmental toxicity affecting reproductive capacity in later 

stages of the life cycle is reported for at least 

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

Developmental Toxicity versus 

Reproductive Toxicity
1

NTP Brief

What is DINP?

DINP is an oily, viscous liquid with the chemical formula C26H42O4. It is a complex substance 

that contains a mixture of DINP isomers such 

as the one shown in Fig. 1. It is one of a group 

of industrially important chemicals known as 

phthalates. Phthalates are primarily used as 

plasticizers to add flexibility to plastics. DINP 

is used to manufacture a broad range of consumer products such as garden hoses, pool liners, flooring tiles, tarps, and toys. It is not used 

in medical devices and finds only limited use in 

food packaging.

Recent information indicates that approximately 178 million kilograms (392 million pounds) 

of DINP were used in the United States in 

1998.

Are People Exposed to DINP?*

Yes. There are several ways that people may 

be exposed to DINP at home or at work. Human exposure to DINP can occur during the 

manufacture of DINP, during the manufacture 

of DINP-containing products, during the use of 

such products, or through the presence of DINP 

in the environment. Food does not appear 

to be a significant source of exposure. While 

inhalation, ingestion, or skin contact may expose people to DINP, consumer exposure is 

thought to occur primarily by ingestion and 

skin contact. Public concern has been raised for 

exposure of infants and children through the 

mouthing of soft toys made of DINP-containing plastics. The U.S. Consumer Product Safety 

Commission (CPSC) convened a Chronic Hazard Advisory Panel on Diisononyl Phthalate 

(CHAP-DINP) to determine if DINP in consumer products poses chronic health hazards. 

This panel considered hazards to children from 

mouthing DINP-containing toys. Their report 

was issued in June 2001, subsequent to release 

of the Phthalates Expert Panel Reports (CPSC, 

2001).

Because of inadequate information on human 

exposure to DINP, the expert panel took the 

conservative position of assuming that general population exposures in the United States 

would be less than 3-30 μg/kg bw/day (micrograms per kilogram body weight per day). 

This is the range of exposures estimated for the 

more widely used phthalate, DEHP. This assumption was supported by exposure calculations (Kohn et al., 2000; David, 2000) using the 

urine data from Blount et al., 2000. Calculated 

daily exposure estimates indicate that 95% of 

the study population exposed to DINP was exposed to less than 1.7 μg/kg bw/day, and that 

the maximum exposure was 22 μg/kg bw/day. 

The CHAP-DINP (CPSC, 2001) estimated 

that, as a result of mouthing DINP-containing toys, children 0-18 months old could be 

exposed to up to 280 μg/kg bw/day. Children 

19-36 months old could be exposed to up to 

70 μg/kg bw/day. A 15 pound child exposed 

to 280 μg/kg bw/day would be exposed to approximately 2000 μg DINP/day. A 30 pound 

NTP Brief on Di-isononyl Phthalate 

(DINP)

O

O

O

O

Figure 1. Chemical structure of the DINP 

isomer, Di (7-methyloctyl) phthalate

* Answers to this and subsequent questions may 

be: Yes, Probably, Possibly, Probably Not, No 

or Unknown
2

NTP Brief

Clear evidence of adverse effects

Some evidence of adverse effects

Limited evidence of adverse effects

Developmental Toxicity

child exposed to 70 μg/kg bw/day would be 

exposed to approximately 1000 μg DINP/day. 

By comparison, a small drop of water weighs 

approximately 30,000 μg and a grain of table 

salt weighs approximately 60 μg.

Can DINP Affect Human Development or 

Reproduction?

Possibly. Although there is no direct evidence 

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

laboratory animals have shown that exposure 

to DINP can adversely affect development, but 

not reproduction, in rodents (Fig. 2). 

Scientific decisions concerning health risks 

are generally based on what is known as the 

“weight-of-the-evidence.” In this case, recognizing the lack of human data, some evidence 

of developmental effects, and limited evidence 

of no reproductive effects in animals, the NTP 

judges the scientific evidence sufficient to 

conclude that DINP might adversely affect 

development of the human fetus if the levels of 

exposure are sufficiently high (Fig. 3).

It is important to note that the levels of DINP 

exposure that lead to adverse effects on development in rodents are generally far higher than 

those experienced by people.

Summary of Supporting Evidence 

As detailed in the expert panel report, studies 

of reproductive and developmental toxicity in 

rats have shown that exposure of the pregnant 

female to relatively high doses of DINP can 

affect development of the kidneys and skeletal system of the fetus and result in reduced 

birth weights. The reproductive toxicity studies reviewed by the expert panel reported no 

evidence of adverse effects on the reproductive 

system of rats. 

Subsequent to release of the expert panel report, a rodent study was conducted to determine if DINP produced antiandrogenic-like 

effects in male rats following exposure to 750 

mg/kg bw/day from gestation day 14 through 

postnatal day 3 (Gray et al., 2000). Treatment 

resulted in female-like areolas/nipples in some 

male pups, and limited evidence of effects on 

the structure of the male reproductive tract. No 

effects were observed for reproductive tract 

development endpoints that included testis 

weight, anogenital distance, age of preputial 

separation, hypospadias or undescended testes. 

This study provides some evidence that DINP, 

like other phthalates such as DEHP and DBP, 

adversely affects development of the male rat 

reproductive system. However, the use of a single, high dose level of DINP limits the utility of 

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

reproductive effects in laboratory animals 

Reproductive Toxicity

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 Toxicity

Developmental Toxicity
3

NTP Brief

this study in evaluating human health risks.

Are Current Exposures to DINP High 

Enough to Cause Concern?

Probably not. More data are needed to better understand the levels to which people are 

exposed to DINP and how these exposures 

vary across the population. Although the general U.S. population presently appears to be 

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

or developmental effects, 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.

The NTP concurs with the conclusions of the 

CERHR Phthalates Expert Panel and has minimal concern for DINP causing adverse effects 

to human reproduction or fetal development. 

This is based on recent estimated DINP exposures in the U.S. general population. The 

CHAP-DINP (CPSC, 2001) convened by the 

U.S. Consumer Product Safety Commission 

concluded, “...the risk to reproductive and developmental processes in humans due to DINP 

exposure is extremely low or nonexistent.”

The NTP has minimal concern for developmental effects in children. 

This is lower than the “low concern” expressed 

by the expert panel and is based on the CHAPDINP estimates of potential DINP exposures in 

children from mouthing DINP-containing toys. 

The exposure levels at which developmental 

effects were reported in rats (143-285 mg/kg 

bw/day) is about 1000 times higher than the 

upper end of the range estimated for children’s 

exposure (70-280 μg/kg bw/day). 

Figure 3. NTP conclusions regarding the possibilities that human development 

or reproduction might be adversely affected by exposure to DINP 

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 or reproductive effects 

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.
4

NTP Brief

References:

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

Pirkle JL, Sampson EJ, Lucier GW, Jackson RJ, 

Brock JW. Levels of seven urinary phthalate 

metabolites in a human reference population. 

Environmental Health Perspectives 108: 979-

982 (2000).

CPSC. Report to the U.S. Consumer Product 

Safety Commission by the Chronic Hazard 

Advisory Panel on Diisononyl Phthalate 

(DINP). June, 2001. (cited June 2002). <http:

//www.cpsc.gov/LIBRARY/FOIA/Foia01/os/

dinp.pdf>

David RM. Exposure to Phthalate Esters. Environmental Health Perspectives 108: A440 

(2000).

Gray LE, Ostby J, Furr J, Price M, Rao Veeramachaneni DN, Parks L. Perinatal exposure to 

the phthalates DEHP, BBP, and DINP, but not 

DEP, DMP, or DOTP, alters sexual differentiation of the male rat. Toxicological Sciences 58: 

350-365 (2000).

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

Shelby MD, Brock JW, Needham LL. Human 

exposure estimates for phthalates. Environmental Health Perspectives 108: A440-A442 

(2000).
I-1

Appendix I

Appendix I. NTP-CERHR Phthalates 

Expert Panel Report on DINP

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

1

, 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 DINP 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>. 1

Phthalate Expert Panel meeting dates were: August 

17-19, 1999, in Alexandria, VA; December 15-17, 

1999, in Research Triangle Park, NC; and July 12-

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

Appendix I

Robert Kavlock, Ph.D. (chair)

EPA/ORD

Research Triangle Park, NC

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

Brown University

Providence, RI

Robert Chapin, Ph.D.

NIEHS

Research Triangle Park, NC

Michael Cunningham, Ph.D.

NIEHS

Research Triangle Park, NC

Elaine Faustman, Ph.D.

University of Washington

Seattle, WA

Paul Foster, Ph.D.

Chemical Industry Institute of Toxicology

Research Triangle Park, NC

Mari Golub, Ph.D.

Cal/EPA

Davis, CA

Rogene Henderson, Ph.D.

Inhalation Toxicology Research Institute

Albuquerque, NM

Irwin Hinberg, Ph.D.

Health Canada

Ottawa, Ontario, Canada

Ruth Little, Sc.D.

NIEHS

Research Triangle Park, NC

Jennifer Seed, Ph.D.

EPA/OPPT

Washington, DC

Katherine Shea, M.D.

North Carolina State University

Raleigh, NC

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

FDA

Rockville, MD

Shelley Tyl, Ph.D.

Research Triangle Institute

Research Triangle Park, NC

Paige Williams, Ph.D.

Harvard University

Cambridge, MA

Tim Zacharewski, Ph.D.

Michigan State University, 

East Lansing, MI

Appendix I. NTP-CERHR Phthalates Expert Panel 

(Name and Affiliation)
�����������������������������������

���������������������

���������������������������

��������������������������������������������

Appendix II

NTP-CERHR EXPERT PANEL REPORT 

on

Di-isononyl Phthalate

October 2000 NTP-CERHR-DINP-00
TABLE OF CONTENTS

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

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

Appendix II
Appendix II

II-i

PREFACE

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

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

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

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

caused by agents to which humans may be exposed.

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

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

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

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

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

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

concern.

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

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

report has been reviewed by the CERHR Core Committee made up of representatives of NTP-participating agencies, by CERHR staff scientists, and by members of the Phthalates Expert Panel. 

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

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

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

scientific evidence that adverse reproductive/development health effects are associated with exposure to specific chemicals or classes of chemicals, including descriptions of any uncertainties that 

would diminish confidence in assessment of risks; and (3) identify knowledge gaps to help establish 

research and testing priorities.

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

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

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

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

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

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

Alexandria, Virginia.

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

CERHR

Sciences International, Inc.

1800 Diagonal Road, Suite 500

Alexandria, VA 22314-2808

Telephone: 703-838-9440
II-ii

Appendix II

A Report of the CERHR Phthalates Expert Panel: 

Name Affiliation

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

 USEPA, Research Triangle Park, NC

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

Robert Chapin, PhD NIEHS, Research Triangle Park, NC

Michael Cunningham, PhD NIEHS, Research Triangle Park, NC

Elaine Faustman, PhD University of Washington, Seattle, WA

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

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

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

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

Ruth Little, ScD NIEHS, Research Triangle Park, NC

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

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

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

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

Paige Williams, PhD Harvard University, Boston, MA

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

With the Support of CERHR Staff:

NTP/NIEHS

Michael Shelby, PhD Director, CERHR

Christopher Portier, PhD Acting Associate Director, NTP

Gloria Jahnke, DVM Technical Consultant

Lynn Goldman, MD Technical Consultant

Sciences International, Inc.

John Moore, DVM, DABT Principal Scientist 

Annette Iannucci, MS Toxicologist 

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

II-1

1.0 CHEMISTRY, USAGE, AND EXPOSURE

1.1 Chemistry

Figure 1: Chemical Structure of a Di-isononyl Phthalate Isomer

Di (7-methyloctyl) phthalate

O

O

O

O

DINP is a complex substance assigned two different CAS Registry Numbers (RN). CAS RN 

68515-48-0 (designated DINP-1 in this document) is manufactured from octene that is converted to 

alcohol moieties consisting mainly of 3,4-, 4,6-, 3,6-, 3,5, 4,5-, and 5,6-dimethyl-heptanol-1. CAS 

RN 28553-12-0 (DINP-2) is produced from n-butene that is converted primarily to methyloctanols 

and dimethylheptanols. The 28553-12-0 CAS RN also represents DINP-3 which is produced from 

n-butene and isobutene that are converted to alcohols, with 60% consisting of methylethyl hexanols. According to the American Chemistry Council (ACC; formerly CMA), DINP-3 is no longer 

produced (1). The ACC (2) has stated that although DINP is a complex substance, it is not variable 

due to the stability of the alcohol manufacturing process. The two types of DINP are considered 

commercially interchangeable. 

DINP is an oily, viscous liquid at standard temperature and pressure.

Table 1: Physicochemical Properties of DINP

Property Value

Chemical Formula C26H42O4

Molecular Weight 419

Melting Point -48 oC

Boiling Point 370 oC

Specific Gravity 0.97 

Solubility in Water Insoluble (<0.001 mg/L)

Log Kow 9

(3)

1.2 Exposure and Usage

Humans may be exposed to DINP by the oral, dermal, and inhalation routes of exposure. Occupational exposure occurs primarily through inhalation and dermal contact, while consumer exposure 

occurs primarily through oral and dermal routes. Exposure of children to DINP through children’s 

products is a public concern. 
II-2

Appendix II

Occupational exposure

DINP, like other phthalate esters, is manufactured within a closed system under negative pressure. 

However, some exposures may occur during the loading and unloading of railroad cars and trucks. 

Slightly higher exposures may occur during the production of PVC products because of elevated 

temperatures and more open processes. ACC (1) cites six studies that indicate exposures are below 1

mg/m3 during production of phthalates and below 2 mg/m3 during production of PVC. As discussed 

in Section 2.2, dermal exposure is not expected to result in significant absorption into the body.

Consumer exposure

The range of products that contain DINP is quite broad. The use, categories, and amounts used of 

DINP in 1998 are given in Table 2. 

Table 2: Calculated 1998 US Consumption of DINP (thousands of metric tons)

End Use Subtotal Total

Film and Sheet 13

Stationary and Wood Veneer 6

Pool Liners 1

Other 6

Flooring 48

Tiles 23

Sheets 25

Artificial Leather 3

Coated Fabrics 21

Tarps 16

Conveyor Belts 1

Other 4

Dip Coating/Slush Molded 30

Gloves 15

Toys 6

Traffic Cones <1

Other ~9

Tubings and Profiles 7

Profiles 5

Garden Hoses 2

Wire and Cables 32

Shoes/Shoe Soles 9

Under-Body Coating 7

Sealants (carpet backing) 8

GRAND TOTAL 178

 (1)
Appendix II

II-3

DINP is a general purpose plasticizer with a broad range of applications used in flexible PVC. It is 

widely used in the toy, construction, and general consumer product markets. It has limited use in 

food packaging and is not used in medical applications.

Because of physicochemical similarities between DINP and DEHP, general exposure to DINP is 

probably very similar to exposure to DEHP, but few monitoring data were located. Based on data 

for other phthalates, one could speculate that environmentally-contaminated food represents a primary route for human exposure. However, the data are scant in support of this view.

DINP’s solubility in water is extremely low; levels are often below the analytical detection limit. 

Vapor pressure is also extremely low, so measured concentrations in air are not available. Modeling 

based on physicochemical properties of DINP can be compared to similar models for DEHP.

Food 

In 1996, dinonyl phthalate (isomer not specified) was identified but not quantified in 4 of 12 infant 

formulas from the UK (4). In a follow-up survey conducted by the Ministry of Agriculture, Fisheries, and Food (MAFF) (5), DINP was not specifically targeted, but there was no evidence of its 

presence in 39 samples of infant formula from the UK. In a UK survey of fatty foods (e.g., dairy 

products, meats, fish, eggs, and oils), DINP was not detected at an analytical limit of 0.01 mg/kg (6). 

Toys 

PVC plastics are often used in children’s products. Different phthalates are constituents of PVC; 

DINP is currently the predominant plasticizer (7). Other phthalates, including DEHP, have been or 

are also used (8, 9). US toy manufacturers began voluntary removal of DEHP from pacifiers and 

nipples in 1986 (10). Few studies pertaining to plasticizers in children’s products were found in the 

peer-reviewed literature. Additional information is available from industry groups and several government agencies. The Expert Panel did not perform a comprehensive review of available data, but 

believes the information it reviewed reflects the general state of knowledge.

As reported by the Consumer Product Safety Commission (CPSC) (7), Chen measured DINP in 31 

of 35 toys and found a concentration range of 15.1–54.4 % dry weight. Health Canada (11) analyzed 41 children’s products made in the US, China, and Thailand for the presence of DINP and 

DEHP. DINP was detectable in 27 of the 41 products in concentrations that ranged from 3.9 to 44% 

dry weight. Criteria for the selection of products were not discussed in any of these surveys. No 

information on market share, length of availability on the market, or estimates of the numbers of 

products in circulation was noted in any study. Only Health Canada listed product number, country 

of origin, manufacturer/distributor, and brand. All studies listed a product description. Marin (9) 

analyzed 15 samples of materials used in toys in Spain. The authors noted that the PVC contained a 

mixture of plasticizers including DINP, DEHP, and DIDP, but reported only the DEHP content. 

The estimation of actual exposure of children to phthalates contained in children’s products has 

been studied. In vitro studies using various agitation and impaction approaches yield a wide range 

of extraction of DINP from toys. CPSC used stainless steel pistons, 11 cm2 of each product, and 

simulant saliva to obtain extraction rates for their 31 DINP-positive children’s products. Migration 
II-4

Appendix II

was log normally distributed with a mean rate of 8.2 – 9.83 μg/11 cm2/hour and a range of 1– 48 

μg/11 cm2/hour. Both CPSC (7) and Health Canada (11) failed to find any correlation between 

release rate of DINP under experimental conditions and total DINP content.

The Dutch Consensus Group reported a small study by Meuling and Rijk (12) using 20 adult volunteers. A control specimen without DINP and three specimens with DINP were used; specimen 

1 contained 38% DINP. Specimens 2 and 3 came from different parts of the same commerciallyavailable teething ring, representing different shapes for mouthing, and contained 43% DINP each. 

All three were 10 cm2 total area. All 20 volunteers were instructed to suck and bite on the control 

specimen for 10–15 minutes, all saliva was collected, volunteers rested 5 minutes and then they 

performed 4 separate sessions on the same test piece of specimen 1, resting 5 minutes between 

each session. This procedure was repeated with half (n=10) of the volunteers on specimen 2 and 

the other half (n=10) on specimen 3. DINP extraction from specimen 1 was 1.38 (0.3–8.3) μg/min, 

from specimen two 2.44 (0.9–8.9) μg/min, and from specimen three 1.63 (0.9–5.7) μg/min. The 

mean across all groups was 1.8 μg/10cm2/min (or 120 μg/11cm2/hour). There was no correlation 

between extraction and pH or protein content of the saliva. Release rates over the various 15-minute 

intervals seemed consistent. The increase in extraction of Specimen 2 was thought to be due to the 

finger-like shape resulting in different mouthing behaviors from those employed on the disk-like 

shape of Specimens 1 and 3.

CPSC (7) reported a similar protocol using 10 adult volunteers and 5 toys, and found a mean 

migration rate of 241.3 μg/11cm2/hour. This rate was 39.5 times higher than the average rate 

obtained by impaction with disks cut from the same 5 toys, but was similar to the ranges in the 

Dutch simulation study. 

Exposure of children to DINP from PVC toys was estimated by Fiala et al. (13) in Austria. DINP 

levels were measured in the saliva of 10 adult volunteers who first sucked on and then sucked and 

chewed on 10–15 cm2 pieces of teether (containing about 36% DINP) for 1 hour. In the experiment where the volunteers only sucked on the sample, the migration rates of DINP ranged from 

297–1,452 μg/dm2/hour with a mean migration rate of 832±397 μg/dm2/hour. Using assumptions of 

an 8 kg body weight, 3-hour exposure time (12), and 10 cm2 mouthing area, mean and maximum 

exposure levels of 31.25 μg/kg bw/day and 54.4 μg/kg bw/day, respectively, were estimated. For 

the experiment where the adults chewed on the sample, migration rates of DINP in 9 adults ranged 

from 768−2152 μg/dm2/hour. Using the same assumptions from the first experiment, a maximum 

exposure level of 84.5 μg/kg bw/day was estimated. 

CPSC (7), the Dutch Consensus Group (12), and Health Canada (11) have attempted to calculate 

daily intake based upon the leaching rates described above. The Dutch Group used Monte Carlo 

simulation and estimates of mouthing time and the leaching rates from the in vivo study of 20 

adults. Mouthing time was derived from parent observations and logging of mouthing time of 42 

children aged 3−35 months (Table 3). Mouthing time was calculated for the time children were 

awake, but not eating, during ten 15 minute observation periods over 2 days. Logs were kept of 

objects mouthed; the objects were divided into those intended for mouthing and those not intended 

for mouthing. The Dutch calculations used total mouthing time excluding time spent mouthing pac-
Appendix II

II-5

ifiers. Because the greatest exposure levels were determined for children within the ages of 3−12 

months, the results for that age group are summarized in Table 4.

Table 3: Total Mouthing Time

Age (months) Sample Size Mean in Minutes (SD) Min (Minutes) Max (Minutes)

3–6 5 36.9 ±19.1 14.5 67.0

6–12 14 44.0 ± 44.7 2.4 171.5

12–18 12 16.4 ± 18.2 0 53.2

18–36 11 9.3 ± 9.8 0 30.9

Table 4: Toy Exposure Estimates for Children Aged 3–12 Months.

Estimated Intake Level (μg/kg bw/day)

Agency Mean 95th Percentile 99th Percentile Maximum

RIVM* 6.53–14.4 20.7–39.7 39.8–77.3 70.7–204

CPSC 5.7 94.3 - -

Health Canada** 44 73.9*** 173.5*** 320

* Exposure range for 3−6 month-old and 6−12 month-old children; 

 range includes results from 3 specimens tested.

** Calculated with mouthing times for teethers and other objects intended for mouthing.

*** Results using Monte Carlo simulations in children aged 3–6 months.

The approach taken by Health Canada used published data and 10,000 Monte Carlo simulations and 

total mouthing time from the Dutch observation study including mouthing of pacifiers, teethers and 

other objects intended for mouthing. CPSC used the same mouthing-time data, but limited its calculations to the mouthing time of objects not intended for mouthing. They performed a log transformation of the time because of the extreme skewness in the sample and calculated a geometric mean 

mouthing time of 12.03 minutes (95% CI 6.2–23.3). Exposure estimates were made using a log 

linear model, the mean leaching rate from mechanical extraction from 31 consumer products and a 

39.5 factor (to adjust for differences between in vitro and in vivo extraction rates). The differences 

in the analyses resulted in quite different exposure estimates, which explains the different conclusions and recommendations of the agencies.

The differences also highlight the uncertainties inherent in these calculations. Because extraction 

of DINP does not correlate with DINP content, because extraction is highly variable across both 

laboratory procedures and human subjects, and because the number and distribution of children’s 

products containing DINP is unknown, the amounts of DINP presented to a child cannot be well 

characterized. Furthermore, the estimates of mouthing behavior in the youngest and potentially 

highest risk group, 3–12 months, are based upon only 19 children. No discussion of developmental 

age, physical condition, ethnicity, or other socio-demographic indicators is included in the small 

parental observation study. These numbers are preliminary estimates at best. Standardization of laboratory techniques with correlation with in vivo simulations, better data on product distribution and 

use, and independent studies of mouthing behavior in babies and young children are needed. None-
II-6

Appendix II

the-less, existing models show this is a potentially significant exposure for young children. A study 

using larger numbers of children has been submitted by Juberg et al. (14), but could not be cited at 

the time of this review. According to the ACC (2), the CPSC and EU Joint Research Laboratory are 

working on standardizing laboratory techniques with in vivo simulations. 

Dermal exposure to DINP from toys may also occur, but has not been studied specifically in children.

Exposure Estimate

Based on the physicochemical characteristics of DINP and limited monitoring data, the Expert 

Panel believes it reasonable to assume that exposure to DINP in the general adult population is 

lower than exposure to DEHP, which is estimated at 3–30 μg/kg bw/day (15). Children may incur 

significantly greater nondietary exposures from mouthing toys and other articles containing DINP.

The summary for Section 1 is located in Section 5.1.1.
Appendix 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

BIBRA (16) (Table 7-1) conducted a 21-day dietary study in 6-week-old F344 rats where groups 

of 5 males and 5 females were fed concentrations of 0, 0.6, 1.2, or 2.5% DINP (M: 639, 1,192, or 

2,195 mg/kg bw/day; F: 607, 1,193, or 2,289 mg/kg bw/day). The test material most likely consisted of a mixture of DINP represented by CAS numbers 68515-48-0 and 28553-12-0 (DINP-1 

and DINP-2). A positive control group of 5 rats per sex was exposed to 1.2% DEHP (M: 1,084 

mg/kg bw/day; F: 1,063 mg/kg bw/day). Body weight and food intake were measured twice weekly. 

On day 21, rats were killed and necropsied. Liver, kidney, and testes were preserved in formalin and 

examined histologically. Peroxisomal proliferation was assessed by measuring activities of peroxisomal proliferation enzymes and by examining liver tissue by electron microscopy. 

A significant decrease in weight gain was observed in the mid- and high-dose groups. Food intake 

was significantly reduced in males. Organ to body weight ratios that were significantly increased 

in all treatment groups included liver (M: 136, 173, and 232%, F: 131, 175, and 237% of control 

values) and kidney (M: 115, 122, and 124%, F: 107, 108, and 114% of control values). Histopathological changes were not observed in kidneys; changes in liver were limited to reduced cytoplasmic basophilia in the mid- and high-dose group and increased cytoplasmic eosinophilia in the 

high-dose group. Palmitoyl-CoA (PCoA) oxidase activity was significantly increased in the midand high-dose groups (M: 452 and 1,035%; F: 376 and 1,104% increases, respectively, compared 

to controls) and an increase in peroxisome numbers was observed by electron microscopy in livers 

from the high-dose group. The activity of 11-hydroxylase and 12-hydroxylase was significantly 

increased in males of all dose groups and in females of the high-dose group. Significant changes 

observed in all treatment groups included increased total liver proteins and reductions in serum 

levels of cholesterol. Serum triglyceride levels were significantly reduced in all treated males, but 

increased in mid- and high-dose females. The testes to body weight ratio was significantly increased 

in the high-dose males (135% of control value), but absolute testes weights were not significantly 

affected. Testicular lesions were not observed with the exception of severe unilateral atrophy in one 

male of the mid-dose group. Treatment with 1,063−1,084 mg DEHP/kg bw/day resulted in similar 

effects including decreased weight gain, increased liver and kidney to body weight ratio, increased 

liver enzyme activities, and reduced serum levels of cholesterol and triglycerides. Moderate testicular atrophy was noted in one male. Peroxisomal proliferation is of particular interest and an increase 

in peroxisome numbers was observed after treatment with DEHP. PCoA activity was significantly 

increased to 683 and 540% of control values for males and females, respectively. The increase in 

peroxisomal enzyme activity in rats treated with 1,063−1,084 mg/kg bw/day DEHP was greater 

than that obtained by treatment with DINP at 1,192−1193 mg/kg bw/day (452 and 376% of control 

values in males and females, respectively).
II-8

Appendix II

This study provides evidence that the liver is a target organ of DINP. A pattern similar to effects 

noted with DEHP is seen: increased liver weight and induction of hepatic peroxisome proliferation. 

The testes do not appear to be a target organ at these dose levels. The study provided a LOAEL of 

0.6% (607[F] and 639[M] mg/kg bw/day) and no NOAEL was identified. 

In a 2-year dietary study, (17) (Table 7-2) systemic effects resulting from DINP-1 exposure in adult 

(6 week old) Fischer 344 rats were evaluated. Groups of 110 rats per sex were fed diets containing 

0, 0.03, 0.3, and 0.6% DINP-1 (males: 0, 15, 152, and 307 mg/kg bw/day; females: 0, 18, 184, and 

375 mg/kg bw/day). Body weight and food intake were measured weekly. Ten rats/sex/group were 

killed and necropsied at 6, 12, and 18 months; the remaining rats were killed and necropsied at 

the end of the 2-year study. Evaluation of hematology, urine, and blood chemistry effects was performed at 6, 12, 18, and 24 months. Histopathological evaluations were conducted on the liver and 

the kidney from all dose groups and in the remaining organs of the control and high-dose groups. 

Evidence of peroxisome proliferation was determined by microscopic examination of livers of 2 

rats/sex/group at 24 months.

Significant reductions in body weight gain were observed in males from 18–24 months in the 152 

mg/kg bw/day group and from 12–24 months in the 307 mg/kg bw/day group. Food intake levels were not reported. Survival was significantly decreased in females of the 184 and 375 mg/kg 

bw/day groups. Liver and kidney to body weight ratios were significantly increased throughout the 

study in both sexes in the mid- and high-dose groups (152–375 mg/kg bw/day). Spleen to body 

weight ratios were significantly increased in males and females of the high-dose group (307–375 

mg/kg bw/day) at 24 months. A small but significant increase in adrenal to body weight ratio was 

reported for females in the 375 mg/kg bw/day group at 6–12 months, and in both sexes in the 

high-dose group (307–375 mg/kg bw/day) at 24 months. Adrenal weights were not listed in tables. 

Dose-related changes in liver included hepatocyte enlargement in high-dose males and females 

throughout treatment. At 24 months, dose-related liver effects included regenerative nodules and 

focal necrosis in males and females of the two highest dose groups, and spongiosis hepatitis in 

males of the high-dose group. In males of the mid- and high-dose groups, consistent increases 

in serum glutamic oxaloacetic transaminase (SGOT) and serum glutamic pyruvic transaminase 

(SGPT) activities were observed. However, for SGOT, statistical significance was obtained only 

at 6 and 12 months in the mid-dose group and at 6–18 months in the high-dose group. In males, 

increases in SGPT activity were statistically significant at 24 months in the mid-dose group and 

at 6 and 18 months in the high-dose group. An increased incidence of mononuclear cell leukemia 

(MNCL) was observed in both sexes of the mid- and high-dose groups. Peroxisome proliferation 

did not occur and there was no evidence of treatment-related lesions in testes or female reproductive organs. The only significant dose-related changes in hematology were a reduction in red blood 

cell count and hemoglobin and hematocrit values in males of the 307 mg/kg bw/day group at 24 

months. Urinalysis results were not listed in tables, but authors reported increased urine volumes in 

high-dose males at all time points and transient increases in potassium and glucose. A NOAEL of 

17 mg/kg bw/day was selected by the authors.

A second 2-year dietary feeding study in F344 rats was reported by Moore et al. (18) (Table 7-

3). Groups of 70–85 F344 rats/sex/group (6 weeks old) were fed concentrations of 0, 500, 1,500, 
Appendix II

II-9

6,000, and 12,000 ppm DINP-1 (males: 0, 29.2, 88.3, 359, or 733 mg/kg bw/day; females: 0, 36.4, 

109, 442, or 885 mg/kg bw/day). Body weight and food intake were measured weekly through 

weeks 16–17 and monthly thereafter. Standard hematological, clinical chemistry, and urinalysis 

parameters were measured every 26 weeks. Peroxisome proliferation was measured in 5 rats/sex in 

control and high-dose groups at weeks 1, 2, and 13, and in 3–5 rats/sex in the control and 2 highest 

dose groups at week 104. Five rats/sex/group were sacrificed and necropsied at weeks 1, 2, and 13. 

Fifteen rats/sex/group were killed and necropsied at week 79. The remaining rats were sacrificed 

and necropsied at week 104. Another group of 55 rats/sex was exposed through diet to 12,000 ppm 

(males 637 mg/kg bw/day; females 774 mg/kg bw/day) DINP for 78 weeks and sacrificed at week 

104 in order to determine if recovery occurs after exposure to DINP has ended. Histopathological 

evaluations were conducted on major organs from rats in all dose groups.

Clinical signs of toxicity were observed in rats exposed to 359 mg/kg bw/day and higher, and 

included hunched posture, decreased activity, bodies that appeared pale and thin, and fewer feces. 

Rats exposed to 733–885 mg/kg bw/day experienced a statistically significant reduction in weight 

gain accompanied by a decrease in food intake. Survival was significantly reduced in the high-dose 

males with only 54% surviving to the end of the study. The body weight effect was shown to be 

partially reversible because male weight gain in the recovery group was not reduced at week 104; 

reduced weight gain in females was less pronounced. Survival was not significantly affected in 

the recovery group. The authors reported that the dose-related depression of body weight gain in 

the two highest doses was associated with clinical chemistry findings or histomorphologic effects 

in liver and kidney. A significant increase in the incidence of anemia, as observed by decreases in 

erythrocyte, hemoglobin, and hematocrit levels, was observed throughout the study in rats exposed 

to 359 mg/kg bw/day and higher, but was not observed in the recovery group. A significant increase 

in kidney to body weight ratio was observed in rats exposed to 359 mg/kg bw/day and higher from 

week 79 to 104 (M: 8.1 and 25% and F: 14.4 and 22% increases in 2 highest dose groups, respectively, at week 104). Liver to body weight ratios were significantly increased in both sexes exposed 

to 359 mg/kg bw/day and higher throughout the study (M: 35 and 61% and F: 26 and 71% 

increases at 2 highest doses, respectively, at week 104). Histological effects observed in kidneys 

of rats exposed to 359 mg/kg bw/day and higher at weeks 79 and 104 included an increased incidence and severity of renal papilla mineralization in males (59/85 and 57/85 at 2 highest doses). An 

increase in tubule cell pigmentation was also reported by the authors, but the incidence of the lesion 

appeared equal among control and dose groups (55−59 sex/dose). Urinalysis findings at week 104, 

which included significant increases in urine output and corresponding decreases in potassium, 

calcium, creatinine, and chloride levels in high-dose males, suggested compromised ability to concentrate in the renal tubule epithelium. Serum urea levels were significantly increased during the 

second half of the study in rats exposed to 359 mg/kg bw/day and higher. Increases in urine volume 

and kidney lesions were observed in the recovery group exposed to 733 mg/kg bw/day and greater 

with severity approximately equal to that of the 359−442 mg/kg bw/day treatment group. Livers of 

rats exposed to 359 mg/kg bw/day and higher appeared enlarged and granular at weeks 79 and 104. 

Histopathological effects in the livers of the high-dose group included diffuse hepatocyte enlargement (37/85 males and 52/85 females), cytoplasmic eosinophilia (43/85 males and 45/85 females), 

and Kupffer cell/bile canaliculi pigmentation (12/85 males and 17/85 females). These effects were 

first detected at weeks 2, 13, and 79, respectively. The authors also reported alterations in serum 

alanine aminotransferase and aspartate aminotransferase activity, but the changes did not appear to 
II-10

Appendix II

be consistent or dose related. Non-neoplastic liver changes were found to be reversible in the recovery group. Peroxisomal enzyme activity was significantly increased at week 104 in females exposed 

to 442 mg/kg bw/day and in both sexes of the high-dose group throughout the study. The recovery 

group was not tested for peroxisomal enzyme activity. Histopathological changes in testes or female 

reproductive organs were not observed. 

Neoplastic effects included a significant increase in liver adenomas (10/80 vs 4/80) and carcinomas 

(11/80 vs 1/80) in male rats of the high-dose group at week 104. At week 104, renal tubule cell carcinoma was observed in 2 males of the high-dose group and 4 males of the recovery group. Mononuclear cell leukemia was found in 45–49% of rats in the 2 highest dose groups. Liver neoplasms 

were not observed in the recovery group, but the incidence of renal tubule cell carcinoma in males 

and mononuclear cell leukemia remained elevated compared to controls. The authors selected a 

NOAEL of 1,500 ppm (88.3−109 mg/kg bw/day) for this study.

In a 2-year dietary study in 6-week-old B6C3F1/CrlBR mice (19) (Table 7-4), groups of 70 mice/

sex/group ate diets that contained 0, 500, 1,500, 4,000, and 8,000 ppm DINP-1 (males: 0, 90.3, 

276, 742, 1,560 mg/kg bw/day; females: 0, 112, 336, 910, 1,888 mg/kg bw/day). Body weights and 

food intake were measured weekly through week 16–17 and monthly thereafter. Standard, hematological, clinical chemistry, and urinalysis parameters were measured every 26 weeks. Peroxisome 

proliferation was measured in five mice/sex in the control and high-dose group at the midpoint and 

end of the study. Fifteen mice/sex/group were sacrificed and necropsied at week 79. The remaining 

mice were sacrificed and necropsied at the end of the 2-year study. Histopathological evaluations 

were conducted on major organs from mice in all dose groups. Another group of mice was exposed 

to 8,000 ppm DINP in the diet for 78 weeks and sacrificed at week 105−106 in order to determine if 

recovery would occur after exposure to DINP ended. 

Toxicological and non-neoplastic effects were observed in mice that received the 2 highest doses, 

742 mg/kg bw/day and greater. A statistically significant reduction in weight gain occurred throughout the study; this reduction was not accompanied by a decrease in food intake. The effect was 

shown to be partially reversible because female weight gain in the recovery group was not reduced 

at week 104; reduced weight gain in males was less pronounced. Clinical signs of toxicity were 

observed and included abdominal swelling in males exposed to 742 mg/kg bw/day and greater, 

and hunched posture, decreased activity, and fewer feces in the high-dose males. Survival was 

significantly reduced in the high-dose males (1,560 mg/kg bw/day), with only 63% of males surviving until the end of the study. Survival was not significantly affected in the recovery group. A 

significant reduction in kidney to body weight ratio was observed in males of the 2 highest dose 

groups (13 and 25% reduction, respectively), whereas a significant increase in liver to body weight 

ratio occurred (7 and 24% increase, respectively). Females exposed to the highest dose (1,888 

mg/kg bw/day) had a 37% increase in liver to body weight ratio from week 79 to 104. Histological 

examination revealed an increased incidence and severity of renal nephropathy in female mice of 

the high-dose group. Urinalysis findings, which included significant increases in urine output and 

corresponding decreases in sodium, potassium, and chloride levels in high-dose mice from week 

52–104, suggested compromised ability to concentrate in the renal tubule epithelium. The effects 

on renal structure and function proved to be partially reversible as they were less pronounced in 

mice of the recovery group by the end of the study. Histopathological liver changes were observed 
Appendix II

II-11

in mice of the highest dose group and included diffuse hepatocyte enlargement (56/70 males and 

65/70 females) and cytoplasmic eosinophilia (67/70 males and 68/70 females) and pigment (64/70 

males and 53/70 females). Other hepatic effects included increased serum alanine aminotransferase 

and aspartate aminotransferase levels in the high dose males at various time points throughout the 

study. Non-neoplastic liver changes were found to be reversible in the recovery group. An increase 

in peroxisomal enzyme activity in mice exposed to 1,560−1,888 mg/kg bw/day indicated that hepatocyte enlargement was due to peroxisomal proliferation. Histopathological changes in testes or 

female reproductive organs were not observed.

Neoplastic effects included increased incidences of hepatic adenomas and carcinomas combined 

in females exposed to 336 mg/kg bw/day (10/60 versus 3/70), and adenomas (15/60 and 13/60 

versus 10/70) and carcinomas (17/60 and 20/60 versus 10/70) in males exposed to 742 and 1,560 

mg/kg bw/day, respectively, and in females exposed to 1,888 mg/kg bw/day (18 adenomas and 18 

carcinomas/70 versus 2 adenomas and 1 carinoma/70). The occurrence of hepatic neoplasms was 

lower in the recovery group compared to the high dose mice exposed for the duration of the study 

with an incidence of 37−39% versus 50−56%. Based on hepatic neoplasms, the authors selected 

a NOAEL of 500 ppm (112 mg/kg bw/day) for females and 1,500 ppm (276 mg/kg bw/day) for 

males.

Hall et al. (20) (Table 7-5) exposed sixteen 25-month-old marmosets (2/sex/group) by gavage for 

13 weeks with DINP (CAS number not provided) in 1% methylcellulose and 0.5% Tween at concentrations of 0, 100, 500, or 2,500 mg/kg bw/day. Clofibrate was administered as a positive control 

at 500 mg/kg bw/day. Analysis was conducted for hematology (weeks 0, 6, and 13), blood chemistry (weeks 0, 4, and 13), estradiol and testosterone levels (week 12), and urine composition (weeks 

0, 5, and 12). The main organs (including but not limited to liver, testes, and epididymides) were 

weighed and examined histologically (testes and epididymides were preserved in Bouin’s). Peroxisomal proliferation was determined by measuring cyanide-insensitive PCoA oxidase activity. 

Clinical signs observed in the marmosets included ungroomed coats and localized reddening of 

the skin around the anus and legs which was likely caused by excretion of test substance in feces. 

One male exposed to 2,500 mg/kg bw/day experienced a 13% weight loss and had reduced activity 

and a hunched posture. Weight loss or decreased weight gain was observed in 2 males and 1 female 

exposed to 2,500 mg/kg bw/day. Peroxisome proliferation was not evident as indicated by a lack of 

dose-related increases in PCoA oxidase activity. There were no DINP-treatment related changes in 

estradiol or testosterone levels, hematology, blood chemistry, organ weights, urine composition, or 

microscopic findings. The authors identified a NOAEL of 500 mg/kg bw/day.

Administration of the positive control, clofibrate, did result in an approximate 100% increase in 

PCoA oxidase activity. Other effects in positive control animals included an increase in 11-hydroxylase activity in males, reduced weight gain, anemia, and a slight increase in relative and absolute 

kidney weight. 

Pugh et al. (21) gavaged 2-year-old (prepubertal) cynomolgus monkeys (4/group) with 0 or 500 

mg/kg bw/day DINP-1 in methylcellulose for 14 days (Web Table 9). According to Short et al. (22), 
II-12

Appendix II

500 mg/kg bw/day is the maximum dose that can be absorbed by the monkeys. On day 15, the animals were sacrificed and the tissues were removed, weighed, and fixed in formalin for histopathological evaluation. Hematology, serum chemistry, and urine analysis were conducted. Peroxisomal 

proliferation was examined by measuring peroxisomal beta oxidation activity and replicative DNA 

synthesis. Gap junctional intercellular communication was determined in liver. There were no clinical signs or changes in body weight gain. A significant increase in blood neutrophil numbers and 

decrease in lymphocyte count were the only effects reported. There were no testicular or hepatic 

lesions and no effects on any of the systemic parameters examined.

Mode of Action

The renal neoplasia in male rats appears to be due to alpha-2-microglobulin nephropathy which is a 

mechanism not considered relevant to humans (23). However, an increased rate of nephropathy was 

seen in female mice exposed to 1,888 mg/kg bw/day which would not be consistent with the alpha2-microglobulin mechanism. The Moore (18) study demonstrated liver tumors in rats only in the 

highest-dose males. Peroxisome proliferation in rats was observed at the highest dose in males and 

females, and the second highest dose in females but not males. No liver tumors were observed in 

either sex at the second highest dose level. In addition, no liver tumors were noted in the recovery 

groups. These results are consistent with a peroxisome proliferation mode of action for hepatic 

tumor induction. Unfortunately, peroxisome proliferation was assayed in mice only at the highest 

dose, and liver tumors were also observed at lower doses.

2.2 Toxicokinetics

Phthalate Moiety

Absorption

Rodents: Dermal

Dermal absorption of 14C-DINP was studied in male Fischer 344 rats (24) in both conditioned (pretreatment with non-labeled DINP) and non-conditioned skin. Following exposure, the dosed area 

was occluded. Under all conditions, the amount absorbed after 7 days ranged from 2 to 4% of the 

dose. Approximately 93−99% of the administered radioactivity was recovered at the site of application. Radioactivity in feces and gut of the exposed rats suggested some excretion via the biliary 

route. In in vitro studies comparing absorption of DEHP through human and rat skin (25), absorption through human skin was slower than through rat skin. Therefore, the dermal absorption rate of 

DINP is also expected to be slower through human versus rat skin. Studies conducted by Deisinger 

et al. (26) have demonstrated that dermal absorption of DEHP from a plasticized film is slower than 

dermal absorption of neat DEHP. It is reasonable to assume that these results apply to DINP.

Rodents: Oral

Oral absorption of 14C-DINP (dose=2,500 mg/kg) was studied (27) in conditioned (pre-treatment 

with non-labeled DINP) and non-conditioned male albino rats. The rats were administered 0.5 mL 

of radiolabeled DINP by gavage and the dose was estimated at approximately 2,500 mg/kg bw by 

the Expert Panel based on the density of DINP and reported rat body weights. Within 72 hours, 

85% of the administered dose was excreted in the feces, most within the first 24 hours. The rest of 
Appendix II

II-13

the dose was excreted in urine (average of 12%) or remained in the tissues (trace amounts). Thus, 

the oral absorption was approximately 12%. In studies at Midwest Research Institute (28), male and 

female Fischer 344 rats were dosed orally either in a single or in 5 daily doses of 50, 150, or 500 

mg/kg. At least 49% of the single low dose was absorbed. Absorption was decreased at the high 

single dose and at all doses following repeated exposures.

Biotransformation

Most of the 14C collected in the urine of rats following a single oral dose of 14C-DINP was in the 

form of phthalic acid or side-chain oxidation products of the monoester (MINP) (28). The relative 

amount of phthalic acid in the urine decreased at the high dose. The monoester itself, as well as 

the diester, was present in only trace amounts. In feces, 8 and 41% of the radioactivity was associated with the diester following administration of a low (50 mg/kg) or a high (500 mg/kg) oral dose 

of 14C-DINP. This indicates saturation of metabolism at the high dose. The remainder of the fecal 

radioactivity was associated with the monoester or its side-chain oxidation products. Major metabolites in the liver were the monoester and its side-chain oxidation products. The same metabolites 

and phthalic acid were in testes. Fat contained the monoester and its oxidation products. Repeated 

exposures revealed similar metabolites in the tissues. In summary, in the rat, DINP was de-esterfied 

to the monoester, which was further metabolized by side-chain oxidation of the ester group or by 

hydrolysis to phthalic acid. Formation of oxidation products appeared to increase following the high 

dose or repeated dosing, while the hydrolysis to phthalic acid decreased (28).

Distribution

In albino rats receiving 0.5 mL of 14C-DINP (approximately 2,500 mg/kg bw as estimated by the 

Expert Panel) after 5 days of dosing with the same amount of unlabeled DINP (27), no tissue studied had over 0.001% per gram of the administered dose after 3 days. The liver contained the most 

radioactivity on a total tissue basis. In male and female Fischer 344 rats receiving single or repeated 

oral doses of 14C-DINP (28), radioactivity also cleared from the tissues rapidly, but analysis of tissues soon (within 1 hour) after the exposure indicated that the highest levels were in liver (4.7% of 

administered dose), kidneys (0.31%), and blood (1.62 %). Fat and testes contained small amounts 

of metabolites. No bioaccumulation occurred over 72 hours postdosing.

Excretion

The major routes of excretion for orally administered DINP in rats were urine and feces, with about 

equal amounts excreted by either route at low doses, but more excreted in feces at high doses (28). 

Repeated dosing caused no accumulation of DINP or its metabolites in blood or tissue, but resulted 

in increased formation and elimination of the monoester side-chain oxidation products (28).

Side Chain-associated Toxicokinetics

A major metabolite of DINP, the monoester, MINP, is further oxidized in the side chain. 

2.3 Genetic Toxicity

DINP was tested in the Ames assay, Chinese hamster ovary (CHO) cells for chromosomal aberrations, the mouse lymphoma forward mutation assay (L5178Y TK -/- cell line), the primary rat hepa-
II-14

Appendix II

tocyte unscheduled DNA synthesis assay, and in an in vitro transformation assay using clone 1–13 

of Balb/c-3T3 A31 mouse cells. Where appropriate, exogenous metabolic activation systems were 

used. Many of the assays were conducted according to GLP standards (29). Based on the results of 

these studies, DINP is not considered mutagenic in bacterial mutation assays and mammalian gene 

assays and is not clastogenic in one cytogenetic assay in vitro with CHO cells and in one in vivo

assay with bone marrow cells of Fischer rats. This suggests that DINP is not genotoxic in vivo or in 

vitro (29)

Cell transformation studies give various results. The experimental conditions in the assays were not 

quite identical and the results are not inconsistent. Such positive results are in accord with those of 

well known peroxisome proliferators (29). DINP tested negative in the L5178Y mouse lymphoma 

mutation assay and the Balb/3T3 cell transformation assay (30). The data from the mutation and 

cell transformation assay were reviewed by OECD.

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

located in Section 5.1.2.
Appendix II

II-15

3.0 DEVELOPMENTAL TOXICITY DATA

3.1 Human Data

There were no human data located for Expert Panel review.

3.2 Experimental Animal Data

Two rat studies evaluating prenatal developmental toxicity of DINP by gavage were reviewed, as 

were the developmental toxicity aspects of a two-generation study in rats. Prenatal developmental 

toxicity of isononyl alcohol, a primary metabolite, was also evaluated.

Using Sprague Dawley rats, Waterman et al. (31) (Table 7-7) evaluated DINP-1 (CAS No. 68515-

48-0) and diisodecyl phthalate (DIDP) in 24 rats/group gavage treated with 0, 100, 500, and 1,000

mg/kg bw/day on gd 6−15. For both compounds, maternal toxicity was observed at 1,000 mg/kg 

bw/day expressed as reduced weight gain and food consumption. Fetal results were presented in 

terms of affected litters and fetuses. Skeletal variations were observed at the 500 and 1,000 mg/kg 

bw/day treatment levels. A dose-related increase in litters with lumbar ribs (25.0, 20.2, 54.2, and 

78.3%) was observed, which was statistically significant at the high dose. A dose-related increase 

in the percent of fetuses with rudimentary lumbar ribs was observed (3.7, 5.4, 18.6, and 34.5%) 

with statistical significance attained in the mid- and high-dose groups. The percent of fetuses with 

supernumerary cervical ribs was statistically increased only in the high dose group (1.6, 1.6, 1.0, 

and 5.7%), but the 2.5-fold increase in litter incidence in the high-dose group was not statistically 

significant (12.5, 12.0, 8.3, and 30.4%). There was a dose-related increase in the percentage of 

litters with dilated renal pelves (0, 12.0, 16.7, and 26.1%) that attained statistical significance at 

the highest dose. The percentage of fetuses with dilated renal pelves was significantly increased at 

all treatment levels (0, 3.7, 4.0, and 4.5% at 0, 100, 500, and 1,000 mg/kg bw/day, respectively). 

The interpretation of results by Waterman et al. (31) included the maternal and developmental 

LOAEL of 1,000 mg/kg bw/day with a NOAEL of 500 mg/kg bw/day, with a conclusion that 

DINP “is not teratogenic or a selective developmental toxicant.” The Expert Panel agreed with the 

authors selection of a maternal NOAEL; however, the Panel concluded that fetal data indicated 

that developmental toxicity was present at 500 mg/kg bw/day. The Expert Panel communicated 

to the study sponsor that there were improved and more current approaches to the analysis of 

fetal incidence data. The sponsor reanalyzed the fetal incidence data of interest using the GEE 

approach (32). This is a pup-level analysis within a model that uses the generalized estimating 

equation approach to account for the litter effect, i.e., the correlation between outcomes measured 

on pups within the same litter. The dose groups were tested pairwise versus controls; this gave 

similar results to a trend test based on a dose-response model fit with all dose levels up to that of 

interest included. These reanalysis data (33), presented below, are consistent with the Expert Panel’s 

interpretation of the vertebral data. 
II-16

Appendix II

Table 5: Mean Percent of Pups in Litter with Effect of Interest 

(significance level)

Dose Group (mg/kg bw/day DINP)

0 100 500 1,000

Skeletal Variations 16.4 15.0

(0.91)

28.3*

(0.05)

43.4**

(0.001)

Visceral Variations 0.5 3.3

(0.08)

3.7

(0.09)

5.8*

(0.04)

Rudimentary Lumbar Ribs 3.5 4.7

(0.18)

18.1**

(0.001)

34.2**

(0.001)

Supernumerary Cervical Ribs 1.6 1.5

(0.81)

1.0

(0.64)

5.5*

(0.05)

*p≤0.05, **p≤0.01

The GEE methodology could not be used to test the dilated renal pelves data because of the zero 

incidence in the control. Two solutions were tried. First, the zero incidence in the control was altered 

by changing one pup to “affected.” Second, an alternative statistical analysis that considers litter effects 

was used (34). The results of these statistical tests show reasonable agreement and are shown below.

Table 6: Mean Percent of Pups in Litter with Dilated Renal Pelves

(significance level using two methods)

Dose Group (mg/kg bw/day DINP)

0 100 500 1,000

Renal Pelves 0.0 3.3 3.7 5.3

 Sig. using added control event

 Sig. using nested analysis

(0.06)

(0.18)

(0.10)

(0.14)

(0.05)*

(0.04)*

These results diminish the Expert Panel’s initial concern that developmental toxicity effects based 

on dilated renal pelves may extend to lower doses. The Panel now concludes that the 100 mg/kg 

bw/day dose is a NOAEL.

Using the model-fitting approach, the sponsor also calculated benchmark doses (BMDs) at the 5 

and 10% excess risk level, based on a multiplicative (or ‘extra’) excess risk function, for the rudimentary lumbar rib variant. At the 5% excess risk level, the BMD05 (and 95% lower confidence 

interval, estimated by a bootstrap approach) was 193 mg/kg bw/day (162 mg/kg bw/day). Benchmark doses were not calculated for other variants. 

Hellwig et al. (35) (Table 7-6) evaluated the comparative developmental toxicity of a number of 

phthalates including three separate DINP materials. The material with CAS RN 68515-48-0 was 

identified as DINP-1, and two materials with CAS RN 28553-12-0, but from different production 

lines, were identified as DINP-2 and DINP-3. See Section 1.1 for description of chemical differences. 

Each DINP was administered by gavage in olive oil at 0, 40, 200, and 1,000 mg/kg bw/day to 8−10 

sperm-positive Wistar females/group on gd 6–15. The dams were killed on gd 20 and implantation 
Appendix II

II-17

sites were examined. Fetuses were weighed and examined for external malformations; half of the 

fetuses were examined for skeletal malformations and the other half for visceral malformations. 

For DINP-1, maternal toxicity at the high dose consisted of reduced food consumption and 

increased relative liver (~6%; not statistically significant) and kidney (~13%) weights. There were 

no treatment-related effects on the number of live fetuses/dam or fetal weight. Developmental toxicity was evident at the highest dose by a statistically significant increase in percent fetuses/litter 

with variations (35.3, 41.5, 29.5, and 58.4% in the 0, 40, 200, and 1,000 mg/kg bw/day groups, 

respectively). These variations consisted of rudimentary cervical and/or accessory 14th rib(s). A 

modest increase in dilated renal pelves in the high dose group was also noted by the Expert Panel 

(8.9% of fetuses in 78% of control litters versus 16.8% of fetuses in 90% of treated litters). There 

were no maternal or developmental effects at 40 or 200 mg/kg bw/day. 

For DINP-2, there was no statistically significant, dose-related evidence of maternal toxicity. 

However, a non-significant increase in relative liver (~5%) and kidney (~7%) weight did occur. 

The authors stated that developmental toxicity effects were limited to an increased fetal incidence 

of accessory 14th lumbar ribs at the high dose. The Expert Panel also noted a modest increase in 

dilated renal pelves in the high-dose group (8.9% of fetuses in 78% of control litters versus 10.6% 

of fetuses in 80% of treated litters).

For DINP-3, maternal toxicity was present at the high dose, expressed as reduced mean body 

weight gain and reduced food consumption during some portions of the treatment period. Relative 

liver weights (~11%) were also increased at the high dose and a non-significant increase in relative kidney weights (~9%) was observed. Developmental toxicity was evidenced by a statistically 

significant increase in percent fetuses/litter with variations at the highest dose (35.3, 29.6, 39.5, and 

60.7% in the 0, 40, 200, and 1,000 mg/kg bw/day groups, respectively). Specific types of developmental toxicity observed in the high-dose group at increased incidences included skeletal retardation (unossified or incompletely ossified sternebrae) and skeletal variations (rudimentary cervical 

and/or accessory 14th rib[s]). The authors assumed that low incidences of soft tissue variations 

(hydroureter), visceral malformations affecting the urogenital tract (agenesis of kidneys and ureters), and skeletal malformations affecting the long bones (shortened and bent humerus and femur) 

observed at the high dose were treatment-related. The Expert Panel also observed that dilated renal 

pelves were slightly increased in the high dose group (8.9% of fetuses in 78% of control litters 

versus 16.7% of fetuses in 100% of treated litters). A maternal and developmental NOAEL of 200 

and LOAEL of 1,000 mg/kg bw/day, respectively, were identified by the Expert Panel and were in 

concurrence with effect levels identified by authors. 

In a two-generation reproductive toxicity study, postnatal weight gain was examined in pups of F0 

and F1 dams exposed to DINP in feed at concentrations of 0, 0.2, 0.4, and 0.8% during mating (0, 

182–197, 356–397, and 696–802 mg/kg bw/day) , gestation (0, 143–146, 287–288, 555–560 mg/kg 

bw/day), and lactation (0, 254–285, 539–553, and 1,026–1,129 mg/kg bw/day) (2, 36). Complete 

details of the experiment are included under Section 4. Weight gain for the F1 pups was reduced 

by DINP; males of the high-dose group were affected on postnatal day (pnd) 0, pups of the midand high-dose groups were affected on pnd 7 and 14 and all dose groups were affected by pnd 21. 

Weight gain of the F2 young during lactation was reduced in primarily the mid- and high-dose 
II-18

Appendix II

groups; females of the low-dose group were affected on pnd 7, females of the mid- and high-dose 

groups were affected on pnd 4, 7, 14, and 21, and males of the mid- and high-dose groups were 

affected on pnd 7, 14, and 21. Postnatal sexual maturation was not examined. The Expert Panel 

identified a developmental NOAEL of 0.2%. 

Hellwig and Jackh (37) evaluated the prenatal development toxicity of two types of isononyl alcohol in Wistar rats. The type 1 alcohol consisted of isomers with a medium degree of branching and 

16% isodecanol and the type 2 alcohol consisted of isomers with a low degree of branching. On gd 

6–15, 10 rats/group were gavaged with the alcohols in water with 0.005% Cremophor EL at concentrations of 0, 1, 5, and 10 mmol/kg bw/day which the authors stated equated to 0, 144, 720, and 

1,440 mg/kg bw/day. A supplementary study was later conducted in 10 rats/group exposed to 0 or 

type 1 or type 2 alcohol at 7.5 mmol/kg bw/day (~1,080 mg/kg bw/day). In the main and supplementary studies, two groups of 10 control rats each were administered water or vehicle. Fetuses and 

dams were evaluated on pnd 20. 

For the type 1 isononyl alcohol, complete maternal lethality occurred in the 1,440 mg/kg bw/day 

group and 1 of 10 dams died at the 1,080 mg/kg bw/day dose. Clinical signs/symptoms were 

reported to have been observed in a dose-related manner in dams that received 720 mg/kg bw/day 

and the two higher doses. A significant reduction in maternal body weight gain and increased fetal 

resorptions were observed in the 1,080 mg/kg bw/day group. Numerical reductions in fetal body 

weight occurred in the 720 mg/kg bw/day group, but were not statistically significant. Malformations that primarily affected the heart were significantly increased in fetuses and litters of the 1,080 

mg/kg bw/day group. Skeletal variations (cervical ribs) or retardations (reduced ossification of sternebrae) were increased in the 1,080 mg/kg bw/day group (statistically significant) and 720 mg/kg 

bw/day group.

Maternal mortality was also observed in the 1,440 mg/kg bw/day group treated with type 2 isononyl alcohol with death occurring in 3/10 dams. Maternal signs and symptoms were observed in 

the three highest doses. Non-significant reductions in body weight gain and marginal increases 

in resorption rates were observed in dams exposed to 720 mg/kg bw/day and higher. Fetal body 

weights were significantly reduced in the 1,440 mg/kg bw/day group. The authors reported that 

malformations were not significantly increased, but also reported that there were significant 

increases in fetuses with skeletal variations and retardations (reduced ossification) from the highdose group (1,440 mg/kg bw/day). It is not clear which type of variation was increased. The authors 

stated that the number of fetuses with malformations (primarily affecting the thoracic vertebrae) 

was elevated in the 1,080 mg/kg bw/day group.

Data available in abstract form (38) and a study in press (39) report that oral exposure of SD rats to 

DINP at 750 mg/kg bw/day on gd 14 through pnd 3 resulted in reproductive malformations in male 

offspring (7.7%). The data were not available to the Panel for evaluation, therefore we merely note 

the existence of the abstract.

The summary for Section 3 is located in Section 5.1.3.
Appendix II

II-19

4.0 REPRODUCTIVE TOXICITY 

4.1 Human Data

There were no human data located for Expert Panel review.

4.2 Experimental Animal Toxicity

The reproductive toxicity of DINP-1 (CAS 68515-48-0) was reported by Waterman et al. (36) 

(Table 7-8). This report describes the results of both a one-generation and a two-generation study. 

In the two-generation study, SD rats (30/group) were given DINP in the diet at 0.2, 0.4, or 0.8% 

(w/w) for 10 weeks prior to mating, and through gestation and lactation. The study sponsor estimated doses of 0, 165, 331, and 665 mg/kg bw/day for males during premating, 0, 182, 356, and 

696 mg/kg bw/day for females during premating, 0, 146, 287, and 555 mg/kg bw/day for females 

during gestation, and 0, 254, 539, and 1,026 mg/kg bw/day for females during lactation (2). Body 

weights and food consumption were measured weekly. After 10 weeks of pre-mating exposure, 

males and females were paired 1:1 within dose groups, and the females were monitored for vaginal 

sperm for up to 3 weeks. F0 males were treated until after the delivery of their last litter and then 

killed and necropsied; females were killed after weaning their litters. Litters were culled to yield 

four male and four female pups per litter on pnd 4. At weaning, one male and one female from each 

litter were selected to grow to adulthood for mating; the remaining animals were examined externally, then killed and discarded without necropsy. F1 animals were fed the same diet as their parents 

throughout the rest of the study. As adults, the F1 rats were mated within dose groups for 3 weeks 

after confirmation of vaginal sperm. Estimated doses for the F1 rats were 0, 189, 379, and 779 

mg/kg bw/day for males during premating, 0, 197, 397, and 802 mg/kg bw/day for females during 

premating, 0, 143, 288, and 560 mg/kg bw/day for females during gestation, and 0, 285, 553, and 

1,129 mg/kg bw/day for females during lactation (2). Dams were allowed to litter and raise young 

until pnd 1, at which time they and their litters were killed, the adults were necropsied, and their 

organs weighed and preserved. Sperm measures were not made; testes were fixed in Bouin’s. 

Weight gain for the F0 rats was unaffected by DINP consumption until pnd 14 and 21, when the 

high-dose dams weighed less than the controls. In the F0 rats, absolute liver weight was increased in 

the females of the mid dose group, and in both males and females of the high-dose group. Absolute 

kidney weight was increased in the two highest dose groups for males and in all exposed female 

groups. Absolute reproductive organ weights (testes, epididymis, prostate, and seminal vesicles) 

were unchanged by DINP. At the high dose, absolute left ovary weight was reduced versus control, 

although the weight of the right ovary was unchanged; the reduced weight of the left ovary appears 

anomalous. Fertility indices for the F0 mating were unchanged by DINP; this includes litter size, 

measures of mating, number of dead offspring, and sex ratio. Weight gain was reduced in both F1

and F2 pups and these results are discussed in detail in Section 3. Body weights during the mating of the F1 generation were variably reduced at the high dose by ~8−10%. When the F1 animals 

were mated within dose groups as adults, DINP caused no change in ability to mate or bear young, 

litter size, pup weight or viability, or sex ratio. At the F1 adult necropsy, absolute liver weight was 

increased in the high-dose females and absolute kidney weight was increased in the high-dose 

males. Absolute reproductive organ weights were unchanged. Livers appeared more eosinophilic in 

all treated F0 and F1 rats; kidneys of the mid- and high-dose males had minimal-to-mild pelvic dilation. Testes were microscopically equivalent to controls. 

 
II-20

Appendix II

The NOAEL for reproduction appears to be at least 0.8%, 665–779 mg/kg bw/day for males and 

696–802 mg/kg bw/day in females. There is no LOAEL for reproduction, as there were no reproductive toxicities observed. The weight gain inhibition at 0.2% seen by pnd 21 in F1 pups suggests a 

developmental LOAEL of 143–285 mg/kg bw/day during gestation through lactation. Using benchmark dose methodology (BMD) the authors reported that 250 mg/kg bw/day represented the 95% 

lower confidence limit for a 5% reduction in body weight. 

In the one-generation study, groups of 30 male or female animals consumed DINP-1 in the feed at 

0, 0.5,1.0, or 1.5% w/w for 10 weeks prior to mating (36). Study sponsor-estimated doses for males 

during the premating period were reported at 301–591, 622–1,157, and 966–1,676 mg/kg bw/day 

in the low- to high-dose groups respectively (2). Doses for the low to high dose females were 363–

624, 734–1,169, and 1,114–1,694 mg/kg bw/day during premating, 377–404, 741–796, and 1,087–

1,186 mg/kg bw/day during gestation, and 490–923, 1,034–1,731, and 1,274–2,246 mg/kg bw/day 

during lactation. The females were exposed throughout mating, gestation, and lactation until pnd 

21. The males were killed immediately after the mating period. At necropsy, the liver, kidneys, and 

reproductive organs were removed and weighed. 

In this one-generation study, body weight gain was reduced at 1 and 1.5% DINP. There were no 

effects on indices of mating or fertility (litter size), and a reproductive NOAEL of 1,000 mg/kg bw/

day was identified. At necropsy, absolute liver and kidney weights were increased in both sexes at 

all dose levels. Testes absolute weights were increased at the high dose; ovary weights were reduced 

by ~30% at the highest dose. Offspring viability was reduced in the high-dose group. Offspring 

body weight gain at pnd 21 was reduced at all dose levels, as in the two-generation study. 

Mode 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 

cytosol (40), rainbow trout hepatic cytosol (41), recombinant human ERs (rhER) overexpressed in 

SF9 insect cells using the baculovirus system (42, 43) and rainbow trout ERs expressed in yeast 

(44). Tritiated E2 was used in the tissue cytosol binding assays while a high affinity fluorescent E2 

derivative was used in the rhER binding assays. Selected phthalate esters have been examined in 

a number of in vitro gene expression assays systems. The assays have used stably transfected cells 

(40), transiently transfected cells (40, 41), yeast based assays (40, 44-46) and vitellogenin induction in rainbow trout hepatocyte cultures (44). DINP exhibited no activity in an in vitro assay that 

measured binding of phthalates to estrogen receptors (40) and in an assay of estrogen-induced 

gene expression (46). The assays did not include the addition of esterases or lipases to metabolize 

DINP to MINP. In vivo assays demonstrated that DINP does not increase uterine wet weight or 

vaginal epithelial cell cornification in immature or mature ovariectomized rats treated with up to 

2,000 mg/kg bw for 4 days. (40). There were no studies located on anti-androgenic activity, but an 

abstract and study in press have reported that gestational DINP exposure demasculinizes male pups 

(38, 39). Thyroid and estrogen serum levels were unaffected in adult marmosets at doses as high as 

2,500 mg/kg bw/day for 13 weeks (20).

The summary for Section 4 is located in Section 5.1.4.
Appendix II

II-21

5.0 DATA SUMMARY & INTEGRATION

5.1 Summary

5.1.1 Human Exposure

DINP, a complex substance of branched, predominantly C-9 isomers, is a general-purpose 

plasticizer for flexible PVC with a broad range of applications. It is widely used in the toy, 

construction, and general consumer product markets. It has limited use in food packaging. It is not 

used for medical applications.

The very limited monitoring data for DINP in air, drinking water, and surface and ground waters 

have usually yielded negative results (i.e., concentrations below detection limits). In the few 

studies of food and infant formula, the levels of DINP have not been quantitated or have been at 

or below the detection limit (0.01 mg/kg) (4-6). Occupational exposures to phthalates are reported 

to be below 1 mg/m3 during the production of phthalates and 2 mg/m3 during the manufacture of 

plasticized PVC (1). 

Toys represent a unique childhood exposure to DINP since it is a major plasticizer used in children’s toys (7, 8, 11). DINP content has been measured at 15.1–54.4% dry weight in 31 toys (7), 

and 3.9–44% dry weight in 27/42 toys (11). Using pneumatic piston impaction in saliva simulant, 

DINP migration ranged from 1.0–48.4 μg/11cm2/hour, but there was no correlation between DINP 

content and migration rate (7). In vivo extraction has been studied using adult volunteers as surrogates for children (7, 12). In a comparison of extraction rates in 10 adults mouthing toys versus laboratory simulation, ratios varied from 22.9 to 72.6 (mean 39.5) for 5 toys (7). RIVM (12) 

tested two different pieces of one toy and a controlled disk using 20 adults and also found higher 

extraction in vivo. Using a 2-day parent observation study of 42 children, ages 3–36 months, mean 

mouthing times have been generated per age category with ranges from 0 minutes/day in older children to 171.5 minutes/day in the 6–12 month age group (see Section 1). These mouthing times have 

been used to model DINP exposure by several groups using a variety of assumptions as indicated in 

Table 7. Dermal exposure may also occur, but has not been studied specifically in children.

Table 7: Toy Exposure Estimates for Children Aged 3−12 Months

Estimated Intake Level (μg/kg bw/day)

Agency Mean 95th Percentile 99th Percentile Maximum

RIVM* 6.53–14.4 20.7–39.7 39.8–77.3 70.7–204

CPSC 5.7 94.3 – –

Health Canada** 44 73.9*** 173.5*** 320

* Exposure range for 3−6 month-old and 6−12 month-old children; 

 range includes results from 3 specimens tested.

** Calculated with mouthing times for teethers and other objects intended for mouthing.

*** Results using Monte Carlo simulations in children aged 3–6 months.
II-22

Appendix II

5.1.1.1 Utility of Data to the CERHR Evaluation

The Expert Panel believes it is reasonable to assume, based on the physicochemical characteristics 

of DINP and existing, though limited monitoring data, that general population exposure to DINP 

(excluding children) is expected to be lower than DEHP, which is estimated at 3–30 μg/kg bw/day 

(15). Children may be in the upper portion of the population range because of their physiologic 

differences compared to adults. The Panel also believes that some small children are likely to have 

exposures exceeding the general population estimates due to non-dietary ingestion from mouthing 

toys and other objects that contain DINP. Current models of non-dietary oral exposure predict that 

older infants and toddlers may incur exposures up to one order of magnitude higher than the upper 

limit of exposure expected in the general population.

5.1.2 General Biological and Toxicological Data

General Toxicity:

There were no human data identified. Animal data consisted of two subhuman primate studies and 

four rodent studies. In a 13-week gavage study, adult marmosets treated with up to 2,500 mg DINP/

kg bw/day (CAS number not specified) experienced weight loss or decreased weight gain, but 

there was no biochemical evidence of peroxisome proliferation or microscopic changes in organs 

examined, including testes and epididymides (20). Prepubertal (2-year-old) cynomolgus monkeys 

that were gavaged with 500 mg/kg bw/day for 2 weeks experienced changes in white blood cell 

numbers, but there were no testicular lesions or hepatic effects, including peroxisome proliferation 

(21). A 21-day repeat-dose dietary study in adult rats focused on peroxisome-proliferating effects 

in liver, and a LOAEL of 607 (F) and 639 (M) mg/kg bw/day was identified; a NOAEL was not 

established (16). Effects included increased liver weight at all dose levels in males and in females, 

dose-related enzymatic evidence of peroxisome proliferation, and alterations in hepatic cytoplasmic basophilia and eosinophilia at the high dose. With the exception of severe unilateral atrophy 

in one male of the mid-dose group, testicular effects were not observed in males dosed with up to 

2,195 mg/kg bw/day. Moderate testicular atrophy was observed in one DEHP-positive control that 

received 1,084 mg/kg bw/day.

There were three chronic (2-year) dietary studies reviewed that were of similar design and included 

toxicopathologic evaluation at several times during the study. Two studies were conducted using 

6-week-old F 344 rats (17, 18), while the third used 6-week-old B6C3F1 mice (19). Lesions in 

testes or female reproductive organs were not observed in any of the 3 studies, with the highest 

doses tested being 885 mg/kg bw/day in rats and 1,888 mg/kg bw/day in mice. Non-neoplastic liver 

lesions and/or changes in liver enzyme activity occurred at doses of 152 mg/kg bw/day and greater 

in rats, and 1,560 (M) to 1,888 (F) mg/kg bw/day in mice. Biochemical evidence of peroxisome 

proliferation was noted throughout the study in both sexes of rats in the Moore (18) study that were 

dosed with 733 (M) and 885 (F) mg/kg bw/day. Female rats receiving 442 mg/kg also had biochemical evidence of peroxisome proliferation when evaluated at the end of the study. Peroxisome 

proliferation was noted in high-dose mice (1,560 [M]; 1,888 [F] mg/kg bw/day), but the mid- and 

low-dose groups were not examined. The Lington et al. (17) rat study evaluated peroxisome proliferation by electron microscopy and saw none in two rats per sex per dose group at the end of the 

study. Non-neoplastic kidney lesions and changes in urinary excretion were seen in rats exposed 

to 307 mg/kg bw/day and higher, and in mice dosed with 1,560 (M) and 1,888 (F) mg/kg bw/day. 
Appendix II

II-23

Indications of anemia, such as reductions in red blood cell numbers and hemoglobin levels, were 

seen in rats exposed to 307 mg/kg bw/day and higher. Hepatic neoplasia was observed only in male 

rats exposed to 733 mg/kg bw/day and in mice exposed to 336 (F) and 742 (M) mg/kg bw/day and 

higher. Renal neoplasia was only observed in male rats of the highest dose group (733 mg/kg bw/

day). The apparent qualitative difference in liver and renal effects (i.e., tumors vs hepatotoxicity) in 

the rat studies may reflect differences in the range of doses tested.

There were no toxicity studies with inhalation exposure.

Mode of Action:

The renal neoplasia in male rats appears to be due to alpha-2-microglobulin nephropathy which is 

a mechanism not considered relevant to humans (23). However, an increased rate of nephropathy, 

was seen in female mice exposed to 1,888 mg/kg bw/day which would not be consistent with the 

alpha-2-microglobulin mechanism. The Moore (18) study demonstrated liver tumors in rats only in 

the highest-dose males. Peroxisome proliferation in rats was observed at the highest dose in males 

and females, and the second highest dose in females but not males. No liver tumors were observed 

in either sex at the second-highest dose level. In addition, no liver tumors were noted in the recovery groups. These results are consistent with a peroxisome proliferation mode of action for hepatic 

tumor induction. Unfortunately, peroxisome proliferation was assayed in mice only at the highest 

dose, and liver tumors were observed at lower doses. 

Toxicokinetics. 

There are no human data. DINP was orally administered to adult male albino rats at doses of 

50, 150, or 500 mg/kg bw/day. It is metabolized by pancreatic lipases in the lumen of the gut 

and rapidly absorbed (49%) as the monoester and rapidly excreted via urine and feces with no 

accumulation in tissues (28). Dermal absorption of DINP is slow (<4% in 7 days) in rats (24). 

Dermal absorption of DINP through human skin is expected to be lower than rat skin based on 

results of an in vitro study conducted with DEHP (47). There is evidence for excretion via the 

biliary route based on radioactivity in feces and GI tract of rats dosed dermally with 14C-DINP. 

There are no inhalation studies available. 

Genetic Toxicity. 

DINP tested negative in experiments of mutagenicity and clastogenicity including the Ames, 

Chinese hamster ovary cell, and rat bone marrow chromosomal aberration, mouse lymphoma 

mutation, unscheduled DNA synthesis, and Balb/c-3T3 mouse cell transformation assays (29, 30).
II-24

Appendix II

Table 8: Summaries of NOAELs and LOAELs and Major Effects in General Toxicity Studies

Protocol & Study

DINP Tested and Doses 

(mg/kg bw/day)

NOAEL

(mg/kg bw/

day)

LOAEL (mg/kg bw/day)

 & Effects

Major Effects at 

Higher Doses

13-week repeat-dose gavage study in adult 

marmosets

16–25 months of age, 1–2 per sex/group 

Doses: 0, 100, 500, 2,500

DINP type not specified

(20)

500 2,500

↓Weight gain or weight loss

No peroxisomal proliferation

No microscopic findings in 

organs

No higher doses in study

2-week repeat-dose gavage study in male 

prepubescent cynomolgus monkeys

2 years of age, 4/group

Doses: 0, 500

DINP-1

(21)

Not 

determined

500

Changes in neutrophil and 

lymphocyte counts

No testicular lesions

No liver effects including 

peroxisomal proliferation

No higher doses in study

21-day repeat-dose dietary study in young 

adult Fischer 344 rats

6 weeks of age at start of study, 

5 rats per sex/group

Doses: (M) 0, 639, 1,192, 2,195;

(F) 0, 607, 1,193, 2,289

Mixture of different DINP types

(16)

None M: 639, F: 607

↑ Liver weight 

↑ Peroxisomal proliferation 

(M)

↑ Kidney weight

↑ Liver weight, and peroxisomal 

proliferation

↑ Kidney weight

↑Testes weight

No testicular lesions

2-year repeat-dose dietary study in 

Fischer 344 rats 

6-week-old at beginning of study, 

110 per sex/group

Doses: (M) 0, 15, 152, 307;

 (F) 0, 18, 184, 375

DINP-1

(17)

M: 15

F: 18

M: 152, F: 184 

Hepatic effects

↑Liver weight

Mononuclear cell leukemia

↑Kidney weight

Hepatotoxicity

↑Liver weight

No testicular lesions

Mononuclear cell leukemia

Anemia

↑Kidney weight and excretion 

changes

No peroxisomal proliferation

2-year repeat-dose dietary study in Fischer 

344 rats 

6 weeks of age at start of study, 70–85 per 

sex/group 

Doses: (M) 0, 29, 88, 359, 733;

(F) 0, 36, 109, 442, 885

DINP-1 

(18)

M: 88 

F: 109

M: 359, F: 442 

Nephrotoxicity 

Excretion changes

Anemia 

↑ Liver weight, peroxisomal 

proliferation (F)

Mononuclear cell leukemia

↑Kidney weight

Hepatic & renal neoplasia at high 

dose (M) 

Anemia

Nephrotoxicity

↑ Liver weight and peroxisomal 

proliferation

Mononuclear cell leukemia

↑Kidney weight

No testicular lesions

2-year repeat-dose dietary study in 

B6C3F1 mice 

6 weeks of age at beginning of study, 70/

sex/group 

Dose: (M) 0, 90, 276, 742 or 1,560;

(F) 0, 112, 336, 910, 1,888

DINP-1 

(19)

M: 276 

F: 112

M: 742, F: 336

Liver neoplasia

↑Liver weight (M)

↓Kidney weight (M)

Liver neoplasia, hepatocyte 

staining variations, 

peroxisomal proliferation, and 

nephrotoxicity (F) at highest 

doses

↑Liver weight 

↓Kidney weight (M)

No testicular lesions
Appendix II

II-25

5.1.2.1 Utility of Data to the CERHR Evaluation 

There are adequate subchronic and chronic data available in rats and mice and adequate subchronic 

data in primates to assess general toxicity by the oral route, including liver and kidney effects (16-

19). No effects have been noted in the male or female reproductive system, although these studies 

were not designed to fully assess this system. 

Toxicokinetic data consist of oral and dermal studies in rodents. The data permit the Panel’s 

conclusion that dermal absorption is slow; oral absorption is rapid for the monoester formed by 

lipases in the gut. Dose-related kinetics of absorption across species is not known. DINP and its 

metabolites are rapidly excreted via urine and feces with no accumulation in tissues.

5.1.3 Developmental Toxicity

There were no human studies located for Expert Panel review.

Two published prenatal developmental toxicity studies in rats were available for DINP (31, 35). The 

protocols for the two studies were similar and included dosing of dams by gavage on gd 6–15 with 

sacrifice and evaluation of fetuses on gd 20–21. Developmental toxicity was also noted in both a 

one-generation and a two-generation toxicity study. The effects on pup body weight are discussed 

below and summarized in Table 9; the reproductive effects are described in Section 5.1.4. 

Hellwig et al. (35) performed their studies in Wistar rats (10/group) at doses of 0, 40, 200, and 

1,000 mg/kg bw/day. Although sample size (n=10) was small, the aggregate of their work can 

logically be considered to be three separate studies of DINP. There was a degree of consistency 

across all studies. Effects were only observed at the highest dose. Relative kidney and liver weights 

were slightly increased in dams of the highest dose group (5–13%), but statistical significance was 

erratic. Fetal viability and body weight were unaffected in all three studies. Skeletal variations 

(rudimentary cervical ribs, accessory 14th ribs) were numerically increased with each DINP with 

the number of affected fetuses per litter significantly higher than controls in two instances. There 

was a tendency to see dilated renal pelves at the highest dose; in one study agenesis of kidneys and 

ureters was assumed by the authors to be DINP-related. Skeletal (shortened and bent humerus and 

femur) malformations were also observed in the high-dose group of this study. It is clear that organ 

effects are associated with kidney and the skeletal system. For maternal and developmental effects, 

a NOAEL of 200 and a LOAEL of 1,000 mg/kg bw/day were identified by the Expert Panel for 

each DINP and are in concordance with effect levels identified by Hellwig et al. (35). 

The prenatal toxicity study of Waterman et al. (31) was more informative than the Hellwig study 

from the standpoint of number of rats per test group and completeness of data reported. Waterman 

et al. (31) tested DINP-1 in Sprague-Dawley rats (25/group) at doses of 0, 100, 500, or 1,000 mg/kg 

bw/day. Maternal toxicity at the highest dose consisted of decreased food intake and weight gain. 

The authors presented and analyzed effects on offspring as percent affected fetuses and percent 

affected litters. Waterman et al. (31) interpreted their results as indicating a LOAEL for maternal 

and developmental toxicity at 1,000 mg/kg bw/day and a NOAEL of 500 mg/kg bw/day. The Panel 

concurred with the maternal NOAEL, but concluded there was developmental toxicity at the 500 

mg/kg bw/day dose. As discussed in Section 3.2, the Panel advised the study sponsor that there 

were more recent and improved methods for the statistical analysis of fetal incidence data. The 
II-26

Appendix II

sponsor performed appropriate reanalyses that the Panel reviewed and found to be consistent with 

the Panel interpretation of skeletal variations. The Panel concludes there is a NOAEL in the study 

at 100 mg/kg bw/day. The BMD estimated a 5% excess risk level was 193 mg/kg bw/day (95% 

LCL=162 mg/kg bw/day) for rudimentary lumbar ribs, as provided by the study sponsor (33).

The Panel noted that developmental toxicity was observed in the prenatal rat studies by Waterman 

and Hellwig. In the study by Waterman, the urinary system was a target of effect as noted by a 

modest increase in dilated renal pelves at the 1,000 mg/kg dose. While only a mild increase in 

dilated renal pelves was observed in the three Hellwig et al. studies, in one instance more severe 

renal effects (hydroureter, agenesis) were seen. In studies by Waterman et al. (31) and Hellwig et al. 

(35), the skeletal system was the target for effect as observed by an increased incidence of cervical 

ribs and accessory 14th (lumbar) ribs. These studies also evaluated the closely related phthalate 

DIDP where the same target organs were identified. An increase in cervical ribs and lumbar ribs 

was observed at the common dose of 1,000 mg/kg bw/day in the two studies. While effect on 

lumbar ribs was more pronounced, the effect on cervical ribs is of greater toxicological concern. 

Cervical ribs are seen infrequently in controls, and their presence may indicate a disruption of gene 

expression. There is evidence that cervical ribs may interfere with normal nerve function and blood 

flow.

Differences in NOAELs between the Waterman et al. (31) and Hellwig et al. (35) studies, 100 and 

200 mg/kg bw/day respectively, may be due to rat strain, and certainly to dose selection. 

The two-generation reproductive study by Waterman et al. (36) suggests an adverse effect on 

weight gain in pups during the perinatal and pre-weaning period of life. Developmental landmarks 

of reproductive tract development, identified as a sensitive target with other phthalates, were not 

examined. F1 mean pup body weight was significantly reduced on pnd 0 in males at 0.8% DINP 

(555 and 1,026 mg/kg bw/day during gestation and lactation, respectively, as calculated by study 

sponsors). On pnd 7 and 14, mean male and female pup body weights were significantly reduced 

at 0.4% (287 and 539 mg/kg bw/day during gestation and lactation, respectively) and 0.8%, and by 

pnd 21, mean male and female body weights were reduced at all dose levels. In the F2 generation, 

mean female pup body weights were significantly reduced at 0.4 and 0.8% on pnd 4, 7, 14, and 

21 and at 0.2% (143 and 285 mg/kg bw/day during gestation and lactation, respectively) at pnd 7. 

Mean male pup body weights were significantly reduced at 0.4 and 0.8% at pnd 7, 14, and 21. The 

LOAEL for developmental effects was therefore identified as 0.2% (143–285 mg/kg bw/day during 

gestation through lactation) by the Expert Panel.

Studies with 2 isononyl alcohols, differing in degree of branching, demonstrated clinical signs and 

symptoms in pregnant rats at doses of 720 mg/kg bw/day and higher (37). Table and text discrepancies in dose values and reported effects at the higher dose levels were noted. Toxicity was more 

severe with type 1 isononyl alcohol, the alcohol that had a higher degree of branching. Maternal 

mortality was seen at the highest dose (1,440 mg/kg bw/day) with both alcohols and in the type 1 

alcohol at 1,080 mg/kg bw/day. Fetal malformations and/or variations occurred at 1,440 mg/kg bw/

day and at 1,080 mg/kg bw/day. Slight effects that may be associated with treatment were observed 

at 720 mg/kg bw/day. A dose of 144 mg/kg bw/day was without effect for both isononyl alcohols.
Appendix II

II-27

Table 9: Summary of NOAELs and LOAELs and Major Effects in 

Developmental Toxicity Studies

Protocol & Study

NOAEL 

(mg/kg bw/day)

[Benchmark dose] 

ED05 in 

mg/kg bw/day

LOAEL (mg/kg bw/day) 

and Effects

Developmental 

Effects Observed at 

Higher Dose Levels

Maternal Developmental Developmental

Prenatal gavage study in 

Wistar rats. 

10/group/study received 0, 

40, 200, or 1,000 mg/kg bw/

day on gd 6–15.

Dams and pups examined in 

late gestation.

DINP-1, DINP-2, and 

DINP-3

(35)

200 

Maternal & 

Developmental

1,000

↑Kidney and liver 

weights.

1,000

↑Cervical and 

lumbar ribsall. 

↑Urogenital 

and skeletal 

malformation 

with DINP-3.

N/A

Prenatal gavage study in

Sprague-Dawley rats.

25 per group received 0, 

100, 500, or 1,000 mg/kg 

bw/day on gd 6–15.

Dams and pups examined in 

late gestation.

DINP-1

(31)

500 (Maternal)

100 ***

(Developmental)

[MLE(95%LCL):

193 (162) for 

lumbar ribs] 

1,000

↓Weight gain.

500

↑ Fetuses with 

vertebral 

variations.

↑ Fetuses and litters 

with visceral 

variations (mainly 

dilated renal 

pelves).

↑ Fetuses and litters 

with lumbar ribs.

↑ Fetuses with 

cervical ribs.

Two generation reproductive 

dietary study in SpragueDawley rats. 

30 per group were fed diets 

with 0, 0.2, 0.4, or 0.8% 

from 10 weeks prior to 

mating (0, 182–197, 356–

397, and 696–802 mg/kg 

bw/day) through gestation 

(0, 143–146, 287–288, and 

555–560 mg/kg bw/day and 

lactation 0, 254–285, 539–

553, and 1,026–1,129 mg/kg 

bw/day during lactation)**.

DINP-1

(36)*

None

[250 (95% LCL) 

for decreased pup 

weight gain]

143−285

↑ Mild 

histological 

liver changes 

in F0 and F1.

↑Kidney weight 

in F0.

143−285

↓ Weight gain on 

pnd 21in F1. 

↓ Weight gain 

on pnd 7 in F2 

females.

↓ Weight gain on pnd 

0 (males), 7, 14, 

and 21 in F1.

↓ Weight gain on pnd 

4 (female), 7, 14, 

and 21 in F2.

* Only maternal and developmental effects were listed in this table. Reproductive and male systemic effects are listed in Table 10.

** Range of doses for F1 and F2 dams.

*** NOAEL selected by Expert Panel is lower than study author’s selection
II-28

Appendix II

5.1.3.1 Utility of Data to the CERHR Evaluation

There are adequate data available in rats to determine that prenatal oral exposure to DINP-1 results 

in developmental toxicity. The results of the Waterman et al. (31) and the Hellwig et al. (35) studies 

were remarkably consistent with respect to DINP-1. In both studies, exposure to DINP-1 resulted 

in increases in lumbar and cervical ribs. In addition, the effective dose levels were similar. Hellwig 

et al. (35) identified a LOAEL of 1,000 mg/kg bw/day and a NOAEL of 200 mg/kg bw/day with 

a sample size of 10/group. The Panel identified an effect level of 500 mg/kg bw/day from the 

Waterman et al. (31) study (sample size of 25/group) and 100 mg/kg bw/day level represented a 

NOAEL. In addition, Hellwig et al. (35) showed some similarities among the three DINPs in that 

each resulted in an increase in lumbar and cervical ribs. It is clear that the urinary and skeletal 

systems are target organs where developmental toxicity is observed. The data from the twogeneration dietary study are sufficient to demonstrate an effect on postnatal growth, with a LOAEL 

of 143–285 mg/kg bw/day and no NOAEL. The reduced growth is consistent in both studies. 

Neither prenatal study extended dosing into the late gestation period which has been shown to be a 

critical window of development for other phthalates. In addition, the study designs did not allow for 

assessment of postnatal sexual maturation. The issue of late gestational exposure was addressed in 

a two-generation reproductive toxicity study reviewed in Section 5.1.4. Confidence in the isononyl 

alcohol study is limited due to table and text discrepancies in dose values and reported effects at the 

higher dose levels. The study is adequate to ascribe maternal and developmental toxicity at these 

higher doses and to assume the lowest dose was without effect.

5.1.4 Reproductive Toxicity

Structural and functional reproductive effects were examined in one- and two-generation feeding 

studies in rats that included in utero exposure during the entire duration of pregnancy (36). In the 

one-generation dose range finding study, rats were administered dietary levels of 0, 0.5, 1.0, or 

1.5% DINP and in the two-generation study, rats were administered dietary levels of 0, 0.2, 0.4, 

or 0.8% DINP. In the two-generation study, reproductive parameters including mating, fertility, 

and testicular histology were unaffected in both generations at the highest dose (0.8%; 665–779 

and 696–802 mg/kg bw/day in males and females, respectively) and this dose was identified as 

the reproductive NOAEL. Developmental effects were observed, including decreased pup weight 

gain (most marked on pnd 21). The effects on pup weight gain are discussed in greater detail under 

Section 5.1.3. Histologic effects included mild hepatic eosinophilia in both sexes of parental rats 

in all dose groups of both generations and dilated renal pelves in F1 parental males of the mid- and 

high-dose groups. The results of the study are consistent with the one-generation pilot study that 

was previously conducted. In the one-generation study, fertility was unaffected in male and female 

rats exposed to dietary DINP concentrations as high as 1.5% (966–1,676 and 1,114–1,694 mg/kg 

bw/day in males and females, respectively). The findings of these studies indicate that male and 

female rat fertility and structure of reproductive organs are unaffected by exposure to DINP at a 

maternal dose of 555–1,129 mg/kg bw/day during gestation and lactation, respectively, and adult 

exposure to concentrations as high as 1,676 mg/kg bw/day in males and 1,694 mg/kg bw/day in 

females.
Appendix II

II-29

Mode of Action

DINP exhibited no activity in an in vitro assay that measured binding of phthalates to rat uterine 

cytosolic estrogen receptors (40) and in an assay of estrogen-induced gene expression (46). The 

assays did not include the addition of esterases or lipases to metabolize the DINP to MINP. In 

vivo assays demonstrated that DINP does not increase uterine wet weight or vaginal epithelial cell 

cornification in immature or mature ovariectomized rats (40). There were no studies located on antiandrogenic activity. Thyroid and estrogen serum levels were unaffected in adult marmosets at doses 

as high as 2,500 mg/kg bw/day for 13 weeks (20).

Table 10: Summary 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

Repro Systemic

Two-generation reproductive dietary study in 

Sprague-Dawley rats. 

30 per group were fed 

diets with 0, 0.2, 0.4, 

or 0.8% (Males: 0, 

165–189, 331–379, and 

665–779 mg/kg bw/day, 

Females: 0, 182–197, 

356–397, and 696–802 

mg/kg bw/day**) from 

10 weeks prior to mating 

through gestation and 

lactation.

DINP-1

(36)*

665–779 (M); 

696–802 (F) 

(Reproductive)

None (Systemic)

No effects on 

reproductive 

structure or 

function.

M: 165–189;

F: 182–197

↑ Mild liver 

effects in F0

and F1.

↑Kidney weight 

in F0 females.

None

* Only effects in parental rats are listed. Developmental effects are listed in Table 9.

 ** Doses during the premating period-Combined for F0 and F1 rats.

5.1.4.1 Utility of Data to the CERHR Evaluation

The data are sufficient to indicate that DINP exposures are not associated with detectable effects 

on reproductive function. The studies did demonstrate consistent effects on the liver (weight and 

histology) and kidney (weight). Given the constraints of the study design, the data demonstrate 

no likely reproductive toxicity at doses up to 779(M)–802(F) mg/kg bw/day in the two-generation 

study or at 1,676(M)–1,694(F) mg/kg bw/day in the one-generation study. However, the studies did 

not assess endpoints of reproductive development shown to be sensitive with other phthalates.
II-30

Appendix II

5.2 Integrated Evaluation

DINP is a complex substance of branched, predominantly C-9 isomers. There are no human data 

from which to assess the health effects associated with DINP exposure; studies of DINP toxicity are 

limited to laboratory animals. In the absence of human data and barring evidence to the contrary, it 

is assumed that the effects observed in laboratory animals are relevant to humans.

Based upon the physicochemical similarities between DINP and DEHP and on limited DINP monitoring data, general population exposures to DINP are expected to be lower than those to DEHP 

which are estimated at 3−30 μg/kg bw/day. It is reasonable to presume that humans would be 

exposed primarily through the oral route. Although data are scant, the ingestion of DINP through 

food does not appear to be common. Children may be exposed to higher levels of DINP (up to 10–

100 fold higher) than adults because infants and small children mouth toys and other articles that 

contain DINP that can migrate into saliva and be swallowed. DINP is not used in medical devices, 

therefore intravenous exposure does not occur.

Dermal absorption of DINP is slow in rats. DINP administered orally to rats is metabolized by gut 

lipases to the monoester, which is rapidly absorbed. DINP and its metabolites are rapidly excreted 

in urine and feces, with no indication of accumulation in tissues with repeated (5×) daily doses. At 

low doses, approximately equal amounts of the DINP-derived material are excreted in urine and 

feces, with the urinary metabolites consisting of the monoester and its oxidation products, while the 

feces contain those metabolites plus the diester. There are no toxicokinetic studies in humans, but 

in vitro studies comparing the dermal uptake of other phthalates in human and rat skin suggest that 

dermal uptake of DINP in humans would be negligible.

Oral exposure to DINP has been shown to cause liver and kidney toxicity in adult rats and mice, 

but not in marmosets. The liver effects are generally consistent with those associated with peroxisome proliferation. Liver tumors have been noted in adult male rats exposed to 733 mg/kg bw/day, 

in female mice exposed to 336 mg/kg bw/day, and in male mice exposed to 742 mg/kg bw/day. 

Kidney tumors were noted in male rats, but these tumors are associated with a mechanism that 

is not believed relevant to humans (alpha-2-microglobulin). However, the increased incidence of 

nephropathy seen in female mice exposed to 1,888 mg/kg bw/day is not consistent with the alpha2-microglobulin mechanism. 

The developmental studies available include examination of effects of prenatal exposure on prenatal 

development, as well as a limited assessment of postnatal developmental effects in one- and twogeneration reproductive studies. The prenatal studies provide consistent results and are sufficient to 

establish that oral exposure to DINP causes fetal skeletal variations (lumbar and cervical ribs) and 

in some cases, urinary tract effects (hydroureter). The Panel was confident that 500 mg/kg bw/day 

was an effect level, and 100 mg/kg bw/day was a NOAEL. For one study, sponsors estimated a 

BMD for a 5% excess risk level of 193 mg/kg bw/day with a 95% lower confidence limit of 162 

mg/kg bw/day. For the second developmental toxicity study, the Expert Panel identified a developmental NOAEL of 200 mg/kg bw/day. In addition, the results of the one- and two-generation 

dietary reproductive toxicity studies demonstrated a consistent reduction in mean pup body weights 

during lactation at doses as low as 143–285 mg/kg bw/day (doses during gestation and lactation) 
Appendix II

II-31

and a NOAEL could not be identified. This effect level is similar to that obtained in the more robust 

prenatal study, although the effects are different and a similar mode of action is not assumed. 

There is evidence that isononyl alcohol, a primary metabolite of DINP, is a developmental and 

maternal toxicant at high (~1,000 mg/kg) oral doses in rats. These doses appear to be greater than 

the doses of DINP that are associated with developmental toxicity, suggesting that effects at lower 

doses are probably associated with the monoester. The Panel does acknowledge there are no data to 

permit a judgment about an interactive effect between the alcohol and monoester metabolite.

Reproductive performance and histological effects on gonads and accessory sex organs were 

assessed in one- and two-generation dietary studies. Parental doses of up to 0.8% in feed (665–779 

[M] and 696–802[F] mg/kg bw/day) did not affect fertility or sex organ histology in either the F0

or F1 male or female pups. A 13-week gavage study in adult marmosets resulted in no evidence 

of microscopic testicular changes at doses that did adversely affect body weight gain (2,500 mg/

kg bw/day). Testicular lesions were not observed in prepubertal cynomolgus monkeys that were 

gavaged for 2 weeks with 500 mg/kg bw/day, reportedly the maximum dose that can be absorbed by 

the monkeys. Chronic 2-year studies in rats and mice gave no gross or histologic evidence of effects 

on testes or ovaries at doses that did cause liver and kidney effects and other clinical signs of toxicity. Thus, the data are sufficient to conclude that neither the reproductive organs nor fertility are 

affected by extended oral exposure to DINP. However, the Panel noted that some endpoints which 

are sensitive to other phthalates (i.e., preputial separation and nipple retention) were not evaluated 

in the two-generation study. The Panel is aware that additional data on reproductive tract development are being developed, but as yet only abstracts are available for review. The Panel also notes 

that the target organs in studies with adult rats, liver and kidney, are also target organs in developmental and multigeneration studies. This increases the Panel’s confidence that these effects are real, 

and that different organ system susceptibilities between adults and young are unlikely.

5.3 Expert Panel Conclusions

DINP is used in toys, construction, and general consumer products. Although data are scant, exposure through food appears to be lower than for DEHP. Therefore, the Expert Panel believes that 

adult exposure to DINP will not exceed levels of 3–30 μg/kg bw/day, the estimates derived for 

DEHP. Exposures to DINP are likely to be below this level, but the Panel cannot quantitate how far 

below. Occupational exposures could occur through inhalation and dermal contact. Limited studies of occupational exposures suggest that inhalation exposure is below 1 mg/m3

 during production 

of DINP and below 2 mg/m3 during production of PVC. Although estimates of dermal exposure 

are not available, the Expert Panel is confident that dermal exposure would not result in significant 

absorption into the body. Children could be exposed to DINP by eating contaminated food. However, DINP has not been detected in a limited survey of infant formula. By analogy, the Expert 

Panel believes that children’s exposure to DINP via food will not exceed those levels estimated for 

DEHP. Additional exposure to children may occur due to mouthing of toys and other objects that 

contain DINP. Current models of non-dietary oral exposure predict that older infants and toddlers 

may incur exposures up to one order of magnitude higher than the upper limit of exposure expected 

for adults.
II-32

Appendix II

The toxicology database is sufficient to determine that oral maternal exposure to DINP can result 

in developmental toxicity to the conceptus. In rats, two prenatal developmental studies have 

shown effects on the developing skeletal system and kidney following oral exposure to DINP. The 

NOAELs for these studies were 100–200 mg/kg bw/day. In addition, developmental toxicity was 

noted in an oral two-generation reproductive toxicity study in rats. In this study, effects on pup 

growth were noted. These effects may be due to prenatal and/or lactational exposures to DINP. The 

LOAEL for the study was 143–285 mg/kg bw/day and a NOAEL was not identified. Based on the 

results of the toxicology studies, oral exposure to pregnant women and oral exposure to children 

should be examined. The Expert Panel has minimal concern for unborn children due to ambient 

maternal exposure to DINP. Based on estimates of exposure to DINP in toys and other objects that 

children may mouth, the Expert Panel has low concern for potential health effects in children. 

The Expert Panel cannot judge the potential health effects in unborn children following maternal 

occupational exposures due to the lack of toxicology data following inhalation exposures and the 

lack of occupational exposure information. 

The oral prenatal developmental toxicity studies and the oral two-generation reproductive toxicity 

study have shown no effects on the reproductive system in rats. The NOAEL for reproductive 

toxicity is 665–779 (M) and 696–802 (F) mg/kg bw/day. The Expert Panel noted that some 

endpoints of reproductive development that have been shown to be sensitive with other phthalates 

were not assessed in the two-generation study of DINP, and therefore the Panel has only moderate 

confidence in the NOAEL. The Expert Panel has minimal concern about DINP resulting in 

reproductive toxicity in humans.

5.4 Critical Data Needs

Critical data needs are discussed under two categories: experimental studies and human exposures.

Experimental studies.

Since some relevant endpoints (i.e., nipple retention) were lacking in many of the studies reviewed, 

uncertainties would be reduced if this additional information were gathered. The Expert Panel recommends a sequential approach for future studies that would focus on obtaining the most critical 

information first; subsequent studies would be dependent upon the results of the initial study. The 

Panel further recognized that data gathering should be an iterative process and that recommendations may change as initial tiers of data are gathered. The Expert Panel recommends that the following sequential steps be considered.

1) Conduct a perinatal developmental study in orally exposed rats that addresses landmarks of 

sexual maturation such as nipple retention, anogenital distance, age at testes descent, age at 

prepuce separation, and structure of the developing reproductive system in pubertal or adult 

animals exposed through development. Although a two-generation reproductive toxicity study 

has evaluated some relevant endpoints, the recommended study would provide greater certainty 

about the lack of such effects with DINP. If DINP does affect these endpoints and the effective 

dose levels are of possible human health concern, then the Expert Panel recommends that the 

following study (2) be conducted. 
Appendix II

II-33

2) Conduct a perinatal developmental study with oral exposure in a non-rodent species. There are 

species differences in the developmental toxicity associated with other phthalates. The developmental effects of DINP have only been examined in the rat. Therefore, there is some uncertainty 

whether other species would exhibit similar responses and whether the rat is an appropriate 

model for assessing potential human risk.

Human exposure.

Human exposure to DINP has not been well studied; there are no reports of levels in biological 

materials (blood, urine, etc.), and the environmental data consist primarily of estimates.

Patterns of use, expected environmental levels, and vulnerability of exposed population groups 

should dictate decisions about measuring DINP in environmental media. For example, determining 

DINP exposures in young children is of highest priority, based on the use patterns and vulnerability 

described above. Workers producing PVC products are a second priority.

Collection of biological samples de novo should be accompanied by environmental measurements 

to provide information on exposure sources. Existing biological samples should be utilized where 

available if they can provide useful information about exposure.

Although information about exposure of young children is a critical data need, manufacturers of 

children’s toys should be polled to determine if their products will continue to contain DINP in the 

future. If so, an estimate of the DINP content should be made by the manufacturer and confirmed 

by independent studies. Salivary extraction of DINP and better estimates of mouthing behavior, 

especially within the potentially highest risk group of 3–12 month-old children, using data from 

more children, should be carried out. 
II-34

Appendix II

6.0 REFERENCES

1. CMA. Comments of the Chemical Manufacturers Association phthalate esters panel in response to 

request for public input on seven phthalate esters. FR Doc. 99-9484. Washington, DC: Chemical 

Manufacturers Association, 1999.

2. American Chemistry Council Phthalate Esters Panel. Comments on Draft CERHR Evaluation of DiIsononyl Phthalate (DINP):ACC, 2000.

3. Staples CA, Peterson DR, Parkerton TF, Adams WJ. The environmental fate of phthalate esters: A 

literature review. Chemosphere 35:667-749(1997).

4. MAFF. Phthalates in infant formulae. Joint food safety and standards group food surveillance 

information sheet, vol 1999:MAFF - UK, 1996;7p.

5. MAFF. Phthalates in infant formulae. Food surveillance information sheet 168: Joint Food Safety 

and Standards Group, 1998.

6. MAFF. Phthalates in food. Joint food safety and standards group food surveillance information 

sheet, vol 1999:MAFF - UK, 1996;9p.

7. CPSC. The risk of chronic toxicity associated with exposure to diisononyl phthalate (DINP) in 

children’s products. Bethesda, MD, 1998.

8. Rastogi SC. Gas chromatographic analysis of phthalate esters in plastic toys. Chromatographia 47:

724-726(1998).

9. Marin ML, Lopez J, Sanchez J, Vilaplana J, Jimenez A. Analysis of potentially toxic phthalate 

plasticizers used in toy manufacturing. Bull Environ Contam Toxicol 60:68-73(1998).

10. Toy Manufacturers of America I-T. Voluntary standard specification for the reduction of DEHP 

(di[2-ethylhexyl]phthalate) in PVC pacifiers and teethers.: Toy Manufacturers of America, Inc., 

1986.

11. Canada H. Updated. Risk assessment on diisononyl phthalate in vinyl children’s products. Ottawa, 

Ontario: Consumer Products Division, Product Safety Bureau, Environmental Health Directorate, 

Health Protection Branch, 1998.

12. RIVM, Milieu RVVE. Phthalate release from soft PVC baby toys RIVM report 613320 002: 

National Institute of Public Health and the Environment, 1998.

13. Fiala F, Steiner I, Kubesch K. Migration of di-(2-ethylhexyl)phthalate (DEHP) and diisononyl 

phthalate (DINP) from PVC articles. (2000).

14. Juberg DR, Thompson KM, Alfano K, Coughlin RJ. An observational study of object mouthing 

behavior by young children. Rochester, New York: International Center for Toxicology and 

Medicine, Submitted to journal.

15. Doull J, Cattley R, Elcombe C, Lake B, Swenberg J, Wilkinson C, Williams G. Expert panel report 

on DEHP.: U.S. Environmental Protection Agency, 1998.

16. BIBRA BIBRA-. A 21-day feeding study of di-isononyl phthalate. Report No. 0495/6/85.: Chemical 

Manufacturers Association, 1985.

17. Lington AW, Bird MG, Plutnick RT, Stubblefield WA, Scala RA. Chronic toxicity a carcinogenic 

evaluation of diisononyl phthalate in rats. Fundam Appl Toxicol 36:79-89(1997).

18. Moore MRCL. Oncogenicity study in rats with di(isononyl)phthalate including ancillary 

hepatocellular proliferation and biochemical analyses. Covance 2598-104 Volume 1 of 5. Vienna, 

VA: Aristech Chemical Corporation, 1998.
Appendix II

II-35

19. Moore MRCL. Oncogenicity study in mice with di(isononyl)phthalate including ancillary 

hepatocellular proliferation and biochemical analyses. Covance 2598-105 Volume 1 of 6. Vienna, 

VA: Aristech Chemical Corporation Performing Laboratory, 1998.

20. Hall M, Matthews A, Webley L, Harling R. Effects of dii-isononyl phthalate (DINP) on peroxisomal 

markers in the marmoset - DINP is not a peroxisome proliferator. The Journal of Toxicological 

Sciences 24:237-244(1999).

21. Pugh G, Isenberg JS, Kamendulis LM, Ackley DC, Clare LJ, Brown R, Lington AW, Smith JH, 

Klaunig JE. Effects of di-isononyl phthalate, di-2-ethylhexyl phthalate, and clofibrate in cynomolgus 

monkeys. Toxicol Sci 56:181-188(2000).

22. Short RD, Robinson EC, Lington AW, Chin EC. Metabolic and peroxisome proliferation studies with 

di (2-ethylhexyl) phthalate in rats and monkeys. Toxicol Ind Health 3:185-194(1987).

23. Caldwell DJ. Review of mononuclear cell leukemia (MNCL) in F-344 rat bioassays and its 

significance to human cancer risk: A case study using alkyl phthalates. Accepted for publication in 

Regulatory Toxicology and Pharmacology(1999).

24. Midwest Research Institute M. Dermal disposition of 14C-diisononyl phthalate in rats 35320. Kansas 

City, MI: Exxon Corporation, Medical Department, Research and Environmental Health, P.O. Box 

235, East Millstone, NJ, 1983.

25. Scott RC, Dugard PH, Ramsey JD, Rhodes C. In vitro absorption of some o-phthalate diesters 

through human and rat skin. Environ Health Perspect 74:223-227(1987).

26. Deisinger PJ, Perry LG, Guest D. In vivo percutaneous absorption of DEHP from DEHP- plasticized 

polyvinyl chloride film in male Fischer 344 rats. Food Chem Toxicol 36:521-527(1998).

27. Hazelton HL-L. Metabolism study of 14C phthalate ester in rats - Final Report 32563. Vienna, VA: 

Esso Research and Engineering Company, 1972.

28. Midwest Research Institute M. Single and repeated oral dose pharmacokinetics of 14 C-labeled 

diisononyl phthalate with cover letter. MRI project No. 7282-B. Kansas City, MI: Exxon 

Corporation, Medical Department, Research and Environmental Health, P.O. Box 235, East 

Millstone, NJ, 1983.

29. OECD. Risk assessment - 1,2-Benzenedicarboxylic acid, di-C8-10-branched alkyl esters C9-rich and 

di-”isononyl”phthalate CAS No.: 68515-48-0 and CAS No.: 28553-12-0: EINECS-No.: 271-090-9 

and EINECS-No.: 249-079-5, 1998.

30. Barber E, Cifone M, Rundell J, Przygoda R, Astill B, Moran E, Mulholland A, Robinson E, 

Schneider B. Results of the L5178Y mouse lymphoma assay and the Balb/3t3 cell invitro 

transformation assay for eight phthalate esters. J Appl Toxicol 20:69-80(2000).

31. Waterman SJ, Ambroso JL, Keller LH, Trimmer GW, Nikiforov AI, Harris SB. Developmental 

toxicity of di-isodecyl and di-isononyl phthalates in rats. Reprod Toxicol 13:1-6(1999).

32. Ryan L. The use of generalized estimating equations for risk assessment in developmental toxicity. 

Risk Analysis 12:439-447(1992).

33. McKee R. Personal communication to Jack Moore, 2000.

34. Chen JJ. Dose-response modeling of growth for develomental toxicity. Environmetrics 7:135-

144(1996).

35. Hellwig J, Freudenberger H, Jackh R. Differential prenatal toxicity of branched phthalate esters in 

rats. Food Chem Toxicol 35:501-512(1997).

36. Waterman SJ, Keller LH, Trimmer GW, Freeman JJ, Nikiforov AI, Harris SB, Nicolich MJ, McKee 
II-36

Appendix II

RH. Two-generation reproduction study in rats given di-isononyl phthalate in the diet. Reprod 

Toxicol 14:21-36(2000).

37. Hellwig J, Jackh R. Differential prenatal toxicity of one straight-chain and five branched-chain 

primary alcohols in rats. Food Chem Toxicol 35:489-500(1997).

38. Ostby J, Price M, Furr J, Lambright C, Hotchkiss A, Parks l, Gray Jr LE. Perinatal exposure to the 

phthalates DEHP, BBP, DINP, but not DEP, DMP or DOTP permanently alters androgen-dependent 

tissue development in Sprague-Dawley rats. Triangle Consortium or Reproductive Biology Jan 29, 

2000(2000).

39. Gray LE, Ostby J, Furr J, Price M, Veeramachaneni DNR, Parks L. Perinatal exposure to the 

phthalates in DEHP, BBP and DINP but not DEP, DMP or DOTP alters sexual differentiation of the 

male rat. Toxicol. Sci. (in press)(2000).

40. Zacharewski TR, Meek MD, Clemons JH, Wu ZF, Fielden MR, Matthews JB. Examination of the 

in vitro and in vivo estrogenic activities of eight commercial phthalate esters. Toxicol Sci 46:282-

293(1998).

41. Jobling S, Reynolds T, White R, Parker MG, Sumpter JP. A variety of environmentally persistent 

chemicals, including some phthalate plasticizers, are weakly estrogenic. Environ Health Perspect 

103:582-587(1995).

42. Bolger R, Wiese TE, Ervin K, Nestich S, Checovich W. Rapid screening of environmental chemicals 

for estrogen receptor binding capacity. Environ Health Perspect 106:551-7(1998).

43. 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).

44. 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).

45. 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).

46. Harris CA, Henttu P, Parker MG, Sumpter JP. The estrogenic activity of phthalate esters in vitro. 

Environ Health Perspect 1997 105:802-811(1997).

47. Scott RC, Dugard PH, Ramsey JD, Rhodes C. In vitro absorption of some o-phthalate diesters 

through human and rat skin. Environ Health Perspect 74:223-227(1987).
Appendix II

II-37

7.0 TABLES

7.1 DINP General Toxicity, Rats..................................................................................................38

7.2 DINP General Toxicity, Rats..................................................................................................39

7.3 DINP General Toxicity, Rats..................................................................................................40

7.4 DINP General Toxicity, Mice .................................................................................................41

7.5 DINP General Toxicity, Marmosets .......................................................................................42

7.6 DINP Developmental Toxicity, Rats ......................................................................................43

7.7 DINP Developmental Toxicity, Rats ......................................................................................44

7.8 DINP Developmental Toxicity, Rats ......................................................................................45

7.9 DINP Developmental Toxicity, Monkeys...............................................................................46

7.10 References..............................................................................................................................47
II-38

Appendix II

Table 7-1: DINP, General Toxicity, Rats

Species, 

Strain, and 

Source

Experimental 

Regimen

Animal 

Number/

Sex

Dose* Body Weight Organ Weight** Liver Effects Hematology Other

Fischer 344 RatBIBRA 1985(1)

Subchronic study :

21 days.

Rats (∼41−43 days old) were fed diets with 0, 0.6, 1.2, or 2.5% DINP and then sacrificed and necropsied. Peroxisome proliferation was studied by electron microscopy examination of liver and measurement of peroxisomal enzyme activity. 

5 05 639(M)

607(F)

NE ↑Li, Ki ↑11-OH(M) and 

12-OH(M).

↑Protein.

NA ↓Serum Ch, 

Tg(M).

5 1,192(M)

1,193(F)

↓ ↑Li, Ki ↑PCoA(M), 11-OH 

and 12-OH(M).

↑Protein.Histological 

changes.

NA ↓Serum Ch, 

Tg(M).

↑Serum Tg(F).

5 2,195(M)

2,289(F)

↓ ↑Li, Ki, 

Te

↑PCoA, 11-OH 

and 12-OH.

↑Peroxisomes.↑Protein.Histological 

changes.

NA ↓Serum Ch, 

Tg(M).

↑Serum Tg(F).No dose-related 

testicular lesions.

↑A group of positive control rats was administered DEHP at 1.2%.

5 1,084(M)

1,063(F)

↓ ↑Li, Ki ↑PCoA, 11-OH 

and 12-OH.

↑Peroxisomes.↑Protein.

NA ↓Serum Ch, 

Tg(M).

No dose-related 

testicular lesions.

 *Dose measured in mg/kg bw/day.**Organ to body weight ratio.↑=Statistically Significant Increase↓=Statistically Significant Decrease

NA=Not AnalyzedNE=No EffectsF=FemaleM=Male

Te=TestesLi=LiverKi=Kidney

11-OH=11-HydroxylasePCoA=Palmitoyl-CoA Oxidas12-OH=12-Hydroxylase

Tg=Triglyceride Ch=Cholesterol
Appendix II

II-39

Table 7-2: DINP, General Toxicity, Rats

Species, 

Strain, and 

Source

Experimental 

Regimen

Animal 

Number/

Sex

Dose* Body Weight Organ Weight** Liver Effects Hematology Other

Fischer 344 

Rat

Lington et 

al. 1997

(2)

Chronic study : 

2 years.

6-week-old rats were fed diets with 0, 0.03, 0.3, and 0.6% DINP1 for 2 years. 10rats/sex/group were sacrificed and necropsied at 6, 12, and 18 months and the rest were killed and necropsied at the end of the study. Hematology, serum chemistry, and urinalysis were evaluated every 6 months. Peroxisome proliferation was examined microscopically in 2 rats/sex/group at 24 months.

110 0110 15(M)

18(F)

NE NE NE NE NOAEL

110 152(M)

184(F)

↓(M; 18–24 mo)

↑Li, Ki(6–24 mo)

Hepatic lesions 

(24 mo)

↑SGOT (M; 6–12 mo)↑SGPT (M; 24 mo)↑MNCL.

NE ↓Survival (F).

110 307(M)

375(F)

↓(M; 12–24 mo)

↑Li, Ki(6–24 mo)↑Sp (24 mo)↑Ad (24 mo)

Hepatocyte 

enlargement (6–24 mo) and lesions (24 mo)

↑SGOT (M; 6–18 mo) ↑SGPT(M; 6 &18 mo) ↑MNCL.No evidence of peroxisomal proliferation

↓RBC, Hb, 

Hct (M; 24 mo)

↓Survival (F).↑Urine volume 

(M; 6–24 mo)

↑Urine K and 

glucose (M; 6–18 mo)

No evidence of testicular damage.

 *Dose measured in mg/kg bw/day**Organ to body weight ratio↑=Statistically Significant Increase↓=Statistically Significant Decrease

NE=No EffectsF=FemaleM=MaleMo=Month

Ad=AdrenalKi=KidneyLi=LiverSp=Spleen

SGOT=Serum Glutamic Oxaloacetic TransaminaseSGPT=Serum Glutamic Pyruvic TransaminaseMNCL=Mononuclear Cell LeukemiaRBC=Red Blood CellHb=HemoglobinK=Potassium
II-40

Appendix II

Table 7-3: DINP, General Toxicity, Rats

Species, 

Strain, and 

Source

Experimental Regimen

Animal 

Number/

Sex

Dose Body Weight Organ Weight** Liver Effects Hematology Other

F344 RatsMoore 1998(3)

Chronic study : 2 years.6-week old rats were fed diets with 0, 500, 1,500, 6,000, or 12,000 DINP. 5 rats/sex/dose were killed on weeks 1, 2, and 13; 15 rats/sex/dose were killed at 79 weeks; and 55 rats/sex/group at 104–106 weeks. Clinical evaluations (hematology, serum chemistry, and urinalysis) were conducted every 26 weeks.Peroxisome proliferation was examined in 5 rats/sex only in the controls and high dose group during weeks 1, 2, and 13 and in 3–5 rat/sex in the control, 359–442 mg/kg bw/day group, and highdose group at week 104.

85 070 29.2(M)

36.4(F)

NE NE NE NE NE

70 88.3(M)

109(F)

NE NE NE NE NOAEL

85 359(M)

442(F)

NE ↑Ki 

(wk 79-104) 

↑Li 

(wk 1-104)

↑PCoA (F, wk 104)↑ASAT, ALAT 

(wk 52, 78, 104)

↑Anemia Kidney lesions 

(M, wk 79–104)

↑ Serum urea 

(wk 26–104)

MNCL (wk 104)

85 733(M)

885(F)

↓ (wk 

9–104)

↑Ki 

(wk 79-104)

↑Li 

(wk 1-104)

↑PCoA (wk 1–104)Lesions 

(wk 2–104)

Neoplasia (M, 

wk 79–104)

↑ASAT, ALAT 

(wk 52, 78, 104)

↓ASAT, ALAT 

(F, wk 26)

↑Anemia ↓Survival (M)

↑Serum urea 

(wk 26–104) 

↑Urine vol with ↓Cl, Ca, K, 

Cre (M, wk 104)

Kidney lesions 

(wk 79–104)

Kidney neoplasm 

(M, wk 104)

MNCL (wk 104)No testicular effects

A group of 55 rats/sex was exposed to the high dose for 78 weeks and sacrificed at 105–106 weeks to study recovery effects.*

55 637(M)

774(F)

↓(F) ↑Ki(F) NE ↑Urine vol with ↓ Cre(M)

MNCL (wk 104)Kidney lesions and 

neoplasm (M)

 *Only effects observed by week 104 listed**Organ to body weight ratio↑=Statistically Significant Increase↓=Statistically Significant Decrease

NE=No EffectsF=FemaleM=Male

Ki=KidneyLi=Liverwk=Weekvol=Volume

ALAT=Alanine aminotransferase Ca=CalciumASAT=Aspartate aminotransferase Cl=ChlorideMNCL=Mononuclear Cell Leukemia Cre=CreatininePCoA=Palmitoyl-CoA Oxidase K=Potassium
Appendix II

II-41

Table 7-4: DINP, General Toxicity, Mice

Species, 

Strain, and 

Source

Experimental Regimen

Animal 

Number/

Sex

Dose* Body Weight Organ Weight*** Liver Effects Hematology Other

B6C3F1/CrlBR miceMoore 1998(4)

Chronic study: 2 years.6-week-old mice were fed diets with 0, 500, 1,500, 4,000, and 8,000 ppm DINP. 15 mice/dose/ sex were evaluated and sacrificed at 79 weeks and 55 mice sex/group at 105–106 weeks. Clinical evaluations (hematology, serum chemistry, and urinalysis) were conducted every 26 weeks. Peroxisome proliferation was examined in 5 mice/sex in the highest dose group and controls during the midpoint and end of study.

70 070 90.3(M)

112(F)

NE NE NE NE NOAEL (F)

70 276(M)

336(F)

NE NE ↑Neoplasia (F) NE NOAEL (M)

70 742(M)

910(F)

↓ (wk 

1–104)

↓Ki(M), ↑Li(M)(wk 79-104)

↑Neoplasia (M) NE NE

70 1,560(M)

1,888(F)

↓ (wk 

1–104)

↓Ki (M),↑Li(wk 79-104)

↑Neoplasia 

and nonneoplastic changes 

↑Serum ASAT, 

ALAT (M)

↑PCoA 

(wk 79–104)

↓WBC 

(wk 26–98)

↓Survival (M)↑Nephropathy(F)↑Serum protein

 (M, week 104)

↑Urinary vol with 

↓Na, Cl, K (week 52–104)

No effects on 

testicular histology

A group of 55 mice/sex was exposed to the high dose for 78 weeks and sacrificed at 105–106 weeks to study recovery effects.**

55 1,377(M)

1,581(F)

↓(M) ↓Ki(M) ↑Neoplasia NE NE

*Dose measured in mg/kg bw/day.

 **Only effects observed by week 104 listed***Organ to body weight ratio ↑=Statistically Significant Increase ↓=Statistically Significant Decrease

NE=No EffectsF=FemaleM=Malewk=Weekvol=Volume

Ki=KidneyLi=LiverWBC=White Blood Cell

ALAT=Alanine aminotransferaseASAT=Aspartate aminotransferasePCoA=Palmitoyl-CoA Oxidase

Cl=ChlorideK=PotassiumNa=Sodium
II-42

Appendix II

Table 7-5: DINP, General Toxicity, Marmosets

Species, 

Strain, and 

Source

Experimental 

Regimen

Animal 

Number/

Sex

Dose* Body Weight Organ Weight Histopathology Hematology Chemistry Other

MarmosetHall et al. 1999(5)

Subchronic study: 

13 Weeks. 

Male and female marmosets (16–25 months old) were gavaged with DINP in 1% methylcellulose and 0.5% Tween. Clofibrate was used as a positive control. Parameters evaluated at sacrifice included estradiol and testosterone concentrations, biochemical evidence of peroxisomal proliferation, and organ weights and histopathology.

1–2 0

2 100 NE NE NE NE NE2 500 NE NE NE NE NE2 2,500 decreasea NE NE NE No change 

in PCoA

Ungroomed coat and reddening of the skin.Thin appearance, hunched posture and reduced activity in one male.a

1–2 500 

Clofibrate

decreasea ↑ Ki to 

body weight ratioa

NE ↑ Anemia ↑ PCoA

↑ 11-OH

 aDose measured in mg/kg bw/day. *Efects were not statistically significant ↑=Statistically Significant Increase

NE=No Effects 11-OH = Lauric acid 11-hydroxylase activity

PCoA=Palmitoyl-CoA Oxidase activity
Appendix II

II-43

Table 7-6: DINP Developmental Toxicity, Rats

Species, Strain, 

and Source Experimental Regimen Numberb Dosed Maternal Effects Fetal Effects

Wistar RatHellwig et al. 1997(7)

Prenatal developmental toxicity study.Three types of DINPamanufactured by different processes were administered in oil by gavage on gd 6–15. Dams were weighed on gd 0, 6, 10, 15, and 20 and sacrificed on gd 20. Maternal uteri were weighed, corpora lutea were counted and implantation sites examined. Fetuses were weighed and examined for gross external malformations. Half of the fetuses were examined for visceral malformations and the other half for skeletal malformations. 

9 0

9–10 40 ↑Kidney to bodyweight 

ratio in DINP-2.

↓Implantation sites/

litter in DINP-2.

↓Live fetuses/litter in DINP-2.

8–10 200 NE NE9–10 1,000 ↑Kidney to bodyweight 

ratio in DINP-1.

↑Liver to bodyweight 

ratio in DINP-3.

↑Fetuses/litter with cervical ribs in:

DINP-1: 11 fetuses/5 litters vs 0 fetusesDINP-2: 4 fetuses/4 litters vs 0 fetusesDINP-3: 12 fetuses/7 litters vs 0 fetuses.

↑Fetuses/litter with lumbar ribs in:

DINP-1: 37 fetuses/10 litters vs 0 fetusesDINP-2: 10 fetuses/5 litters vs 0 fetusesDINP-3: 34 fetuses/8 litters vs 0 fetuses.

↑Fetuses/litter with hydroureter in

DINP-3: 2 fetuses/8 litters vs 4 fetuses/3 litters.

↑Fetuses/litter with malformations in:

DINP-3 (7.3 vs 4.3%).c

↑ Dilated renal pelves in DINP-1,2,3.

a (1) CAS RN: 68515-48-0; (2) CAS RN: 28553-12-0; (3) CAS RN: 28553-12-0 (2 and 3 by different manufacturing process).

b Number of litters examined per type of DINP. 

c Skeletal and visceral malformations (humorous, femur, kidney, and ureter); not statistically significant.

d Dose measured in mg/kg bw/day.

↑= Statistically Significant Increase↓= Statistically Significant Decrease

NE= No Effect 
II-44

Appendix II

Table 7-7: DINP, Developmental Toxicity, Rats

Species, Strain, 

and Source Experimental Regimen Number* Dose** Maternal Effects Fetal Effects

Sprague-Dawley RatWaterman et al. 1999(8)

Prenatal developmental toxicity study.DINP-1 administered in oil by gavage on gd 6–15. Sacrificed on gd 21. Dams weighed on gd 0, 6, 9, 12, 15, 18, and 21. Maternal uterus and ovaries were weighed, corpora lutea were counted and implantation sites examined. Fetuses were weighed, sexed, and examined for gross external malformations. Half of the fetuses were examined for visceral malformations and the other half for skeletal malformations.

24 025 100 NE ↑ % Fetuses with dilated renal pelves 

(3.7 vs 0%).

24 500 NOAEL ↑ % Fetuses with dilated renal pelves 

(4 vs 0%). 

↑ % Fetuses with lumbar ribs

 (19 vs 4%).

23 1,000 ↓ Weight gain 

(transient).

↓ Food intake 

(transient).

↑% Litters with dilated renal pelves 

(26 vs 0%).

↑ % Fetuses with dilated renal pelves 

(4.5 vs 0%).

↑ % Litters with lumbar ribs 

(78 vs 25%).

↑ % Fetuses with lumbar ribs 

(35 vs 4%).

↑ % Fetuses with cervical ribs 

(6 vs 2%).

* Number of litters examined.

** Dose measured in mg/kg bw/day.

↑ = Statistically Significant Increase↓ = Statistically Significant Decrease

NE = No Effect
Appendix II

II-45

Table 7-8: DINP, Reproductive Toxicity, Rats

Species, Strain, 

and Source Experimental Regimen

Animal 

Number/

Sex

Dose* Effects

CD RatsWaterman et al. 2000(9)

Two generation reproductive toxicity study. DINP administered in feed for 10 weeks prior to mating at 0, 0.2, 0.4, or 0.8%. Males treated until delivery of last litter and females through gestation to lactation. Breeding pairs housed together for 3 weeks; body weight and food intake was measured weekly. One male and female from each litter reared to adulthood and remaining pups were examined and discarded.

30 030 165/182/146/254 ↑Kidney weight in F0 females.

↓Weight gain in F1 pups.

30 331/356/287/539 ↑Liver weight in F0 females.

↑Kidney weight in F0 males and females.↓Weight gain in F1 pups.

30 665/696/555/1,026 ↓Weight gain in F0 females (pnd 14, 21).

↑Liver weight in F0 males and females.↑Kidney weight in F0 males and females.↓Left ovary weight in F0 females.↓Weight gain in F1 pups.No effect on weights of male reproductive 

organs, testicular histology, or litter size, mating, offspring survival, and sex ratio.

One male and female F1 rat/litter continued to receive the same doses as parental rats and were then mated within dose groups during adulthood.

30 189/197/143/285 NE30 379/397/288/553 ↓F2 pup weight gain during lactation.30 779/802/560/1,129 ↓Body weight in F1.

↑Liver weight in F1 females and 

kidney weight in F1 males. 

↓F2 pup weight gain during lactation.No effects on mating, fertility, litter size, 

pup weight, survival, or sex ratio, sex organ weights or testicular histology.

*Doses (in mg/kg bw/day) in males during premating/females during premating/females during gestation/females during lactation.↑=Statistically Significant Increase↓=Statistically Significant Decrease
II-46

Appendix II

Table 7-9: DINP, General Toxicity, Monkeys 

Species, 

Strain, and 

Source

Experimental 

Regimen

Animal 

Number/

Sex

Dose* Body Weight Organ Weight Histopathology Hematology Chemistry Other

Cynomolgus monkeyPugh et al. 2000(6)

Subacute study: 

2 weeks. 

2-year-old males were gavaged with DINP in 0.5% methylcellulose for 2 weeks and then sacrificed and necropsied.

4 04 500 NE No effects 

on organ to bodyweight ratios

No testicular or hepatic lesions.

↑Neutrophils↓Lymphocytes

NE No peroxisome 

proliferationNo effects on gap junctional intracellular communication

*Doses measured in mg/kg bw/day.NE=No Effects↓=Statistically Significant Decrease↑=Statistically Significant Increase
Appendix II

II-47

References

1. BIBRA BIBRA-. A 21-day feeding study of di-isononyl phthalate. Report No. 0495/6/85: 

Chemical Manufacturers Association (1985).

2. Lington AW, Bird MG, Plutnick RT, Stubblefield WA, Scala RA. Chronic toxicity a 

carcinogenic evaluation of diisononyl phthalate in rats. Fundam Appl Toxicol 36:79-89 

(1997).

3. Moore MRCL. Oncogenicity study in rats with di(isononyl)phthalate including ancillary 

hepatocellular proliferation and biochemical analyses. Covance 2598-104 Volume 1 of 5. 

Vienna, VA: Aristech Chemical Corporation (1998).

4. Moore MRCL. Oncogenicity study in mice with di(isononyl)phthalate including ancillary 

hepatocellular proliferation and biochemical analyses. Covance 2598-105 Volume 1 of 6. 

Vienna, VA: Aristech Chemical Corporation Performing Laboratory (1998).

5. Hall M, Matthews A, Webley L, Harling R. Effects of dii-isononyl phthalate (DINP) on 

peroxisomal markers in the marmoset - DINP is not a peroxisome proliferator. The Journal 

of Toxicological Sciences 24:237-244 (1999).

6. Pugh G, Isenberg JS, Kamendulis LM, Ackley DC, Clare LJ, Brown R, Lington AW, Smith 

JH, Klaunig JE. Effects of di-isononyl phthalate, di-2-ethylhexyl phthalate, and clofibrate in 

cynomolgus monkeys. Toxicol Sci 56:181-188 (2000).

7. Hellwig J, Freudenberger H, Jackh R. Differential prenatal toxicity of branched phthalate 

esters in rats. Food Chem Toxicol 35:501-512 (1997).

8. Waterman SJ, Ambroso JL, Keller LH, Trimmer GW, Nikiforov AI, Harris SB. 

Developmental toxicity of di-isodecyl and di-isononyl phthalates in rats. Reprod Toxicol 13:

1-6 (1999).

9. Waterman SJ, Keller LH, Trimmer GW, Freeman JJ, Nikiforov AI, Harris SB, Nicolich MJ, 

McKee RH. Two-generation reproduction study in rats given di-isononyl phthalate in the 

diet. Reprod Toxicol 14:21-36 (2000).
Appendix III

PUBLIC COMMENTS

ON THE PHTHALATES

EXPERT PANEL REPORTS

�����������������������������������

���������������������

���������������������������

��������������������������������������������
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III
Appendix III