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Provisional Peer-Reviewed Toxicity Values for 

Diethylene Glycol Monobutyl Ether 

(DGBE, CASRN 112-34-5) 

Superfund Health Risk Technical Support Center 

National Center for Environmental Assessment 

Office of Research and Development 

U.S. Environmental Protection Agency 

Cincinnati, OH 45268 
Commonly Used Abbreviations 

BMD Benchmark Dose 

IRIS Integrated Risk Information System 

IUR inhalation unit risk 

LOAEL lowest-observed-adverse-effect level 

LOAELADJ LOAEL adjusted to continuous exposure duration 

LOAELHEC LOAEL adjusted for dosimetric differences across species to a human 

NOAEL no-observed-adverse-effect level 

NOAELADJ NOAEL adjusted to continuous exposure duration 

NOAELHEC NOAEL adjusted for dosimetric differences across species to a human 

NOEL no-observed-effect level 

OSF oral slope factor 

p-IUR provisional inhalation unit risk 

p-OSF provisional oral slope factor 

p-RfC provisional inhalation reference concentration 

p-RfD provisional oral reference dose 

RfC inhalation reference concentration 

RfD oral reference dose 

UF uncertainty factor 

UFA animal to human uncertainty factor 

UFC composite uncertainty factor 

UFD incomplete to complete database uncertainty factor 

UFH interhuman uncertainty factor 

UFL LOAEL to NOAEL uncertainty factor 

UFS subchronic to chronic uncertainty factor 

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PROVISIONAL PEER-REVIEWED TOXICITY VALUES FOR 

DIETHYLENE GLYCOL MONOBUTYL ETHER (DGBE, CASRN 112-34-5) 

Background

 On December 5, 2003, the U.S. Environmental Protection Agency's (U.S. EPA) Office of 

Superfund Remediation and Technology Innovation (OSRTI) revised its hierarchy of human 

health toxicity values for Superfund risk assessments, establishing the following three tiers as the 

new hierarchy: 

1) U.S. EPA's Integrated Risk Information System (IRIS) 

 2) Provisional Peer-Reviewed Toxicity Values (PPRTVs) used in U.S. EPA's Superfund 

Program 

 3) Other (peer-reviewed) toxicity values, including: 

< Minimal Risk Levels produced by the Agency for Toxic Substances and Disease 

Registry (ATSDR) 

< California Environmental Protection Agency (CalEPA) values 

< EPA Health Effects Assessment Summary Table (HEAST) values 

 A PPRTV is defined as a toxicity value derived for use in the Superfund Program when 

such a value is not available in U.S. EPA's IRIS. PPRTVs are developed according to a Standard 

Operating Procedure (SOP) and are derived after a review of the relevant scientific literature 

using the same methods, sources of data, and Agency guidance for value derivation generally 

used by the U.S. EPA IRIS Program. All provisional toxicity values receive internal review by 

two U.S. EPA scientists and external peer review by three independently selected scientific 

experts. PPRTVs differ from IRIS values in that PPRTVs do not receive the multiprogram 

consensus review provided for IRIS values. This is because IRIS values are generally intended 

to be used in all U.S. EPA programs, while PPRTVs are developed specifically for the Superfund 

Program. 

 Because new information becomes available and scientific methods improve over time, 

PPRTVs are reviewed on a 5-year basis and updated into the active database. Once an IRIS 

value for a specific chemical becomes available for Agency review, the analogous PPRTV for 

that same chemical is retired. It should also be noted that some PPRTV documents conclude that 

a PPRTV cannot be derived based on inadequate data. 

Disclaimers 

 Users of this document should first check to see if any IRIS values exist for the chemical 

of concern before proceeding to use a PPRTV. If no IRIS value is available, staff in the regional 

Superfund and Resource Conservation and Recovery Act (RCRA) program offices are advised to 

carefully review the information provided in this document to ensure that the PPRTVs used are 

appropriate for the types of exposures and circumstances at the Superfund site or RCRA facility 

in question. PPRTVs are periodically updated; therefore, users should ensure that the values 

contained in the PPRTV are current at the time of use. 

 It is important to remember that a provisional value alone tells very little about the 

adverse effects of a chemical or the quality of evidence on which the value is based. Therefore, 

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users are strongly encouraged to read the entire PPRTV document and understand the strengths 

and limitations of the derived provisional values. PPRTVs are developed by the U.S. EPA 

Office of Research and Development’s National Center for Environmental Assessment, 

Superfund Health Risk Technical Support Center for OSRTI. Other U.S. EPA programs or 

external parties who may choose of their own initiative to use these PPRTVs are advised that 

Superfund resources will not generally be used to respond to challenges of PPRTVs used in a 

context outside of the Superfund Program. 

Questions Regarding PPRTVs

 Questions regarding the contents of the PPRTVs and their appropriate use (e.g., on 

chemicals not covered, or whether chemicals have pending IRIS toxicity values) may be directed 

to the U.S. EPA Office of Research and Development’s National Center for Environmental 

Assessment, Superfund Health Risk Technical Support Center (513-569-7300), or OSRTI. 

INTRODUCTION 

Diethylene glycol monobutyl ether (DGBE) is a colorless liquid. Its synonyms include 

2-(2-butoxyethoxy)-ethanol, butoxydiglycol, butadigol, and butyl carbitol. The formula for 

DGBE is C8H18O3 (Figure 1) with a molecular weight of 162.23, boiling temperature of 

230.4 C°, and melting point of -68.1C°. 

Figure 1. Diethylene Glycol Monobutyl Ether Structure 

There is no RfD assessment for DGBE on IRIS (U.S. EPA, 2008) or in the HEAST 

(U.S. EPA, 1997) or Drinking Water Standards and Health Advisories list (U.S. EPA, 2006). 

Subchronic and chronic RfDs of 0.04 and 0.004 mg/kg-day, respectively, were derived for 

DGBE in a draft Health and Environmental Effects Document (HEED) on glycol ethers 

(SRC, 1992). The RfDs in the draft HEED were based on a subchronic LOAEL of 

36 mg/kg-day for hematological effects in rats (Hobson et al., 1987) and an UF of 1,000 

(subchronic RfD) or 10,000 (chronic RfD). An earlier Health Effects Assessment (HEA) on 

glycol ethers (U.S. EPA, 1984) did not derive acceptable subchronic or chronic oral intake 

values for DGBE. This HEA is the only document relevant to DGBE in the Chemical 

Assessments and Related Activities (CARA) list (U.S. EPA, 1991, 1994a). 

There is no RfC assessment for DGBE on IRIS (U.S. EPA, 2008) or in the HEAST 

(U.S. EPA, 1997). No occupational exposure limits have been recommended or promulgated by 

the American Conference of Governmental Industrial Hygienists (ACGIH, 2001, 2007), the 

National Institute of Occupational Safety and Health (NIOSH, 2005), or the Occupational Safety 

and Health Administration (OSHA, 2008). 

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 3

 DGBE is not listed in the HEAST cancer table (U.S. EPA, 1997), and no carcinogenicity 

assessment is available on IRIS (U.S. EPA, 2008) or indicated in the Drinking Water Standards 

and Health Advisories list (U.S. EPA, 2006). The carcinogenicity of DGBE has not been 

assessed by the International Agency for Research on Cancer (IARC, 2008) or the National 

Toxicology Program (NTP, 2005, 2008). 

 There are no Toxicological Profiles available from ATSDR (2008) or World Health 

Organization Environmental Health Criteria documents (WHO, 2008) for DGBE. CalEPA 

(2002, 2005a,b) has not derived oral or inhalation recommended exposure limits (RELs) or a 

cancer potency factor for DGBE. 

 Literature searches were conducted from the 1960s through August 2009 for studies 

relevant to the derivation of provisional toxicity values for DGBE. Databases searched included: 

MEDLINE (including cancer subset), TOXLINE (Special), BIOSIS, TSCATS 1/TSCATS 2, 

CCRIS, DART/ETIC, GENETOX, HSDB, RTECS, and Current Contents (February–August 

2009). 

REVIEW OF PERTINENT DATA 

Human Studies 

No pertinent data were located regarding health effects of DGBE in humans following 

oral or inhalation exposure. 

Animal Studies 

Animal studies are summarized in Table 1. 

Oral Exposure 

Short-term Studies—As a preliminary study to a 13-week oral subchronic toxicity 

investigation discussed above, Johnson et al. (2005) administered DGBE (99.2% pure) in 

drinking water to Fischer 344 (F344) rats (five/sex/group) at doses of 0, 1,000, 1,500, or 

2,000 mg/kg-day for 2 weeks. Animals were evaluated for clinical signs, body weight, food 

consumption, water consumption, urinalysis variables, serum chemistry, and hematological 

variables, and were given a complete gross necropsy at termination. There were no effects on 

mortality and no clinical signs other than urine soiling (all doses). Body weight, food 

consumption, and water consumption were reduced at all doses (but not statistically significant), 

with effects greatest in high-dose males (14% lower body weight than controls by study 

termination). Red blood cell count (RBC), hemoglobin (Hgb), and hematocrit (Hct) were 

decreased (not statistically significant) relative to controls at all doses, with the largest decreases 

observed in the 2,000 mg/kg-day dose group (all variables: 10–13% decrease for males, 5–7% 

decrease for females). The data showed no statistically significant treatment-related adverse 

effects on serum chemistry, urinalysis, or organ weights. The only effect observed at gross 

necropsy was urine staining of the perineal fur (all doses). 
Table 1. Summary Table for Diethylene Glycol Monobutyl Ether

Species and 

Study Type Exposure Critical Effects NOAEL LOAEL Comments 

Oral Studies Rat Short-term 

0, 51, 94, 210, 650, 970, or 1,830 mg/kg-day in drinking water for 30 days. 

Reduced water intake at ≥94 mg/kg-day and histopathological changes in the liver (congestion and slight cloudy swelling) and kidney tubules (cloudy swelling and increased secretion) at 

≥650 mg/kg-day. 

Cannot be determined 

Cannot be determined 

Endpoints included mortality, water intake, body weight and histology of the liver, kidneys, spleen, and testis. Small groups of rats (five/sex/dose). This is a range-finding study that is limited in scope and inadequately reported. 

Rat Subchronic 

891, 1,782, or 3,564 mg/kg-day by gavage on 5 days/week, for 6 weeks. (average daily dose: 636, 1,273, or 2,546 mg/kg-day) 

Hyperkeratosis of the stomach at ≥891 mg/kg-day and hematologic effects (reduced erythrocyte count, hemoglobin concentration, and MCHC and increased MCV) at ≥1,782 mg/kg-day. 

891 mg/kg-day (average daily dose of 636 mg/kg-day) 

1,782 mg/kg-day (average daily dose of 1,273 mg/kg-day) Hematological effects 

Endpoints included clinical signs, fooconsumption, body weight, hematology, clinical chemistry, organ weights, gross pathology, and histopathology. Human relevance of gastric hyperkeratosis is questionable due to bolus exposure and likely direirritant properties of DGBE. 

Rat Subchronic 

0, 70, 330, or 1,630 mg/kg-day (males) or 0, 50, 250, or 1,270 mg/kg-day (females) by gavage on 5 days/week, for 13 weeks. (Average daily dose: 0, 50, 236, or 1,164 mg/kg-day [males]; 0, 36, 179, or 907 mg/kg-day [females].) 

Decreased total WBC and lymphocyte counts and MCHC in females at ≥50 mg/kg-day. Mortality in both sexes at ≥250/330 mg/kg-day (some of the deaths were due to gavage error). 

N/A 50 mg/kg-day 

(average daily dose of 36 mg/kg-day) Lymphopenia 

Endpoints included clinical signs, fooconsumption, body weight, hematology, clinical chemistry, urinalysis, organ weights, gross pathology, and histopathology. The histological examinations did not include the stomach. 

4
Table 1. Summary Table for Diethylene Glycol Monobutyl Ether

Species and 

Study Type Exposure Critical Effects NOAEL LOAEL Comments 

Rat Subchronic 

0, 50, 250, or 1000 mg/kg-day in drinking water for 13 weeks. 

Decreased RBC count, hemoglobin, and hematocrit at ≥250 mg/kg-day. Other effects only occurred at 1,000 mg/kg-day and mainly involved the liver; these included increases in organ weight and hepatic cytochrome P450s and UGT levels, decreases in serum total protein, cholesterol, and serum AST, and hepatocyte hypertrophy and individual hepatocyte degeneration. 

50 mg/kg-day 250mg/kg-day 

Reduced RBC count and Hgb in both sexes. 

Endpoints included clinical signs, fooand water consumption, body weighthematology, clinical chemistry, urinalysis, functional observational battery, sperm analysis, liver metabolenzymes, organ weights, gross pathology, and histopathology. Study possibly useful for subchronic and chronic p-RfD derivation. 

Rat Reproductive 

0, 250, 500, or 1,000 mg/kg-day by gavage in water. Untreated males were bred to treated females and vice versa. The males were treated for 60 days prior to mating, and the females were treated from 14 days prior to mating to GD 13 or until weaning of the offspring. 

Slight reduction in pup weight during the last week of lactation in the offspring of the females dosed with 1,000 mg/kgday. No maternal toxicity or reproductive effects were observed. 

500 mg/kg-day 

1,000 mg/kg-day Decreased pup weight during the last week of lactation. 

Endpoints included body weight in both sexes; numbers of corpora luteaimplants, resorptions, viable embryos and live or dead pups; and clinical condition, body weight, and external development in pups. 

5
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Table 1. Summary Table for Diethylene Glycol Monobutyl Ether

Species and 

Study Type Exposure Critical Effects NOAEL LOAEL Comments 

Rat Reproductive and Developmental 

0, 25, 115, or 633 mg/kg-day in diet on GD 0–20. Rats were killed on GD 20 for uterine and fetal examinations or allowed to deliver 

pups that were reared 

until 10 weeks of age. 

Reduced maternal body weight gain during pregnancy at ≥25 mg/kgday (the effect was variable and not doserelated). No prenatal or postnatal developmental toxicity. 

633 mg/kg-day 

Maternal endpoints included clinical signs, food consumption, and body weight. Developmental endpoints included numbers of corpora lutea, implantations, litters, and live fetuses per litter, pre and postimplantation losses, fetal and placental weights, seratio, external and oral cavity anomalies, skeletal and internal anomalies, gestation length, numbers of live newborns, and pup body weigand survival. 

Mouse Developmental 

0 or 500 mg/kg-day gavage in water on GD 7–14. 

No maternal or developmental toxicity. 

500 mg/kg-day 

Developmental toxicity screening assay. Study endpoints included maternal body weight, fetal survival, pup perinatal and postnatal survival, and pup body weight. Teratogenicity not evaluated. 

Inhalation Study Rat Subchronic 

0, 2, 6, or 18 ppm (0, 13, 40, or 119 mg/m3) for 

6 hours/day, 5 days/week, for up to 22 exposures in 5 weeks. (average daily concentration: 0, 2.3, 7.1, or 21.3 mg/m3) 

Liver histopathology (slight hepatocyte vacuolization consistent with fatty change) in females at ≥6 ppm and gross paleness of the liver in 3/10 females at 18 ppm. 

2 ppm (average concentration of 13 mg/m3)

6 ppm (average concentration of 40 mg/m3) 

Endpoints included clinical signs, boweight, hematology, clinical chemistrurinalysis, organ weights, gross pathology, and histopathology. Histological exams were comprehensive at 0 and 18 ppm but limited to the liver at 2 and 6 ppm. Study used to derive subchronic and chronic RfCs in previous PPRTV document. 
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 In an earlier drinking water study, groups of Sherman rats (5/sex/group) were exposed to 

DGBE (reported as butyl carbitol) in drinking water at reported doses of 0, 51, 94, 210, 650, 970, 

or 1,830 mg/kg-day for 30 days (Smyth, 1940; Smyth and Carpenter, 1948). The study was 

summarized only briefly. Endpoints included mortality, water intake, body weight, and 

histology of the liver, kidney, spleen, and testis. No effects were observed at 51 mg/kg-day. 

Deaths did not occur at any dose. Water intake was reduced at ≥94 mg/kg-day, but growth was 

not statistically significantly affected at any dose level. Histopathological changes occurred at 

≥650 mg/kg-day that included cloudy swelling and increased secretion in the kidney tubules and 

congestion and slight cloudy swelling of the liver. The limited scope and inadequate reporting of 

this study precludes identification of a NOAEL or LOAEL. 

Subchronic Studies—Groups of 10 male albino rats (Charles River COBS, CD, BR) 

were administered 0, 891, 1,782, or 3,564 mg/kg-day doses of undiluted DGBE (>99.5% pure) 

by gavage on 5 days/week, for 6 weeks (Kodak, 1984). The group of control rats was treated 

with distilled water equal in volume to that of the highest-dose group. Endpoints included 

clinical condition, food consumption and body weight, measured hematology values (Hgb, Hct, 

RBC, and total and differential white blood cell counts [WBC]) and calculated red blood cell 

indices (mean corpuscular volume [MCV], mean corpuscular hemoglobin [MCH], and mean 

corpuscular hemoglobin concentration [MCHC]), several serum chemistry indicators of liver 

damage and kidney function (alanine aminotransferase [ALT or SGPT], aspartate 

aminotransferase [AST or SGOT], alkaline phosphatase [ALP], lactate dehydrogenase [LDH], 

blood urea nitrogen [BUN], creatinine, and glucose), organ weights (liver, kidneys, heart, testes, 

brain, and spleen), and comprehensive gross and histopathology (~30 tissues). Deaths attributed 

to gavage error occurred in the high- (4/10), mid- (2/10), and low- (1/10) dose groups. An 

additional two rats in the high-dose group were killed in moribund condition. Effects observed 

in the high-dose group included clinical signs of toxicity (dyspnea, prostration, and unkempt hair 

coat, as well as bloody urine and blood around the nose and mouth in one animal), reduced 

body-weight gain and food consumption, and histopathological changes in the spleen 

(congestion, red pulp hypocellularity, and hemosiderin-like pigmentation). Statistically 

significant effects occurring in the mid- and high-dose groups included hematologic indications 

of RBC damage (clear dose-related reductions in RBC [14% and 28%], Hgb [11% and 15%], 

and MCHC [9% and 18%] and increases in MCV [14% and 45%] and MCH [18% at high dose 

only], with no effect on total or differential WBC), decreased serum glucose concentration, 

increased absolute and relative liver and spleen weights, and histopathological changes in the 

kidneys (hyaline droplet degeneration, proteinaceous casts, and hemosiderin in the proximal 

tubules). The investigators noted that the renal proteinaceous casts and hemosiderin appeared to 

be compound-related but may have been secondary to the hematological effects, and that the 

significance of the hyaline droplet degeneration is uncertain because it was also seen in all 

10 control rats. The only clear treatment-related effect at the low dose was hyperkeratosis of the 

stomach (10/10), which was also observed in the mid- and high-dose groups (10/10 and 8/10 

incidences, respectively) compared to 0/10 in the controls. The human relevance of the gastric 

hyperkeratosis is questionable due to the bolus (undiluted gavage) type of exposures and likely 

direct irritant properties of DGBE, as indicated by aspiration-related respiratory tract lesions 

after gavage exposure in the Hobson et al. (1987) study summarized below. Based on the 

hematological effects, the NOAEL and LOAEL of 891 and 1,782 mg/kg-day were identified, and 

the corresponding average daily doses were 636 and 1,273 mg/kg-day. 

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 A 13-week study was conducted, in which groups of 16 male and 16 female F344 rats 

were treated with DGBE (99% pure) by gavage in water (0.2% of body weight) 5 days/week 

(Hobson et al., 1987). Reported dose levels were 0, 70, 330, and 1,630 mg/kg-day in the males 

and 0, 50, 250, and 1,270 mg/kg-day in the females. An interim sacrifice was conducted at 

6 weeks, after which all groups except the high dose consisted of 10 rats/sex; the high-dose 

group had 4 rats/sex due to early mortality, as discussed below. Endpoints included clinical 

condition, food consumption and body weight, hematology (RBC and platelet counts, Hgb, Hct, 

MCV, MCHC, and total and differential WBC), serum chemistry (11 indices, including BUN), 

urine chemistry (14 indices, including blood in urine), organ weights, and comprehensive gross 

and histopathology (~30 tissues, but not including stomach). A dose-related increase in 

mortality occurred in mid- and high-dose rats of both sexes. Deaths occurred in mid-dose males 

starting in Week 8, mid-dose females starting in Week 4, and high-dose rats of both sexes 

starting in Week 1. Surviving to termination were 10, 10, 4, and 2 males and 10, 9, 8, and 1 

females in the control, low-, mid-, and high-dose groups, respectively. The small numbers of 

survivors in the high-dose group precluded statistical analysis of endpoints evaluated at 

termination for this group. 

 Other effects in the mid- and high-dose rats of both sexes included inflammatory lesions 

in the respiratory tract (Hobson et al., 1987). Respiratory tract lesions that were commonly 

observed in the females at ≥250 mg/kg-day and males at ≥330 mg/kg-day included acute rhinitis, 

laryngitis, tracheitis, pulmonary congestion, and edema. Mild squamous metaplasia of the nasal 

epithelium was found in two mid-dose and five high-dose males and one high-dose female. The 

study authors suggested that these effects, as well as sporadic observations of foreign body 

pneumonia and acute pleuritis, indicated that gavage-related aspiration of DGBE may have 

occurred, contributing to mortality in some of the rats. Study authors reported that over the 

entire study, respiratory tract lesions in three males and six females in the mid-dose groups and 

four males and five females in the high-dose groups were consistent with gavage accident. 

Although a number of the deaths were clearly attributable to gavage error, the dose-related 

distribution of the mortality suggests that some deaths were related to systemic toxicity. 

Histological findings in the other mid- and high-dose rats (i.e., unclear pathogenesis of 

respiratory lesions) and experimental observations by the dosing technicians also did not 

completely support a definitive relationship between gavage procedure and mortality. The lack 

of histological examinations of the stomach precludes possibly corroborating the gastric 

hyperkeratosis observed in the Kodak (1984) 6-week study. Gastric hyperkeratosis is not 

necessarily expected to have been induced in the Hobson et al. (1987) study due to testing of 

lower doses diluted in water compared to higher undiluted doses in the Kodak (1984) study 

discussed above. 

 Hematological effects after 13 weeks of DGBE exposure (Hobson et al., 1987) included 

dose-related decreases in total WBC and lymphocyte counts in surviving females at 50 (by 

32.1% in WBC and 34.9% in lymphocyte counts) and 250 mg/kg-day (by 42.6% in WBC and 

47.9% in lymphocyte counts) (see Table 2). No decrease in these parameters was seen in the 

lone surviving female at 1,270 mg/kg-day or in males at any dose at 13 weeks. RBC and Hgb 

were not significantly reduced at any dose in either sex at 13 weeks. However, the researchers 

reported that at 6 weeks (i.e., before all the animals died), RBC and Hgb were significantly 

decreased (data not reported) in high-dose males and females, and there were dose-dependent 

hemoglobinuria in males and females exposed to DGBE. In addition, lymphocyte counts were 

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significantly decreased (data not reported) in high-dose DGBE males; nevertheless, neutrophil 

counts were significantly increased in mid- and high-dose males at the interim sacrifice. At 

13 weeks, nonhematological effects of DGBE that were statistically significant (p < 0.05) 

included increased absolute liver weight in males at ≥70 mg/kg-day, increased relative liver 

weight in males at 330 mg/kg-day, increased BUN and serum alkaline phosphatase values in 

males at ≥330 mg/kg-day, and increased hyaline droplet formation in the renal tubular epithelium 

in females at 1,270 mg/kg-day. The hyaline droplets in females were considered by the 

researchers to be indicative of hemoglobinuria. Renal hyaline droplets were commonly observed 

in males of all dose and control groups and thereby regarded as being within normal physiologic 

limits. The available data indicate that hematological changes are the most sensitive effects of 

DGBE and that the 50 mg/kg-day-dose (average daily dose of 36 mg/kg-day) is most 

appropriately classified as a LOAEL based on the induction of lymphopenia in only females, and 

no lymphopenia in the high-dose female rat. MCHC was also reduced in females at 

≥50 mg/kg-day, but is not clearly reflective of an adverse effect in the absence of significant 

decreases in RBC count, Hgb, or Hct. 

Table 2. Data Sets for Leukopenia/Lymphopenia in Female Rats 

Exposed to DGBE by Gavage for 13 Weeksa

Duration-Adjusted Doseb (mg/kg-day) 

Endpoint (Mean ± SD) 0 35.7 178.6 

Number Examined 7 9 7 

WBC count (× 103/mm3) 4.46 ± 0.73 3.03 ± 0.59c

(67.9%)d

2.56 ± 0.31c

(57.4%) 

Lymphocyte count (× 103/mm3) 3.55 ± 0.6 2.31 ± 0.42c

(65.1%) 

1.85 ± 0.31c

(52.1%) a

Hobson et al. (1987) 

b

Duration-Adjusted Dose = Dose × 5/7 days c

Statistically significantly different from controls (p < 0.05) 

d

Percentage relative to the control 

Note: the results from the high-dose group are not presented because this dose group was dropped from further 

analysis due to high mortality. 

Johnson et al. (2005) administered DGBE (99.2% pure) in drinking water daily to groups 

(10 rats/sex) of F344 rats at doses of 0, 50, 250, or 1,000 mg/kg-day for 13 weeks. Animals were 

observed twice daily for general appearance and were assessed weekly for detailed clinical 

evaluation, body weight, and food and water consumption. Ophthalmological examinations and 

a behavioral functional observational battery (FOB) were administered before initiation of 

exposure and during the last week of the study. Clinical pathology, including hematology, serum 

chemistry, and urinalysis, was assessed prior to sacrifice and necropsy. Evaluations of sperm 

and liver metabolic enzymes were also made prior to necropsy. A comprehensive necropsy, 

including organ-weight determinations, was conducted for all rats. Comprehensive 

histopathological evaluations were made for control and high-dose rats, with subsequent 

evaluations of selected organs from the low- and mid-dose groups as indicated based on findings 

at the high dose. 

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All rats survived to study termination, and other than some urine soiling in the high-dose 

group, there were no treatment-related clinical signs (Johnson et al., 2005). Decreases in water 

and food consumption and concomitant reductions in body weight were noted in both sexes at 

the high dose throughout the study, with the greatest reductions noted in males; body weight at 

the end of the study was reduced by 10% and 6% in males and females, respectively, in 

comparison with controls. Water consumption among high-dose rats was 7–8% less than 

controls throughout the study, and weekly food consumption among high-dose rats was 

5–11% less than control values. No treatment-related effects on body weight or food or water 

consumption were noted in the mid- and low-dose groups. No treatment-related effects on 

urinalysis, ophthalmology or the FOB (sensory evaluation, rectal temperature, grip performance, 

motor activity) were observed for any treatment group (data not shown). There were small, but 

statistically significant, dose-related decreases in RBC hematological variables relative to 

controls in mid- (RBC, Hgb) and high-dose (RBC, Hgb, Hct) rats of both sexes. The decreases 

ranged from 2–4% at 250 mg/kg-day and 5–9% at 1,000 mg/kg-day. RBC counts in both sexes 

and Hgb in males were outside of the historical control range for the similar study in the same 

strain rats (see Table 3). There were no other treatment-related effects on hematological 

variables, including total and differential WBC (data not shown). 

Table 3. Data Sets for Anemia in Male and Female Rats 

Exposed to DGBE in Drinking Water for 13 Weeksa

Dose (mg/kg-day) 

Endpoint 0 50 250 1000 

Males

Number Examined 10 10 10 10 

RBC Count (× 106/μL) 9.27 ± 0.35 9.13 ± 0.22 

(98.5%)c

8.94 ± 0.34b

(96.4%)

8.53 ± 0.31b

(92.0%)

Hgb (g/dL) 15.8 ± 0.5 15.7 ± 0.4 

(99.3%)

15.3 ± 0.4b

(96.8%)

14.8 ± 0.4b

(93.6%)

Females

Number Examined 10 10 10 10 

RBC Count (× 106/μL) 8.26 ± 0.22 8.06 ± 0.31 

(97.6%)

8.07 ± 0.24b

(97.7%)

7.54 ± 0.17b

(91.3%)

Hgb (g/dL) 15.6 ± 0.3 15.2 ± 0.4 

(97.4%)

15.2 ± 0.4b

(97.4%)

14.8 ± 0.3b

(94.9%) a

Johnson et al. (2005) 

b

Statistically significantly different from controls (p < 0.05) c

Percentage relative to the control 

Note: historical control ranges: RBC in males 9.11–9.22 (106

/μL), RBC in females 7.81–8.20 (106/μL), Hgb in males 

15.0–15.6 (g/dL), and Hgb in females 14.3–14.9 (g/dL).

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 Clinical chemistry findings were generally unremarkable, although there were slight 

decreases in serum protein, cholesterol, and AST in high-dose males and females (Johnson et al., 

2005). Liver metabolic enzymes were slightly increased relative to controls only for high-dose 

animals as follows: EROD1

 (24% both sexes); PROD2 (38% males, 24%, females); and UGT3

(17% males, 16% females). Organ weight results included increased absolute and relative liver 

and kidney weights in the high-dose groups of both sexes, despite decreased body weight in 

these groups. Absolute and/or relative spleen weights were increased in all male and female 

dose groups, but the changes from controls were small (absolute weight changes ≤6% in males 

and ≤3% in females) and did not increase with dose. The only treatment-related 

histopathological changes were noted in the livers of high-dose females. These changes were 

considered by the study authors to be treatment-related and consisted of hepatocellular 

hypertrophy (“very slight” in 6/10 high-dose females) and foci of necrotic cells in the 

centrolobular region. No treatment-related histopathologic changes were noted in any other 

tissues, including kidneys, spleen, or male livers (incidence data not shown). No treatmentrelated effects on sperm count, sperm morphology or motility, or male reproductive tissues were 

observed (data not shown). The NOAEL for this study is 50 mg/kg-day, and the LOAEL is 

250 mg/kg-day on the basis of significantly reduced RBC count and Hgb in both sexes. 

Chronic Studies—No pertinent data were located regarding health effects of DGBE in 

animals following chronic oral exposure. 

Reproductive/Developmental Studies—Groups of 25 male and 25 female Charles River 

CD rats were treated with 0, 250, 500, or 1,000 mg/kg-day doses of DGBE (95% pure) by 

gavage in distilled water in a fertility study (Nolen et al., 1985). Untreated males were bred to 

treated females and vice versa. The males were treated for 60 days prior to mating, and the 

females were treated from 14 days prior to mating to gestation day (GD) 13 or until weaning of 

the offspring. Half of the females in each group were killed on GD 13 for examination of uterine 

contents. The remaining females delivered their young, and the offspring were followed to 

weaning. Endpoints included weekly body weight in both sexes; numbers of corpora lutea, 

implants, resorptions, viable embryos, and live or dead pups; and clinical condition, body weight, 

and external development in pups. The only effect attributable to treatment was a slight (~8%) 

reduction in pup weight during the last week of lactation among the offspring of the females 

dosed with 1,000 mg/kg-day relative to the control group. The NOAEL and LOAEL were 500 

and 1,000 mg/kg-day, respectively, for mild toxicity to the neonate, while no maternal or 

reproductive effects were found even at 1,000 mg/kg-day. 

These findings are supported by another study in rats. Groups of 19–21 pregnant Wistar 

rats were fed DGBE (purity not reported) in the diet at reported intake levels of 0, 25, 115, or 

633 mg/kg-day from GD 0–20 (Ema et al., 1988). Some 14 or 15 rats in each group were killed 

on Day 20 for uterine and fetal examinations, and the remaining 5 or 6 rats/group were allowed 

to deliver spontaneously. From each litter, eight pups (four rats/sex) were reared until 10 weeks 

of age. Maternal endpoints included daily evaluations for clinical signs of toxicity and effects on 

food consumption and body weight. Prenatal developmental endpoints included numbers of 

litters and corpora lutea, implantations and live fetuses per litter, pre and postimplantation losses, 

1

Ethoxyresorufin-O-dealkylase. 

2

Pentoxyresorufin-O-dealkylase. 

3

UDP-glucuronosyltransferase. 

 11
FINAL 

9-28-2009 

fetal and placental weights, sex ratio, external and oral cavity anomalies, skeletal anomalies 

(approximately half the fetuses in each litter), and internal anomalies (remaining half of the 

fetuses). Postnatal endpoints included gestation length, numbers of live newborns, body weight, 

and survival rate. The only effect noticed was reduced maternal body-weight gain during 

pregnancy at ≥25 mg/kg-day (12–18% lower than controls, p < 0.05). The maternal body-weight 

gain is not only affected by changes in the maternal weight, but also changes in litter size and 

fetal body weight. Thus, the absolute changes in maternal weight during the gestation in any of 

these dose groups were relatively small (<10%); therefore, this endpoint is not considered to be 

an adverse response to DGBE. The lack of any prenatal or postnatal developmental effects 

indicates that 633 mg/kg-day is a NOAEL for developmental as well as maternal toxicity. 

Similar results were found in mice. In a screening assay, groups of 50 pregnant CD-1 

mice were treated with 0 or 500 mg/kg doses of DGBE (>99% pure) by gavage in aqueous 

solution on GD 7–14 (Schuler et al., 1984). Study endpoints included pup survival in utero 

(percent of live litters/pregnant survivors), pup perinatal and postnatal survival (numbers of live 

and dead pups per litter and pup survival to age 2.5 days), and pup body weights (at birth and age 

2.5 days). Maternal indices included body weight on GD 7, GD 18, and Day 3 postpartum. No 

treatment-related maternal, litter, or pup effects were observed, indicating that 500 mg/kg-day 

was a NOAEL for maternal and developmental toxicity. 

Inhalation Exposure 

Chronic Studies—No pertinent data were located regarding health effects of DGBE in 

animals following chronic inhalation exposure. 

Subchronic Studies—In a subchronic inhalation study, groups of 15 male and 15 female 

F344 rats were exposed to 0, 2, 6, or 18 ppm of DGBE (~98.6% pure) (0, 13, 40, or 119 mg/m3

) 

for 6 hours/day, 5 days/week, for up to 22 exposures in 5 weeks (Gushow et al., 1984). The 

119 mg/m3 concentration was the highest sustainable vapor concentration of DGBE due to its 

low vapor pressure. RBC fragility was tested in five rats/sex/group following the 15th exposure. 

Endpoints evaluated in the remaining 10 rats/sex/group included clinical condition and body 

weight throughout the study; hematology (RBC, total and differential WBC, platelet, Hgb, Hct, 

MCV, MCH, and MCHC), clinical chemistry (BUN, ALT, ALP, glucose, albumin, total protein, 

and total bilirubin), and urine chemistry (pH, glucose, ketones, bilirubin, urobilinogen, occult 

blood, protein, and specific gravity) after 16–22 exposures; and organ weights, gross pathology, 

and histology after 22 exposures. The histological examinations were comprehensive (including 

nasal turbinates and lungs) in the control and 119-mg/m3 groups but limited to the liver in the 

13 and 40 mg/m3 groups because this was the only tissue identified as a possible target at 

119 mg/m3

. Statistically significant (p < 0.05) decreases in serum glucose levels were observed 

in females at ≥13 mg/m3 and males at ≥40 mg/m3, but the reductions were minimal 

(9–16% lower than controls). Other effects included slightly increased WBC in males (within 

the normal range of variation and not considered to be toxicologically significant by the 

researchers) and hepatic changes in both sexes at ≥40 mg/m3

. The hepatic effects included 

inconsistent changes in relative liver weight (decreased in males at ≥40 mg/m3

 and increased in 

females at 119 mg/m3

) with no effects on absolute liver weight, histological changes (slight 

hepatocyte vacuolization consistent with fatty change) in females at ≥40 mg/m3

, and gross 

paleness of the liver in 3/10 females at 119 mg/m3

. Hepatocytes in the control and 13-mg/m3 

females were vacuolated to a somewhat lesser degree (i.e., very slight), indicating that the effect 

 12
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9-28-2009 

was only minimally increased at ≥40 mg/m3

. No other biologically significant effects were 

found in clinical chemistry, hematology, and urine chemistry. The low concentration of DGBE 

(i.e., 13 mg/m3) is identified as a NOAEL based on significant incidences of hepatocyte 

vacuolization consistent with fatty change that occurred in female rats at ≥40 mg/m3 and is 

supported by increased relative liver weight at 119 mg/m3. 

Other Studies 

There are several short-term and subchronic toxicity studies (Proctor and Gamble, 1982; 

Bio/dynamics, Inc., 1989; Auletta et al., 1993), reproductive toxicity studies (Bio/dynamics, Inc., 

1989; Auletta et al., 1993), and a subchronic neurotoxicity study (Bio-Research Labs, 1989) with 

dermal exposure to DGBE. The only changes reported by Proctor and Gamble (1982) 

(hematological parameters or hematuria) and those (occult blood in the urine) reported by 

Bio/dynamics, Inc. (1989) and Auletta et al. (1993) are consistent with the effects that were seen 

following oral exposure to DGBE. 

Genotoxicity

The genotoxicity of DGBE has been tested in several assay systems. In in vitro assays, 

DGBE was negative for reverse mutation in Salmonella typhimurium with or without metabolic 

activation (Thompson et al., 1984; Zeiger et al., 1992), unscheduled DNA synthesis in primary 

rat hepatocytes (Thompson et al., 1984), forward mutation (HGPRT locus) in Chinese hamster 

ovary (CHO) cells (Gollapudi et al., 1993) and sister chromatid exchanges (SCEs) in CHO cells 

(Thompson et al., 1984), and weakly positive for forward mutations in mouse lymphoma 

L5178Y cells (Thompson et al., 1984). In vivo testing of DGBE was negative for sex-linked 

recessive mutations in Drosophila melanogaster (Thompson et al., 1984) and induction of 

micronuclei in bone marrow cells of mice (Gollapudi et al., 1993). 

DERIVATION OF PROVISIONAL SUBCHRONIC AND CHRONIC ORAL RfD 

VALUES FOR DIETHYLENE GLYCOL MONOBUTYL ETHER (DGBE)

Information on systemic effects of repeated oral exposure to DGBE is available from 

several studies in rats as summarized in Table 1. A range-finding study was conducted in which 

groups of five Sherman rats/sex that were exposed to DGBE in drinking water at doses ranging 

from 51–1,830 mg/kg-day for 30 days had no effects at 51 mg/kg-day, reduced water intake at 

≥94 mg/kg-day, and histopathological changes in the kidneys and liver at ≥650 mg/kg-day 

(Smyth, 1940; Smyth and Carpenter, 1948). However, a NOAEL or LOAEL cannot be 

identified from this study due to the limited number of endpoints investigated and inadequate 

reporting. A more recent range-finding study (Johnson et al., 2005) in which groups of five 

F344 rats/sex were exposed to DGBE in drinking water at doses of 1,000–2,000 mg/kg-day for 

2 weeks noted decreases in body weight, food consumption, water consumption, RBC, Hgb, and 

Hct; however, these changes were not statistically significant in comparison with controls. 

In a 6-week gavage study, DGBE was administered to groups of 10 male rats (Charles 

River COBS, CD, BR) at doses of 891, 1,782, or 3,564 mg/kg-day on 5 days/week, for 6 weeks 

(Kodak, 1984). This study identified a hematological NOAEL and LOAEL of 891 and 

1,782 mg/kg-day (average daily doses of 636 and 1,273 mg/kg-day), respectively, based on 

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9-28-2009 

reduced RBC counts, Hgb, and MCHC, with related effects, including bloody urine, occurring at 

3,564 mg/kg-day. In a 13-week study, DGBE was administered to groups of 16 male/16 female 

F344 rats at doses of 0, 50/70, 250/330, or 1,270/1,630 mg/kg-day by gavage on 5 days/week, for 

13 weeks (Hobson et al., 1987). This study identified a LOAEL of 50 mg/kg-day (36 mg/kg-day 

average daily dose) based on leukopenia and lymphopenia in female rats (Hobson et al., 1987). 

In a subsequent comprehensive drinking water study, groups of 10 F344 rats/sex were exposed to 

DGBE in drinking water at doses of 0, 50, 250, and 1,000 mg/kg-day for 13 weeks 

(Johnson et al., 2005). This study identified a NOAEL of 50 mg/kg-day based on the critical 

effects of decreased RBC and Hgb noted at doses ≥250 mg/kg-day in both sexes. 

The database for DGBE consistently indicates that hematological effects are the most 

sensitive response after exposure to DGBE; however, inconsistent findings of leukopenia and 

lymphopenia were reported. The most sensitive response was reported from Hobson et al. 

(1987), noting leukopenia and lymphopenia in female rats at doses ≥50 mg/kg-day. However, 

confidence in the findings from this study (Hobson et al., 1987) is limited by gavage accidents 

and associated deaths and respiratory tract pathology in the higher-dose groups 

(≥250 mg/kg-day), inconsistent changes in neutrophil counts (increased) in male rats in the midand high-dose groups (≥330 mg/kg-day) at 6 weeks, lack of leukopenia and lymphopenia in male 

rats at the end of experiment, and no similar responses in the lone surviving female in the 

highest-dose group. In addition, neither leukopenia nor lymphopenia was reported in the same 

strain of rats in a 13-week drinking water study (Johnson et al., 2005) at dose levels as high as 

1,000 mg/kg-day, and no similar effects were found in male rats in the Kodak (1984) 6-week 

gavage study (females were not tested) at doses at high as 3,564 mg/kg-day. In contrast to the 

leukopenia and lymphopenia responses, consistent findings of decreased RBC counts and Hgb 

levelswere reported in male rats treated with DGBE ≥1,782 mg/kg-day for 6-weeks (Kodak, 

1984), in male rats at a dose level of 1,630 mg/kg-day and female rats at a dose level of 

1,270 mg/kg-day at 6-week interim sacrifice (Hobson et al., 1987), and in both sexes at a dose 

level of ≥250 mg/kg-day at the end of 13 weeks of exposure (Johnson et al., 2005). Therefore, 

decreased RBC counts and Hgb levels after 13-weeks of treatment (Johnson et al., 2005) are 

considered the critical effects for the DGBE-induced hematological response. 

There are no indications that reproductive or developmental toxicity are effects of 

concern for DGBE, as multiple oral studies in rats and mice found no effects at doses of 500 and 

633 mg/kg-day and only a mild, transitory effect on neonate weight at 1,000 mg/kg-day 

(Ema et al., 1988; Nolen et al., 1985; Schuler et al., 1984). Dermal studies also showed no 

effects on reproduction or development (Bio/dynamics, Inc., 1989; Nolen et al., 1985). 

There are no indications that neurotoxicities are effects of concern for DGBE, as an oral 

study in rats found no behavioral effects at doses of 50, 250, or 1,000 mg/kg-day (Johnson et al., 

2005). A dermal study in rats also showed no effects on a number of neurotoxicity endpoints at 

doses corresponding to 200, 600, or 2,000 mg/kg (Bio-Research Labs, 1989). 

Based on the available studies, a potential point of departure (POD) for the derivation of 

p-RfD values is the NOAEL of 50 mg/kg-day for anemia (reduced RBC counts, Hgb levels) in 

rats (Johnson et al., 2005). The relevant data sets for these endpoints are shown in Table 3 

(Johnson et al., 2005). No duration adjustment for the doses for the Johnson et al. (2005) study 

is needed because exposure was continuous for the duration of the study. Attempts to apply 

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BMD modeling to the data sets (RBC and Hgb) were successful for RBC counts in male (BMDL 

of 81 mg/kg-day) and female rats (BMDL of 280 mg/kg-day), and for Hgb levels in males 

(BMDL of 328 mg/kg-day) but not in females (Appendix A). Because the changes in Hgb levels 

in females were comparable to those in males (see Table 3), a POD for this endpoint is expected 

to be at the range close to 328 mg/kg-day based on male data. Thus, the lowest BMDL of 

81 mg/kg-day (based on changes in RBC in male rats) among these four endpoints is the most 

sensitive POD and is used for the derivation of both subchronic and chronic p-RfDs. 

Subchronic p-RfD 

A subchronic p-RfD is derived by applying an UF of 300 to the BMDL of 81 mg/kg-day 

as follows: 

 Subchronic p-RfD = BMDL ÷ UF 

 = 81 mg/kg-day ÷ 300 

 = 0.3 mg/kg-day or 3 × 10-1 mg/kg-day

The composite UF is composed of the following factors: 

• An UF of 10 is applied for interspecies extrapolation to account for potential 

pharmacokinetic and pharmacodynamic differences between rats and humans. 

• An UF of 10 for intraspecies differences is used to account for potentially 

susceptible individuals in the absence of information on the variability of 

response in humans. 

• An UF for LOAEL to NOAEL extrapolation is not needed because the POD is a 

BMDL. 

• An UF of 3 (100.5) is applied to account for deficiencies in the database. The 

database includes three subchronic studies in rats, two developmental studies in 

rats and mice, and a one-generation reproductive study in rats; however, the 

database lacks a multigeneration reproduction study and a subchronic study in a 

second species. 

Confidence in the principal study (Johnson et al., 2005) is medium. The principal study 

was well designed, and the critical effect is consistent with that found from other short-term 

studies (Kodak, 1984, Hobson et al., 1987). However, a 13-week study (Hobson et al., 1987) 

exposed F344 rats to DGBE by gavage for the same duration but did not identify a similar 

response at the end of treatment at doses of ≥1,270 mg/kg-day. Instead, this study (Hobson et 

al., 1987) identified a LOAEL of 50 mg/kg-day for decreased WBC and lymphocyte counts, 

which were not shown by Johnson et al. (2005). Confidence in the database is medium due to 

some inconsistent findings from subchronic studies (Johnson et al., 2005; Hobson et al., 1987), a 

lack of a multigeneration reproduction study, and a subchronic study conducted in a second 

species. Overall, confidence in the subchronic p-RfD is medium. 

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Chronic p-RfD

A chronic p-RfD is similarly derived by applying an UF of 3,000 to the BMDL of 81 

mg/kg-day as follows: 

 Chronic p-RfD = BMDL ÷ UF 

 = 81 mg/kg-day ÷ 3,000 

 = 0.03 mg/kg-day or 3 × 10-2 mg/kg-day

The composite UF is composed of the following factors: 

• An UF of 10 is applied for interspecies extrapolation to account for potential 

pharmacokinetic and pharmacodynamic differences between rats and humans. 

• An UF of 10 for intraspecies differences is used to account for potentially 

susceptible individuals in the absence of information on the variability of 

response in humans. 

• An UF for LOAEL to NOAEL extrapolation is not needed because the POD is a 

BMDL. 

• An UF of 10 was applied for using a study with a subchronic duration of exposure 

to approximate chronic exposure. 

• An UF of 3 (100.5) is applied to account for deficiencies in the database. The 

database includes critical three subchronic studies in rats, two developmental 

studies in rats and mice, and a one-generation reproductive study in rats; however, 

the database lacks a multigeneration reproduction study and a subchronic study in 

a second species. 

Confidence in the principal study (Hobson et al., 1987) is medium, as discussed above for 

the subchronic p-RfD. Confidence in the database is low due to some inconsistent findings from 

subchronic studies (Johnson et al., 2005; Hobson et al., 1987), a lack of chronic oral toxicity 

studies, and a lack of a multigeneration reproduction study. Overall, confidence in the chronic 

p-RfD is low. 

DERIVATION OF PROVISIONAL SUBCHRONIC AND CHRONIC INHALATION RfC 

VALUES FOR DIETHYLENE GLYCOL MONOBUTYL ETHER (DGBE)

Information on the inhalation toxicity of DGBE is essentially limited to the results of a 

comprehensive 5-week study in which rats were exposed to 0, 13, 40, or 119 mg/m3

 for 

6 hours/day, 5 days/week (Gushow et al., 1984). Evaluation of clinical condition, body weight, 

hematology (including RBC fragility), blood and urine chemistry, organ weights, gross 

pathology, and histopathology (including nasal turbinates and lungs) showed slight hepatocyte 

vacuolization consistent with fatty change in female rats at concentrations of ≥40 mg/m3

 and 

increased relative liver weight and gross paleness of the liver at 119 mg/m3

. Although there 

were inconsistent changes (decreases) in relative liver weight in male rats at ≥40 mg/m3

 with no 

effects on absolute liver weight, the hepatic changes in female rats were considered biologically 

significant. The potential for hepatic effects is supported by findings of increased liver weight 

and congestion/cloudy swelling in subchronic oral studies of DGBE (Johnson et al., 2005; 

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Hobson et al., 1987; Kodak, 1984; Smyth, 1940; Smyth and Carpenter, 1948). Exposure to 

concentrations higher than 119 mg/m3 could not be tested because this level was the maximum 

sustainable vapor concentration of DGBE. 

 Subchronic and chronic p-RfCs were derived for DGBE using the data for hepatocelluar 

vacuolization in female rats (Gushow et al., 1984) based on U.S. EPA (1994b) RfC 

methodology. BMD modeling was conducted for the incidence data on hepatocyte vacuolization 

(Table 4), and detailed results are presented in Appendix A. Because the exposure was not 

continuous in the principal study (Gushow et al., 1984), duration-adjusted concentrations and 

human equivalent concentration (HEC) (see Table 4) were calculated before BMD modeling. 

An example of the duration adjustment calculation is presented below: 

 NOAELADJ = NOAEL mg/m3 × 6/24 hr × 5/7 d 

 = 13 mg/m3 × 6/24 hr × 5/7 d 

 = 2.3 mg/m3 

For purposes of deriving a p-RfC based on extrarespiratory effects, DGBE was treated as a 

Category 3 gas. The HEC was calculated assuming periodicity was attained using the following 

equation: 

 NOAELHEC = NOAELADJ × (Hb/g)A/(Hb/g)H

where (Hb/g)A/(Hb/g)H is the ratio of blood:gas (air) partition coefficients of the chemical in the 

test animals and humans. Since a blood:gas partition coefficient is not available for DGBE in 

humans or rats, a unity value is assumed for the ratio (U.S. EPA, 1994b). The BMD modeling 

resulted in an estimated BMC10ADJ of 1.7 mg/m3 and BMCL10ADJ of 0.32 mg/m3, and the 

BMCL10ADJ is equivalent to BMCL10HEC. 

Table 4. Data Sets for Hepatic Changes in Female Rats 

Exposed to DGBE Through Inhalation for 5 Weeksa

Concentration (mg/m3) 

Endpoint 0 13 40 119 

Females

Duration-adjusted concentration 

(mg/m3

)

b 0 2.3 7.1 21.3 

Human equivalent concentration 

(mg/m3

)

c 0 2.3 7.1 21.3 

# Examined 3 4 10 10 

Slight vacuolization consistent 

with fatty change 3 4 9d 10d

a

Gushow et al. (1984) 

b

The duration adjustment was calculated as follows: 

ConcADJ = 119 mg/m3

 × 6/24 hr × 5/7 d = 21.3 mg/m3 

c

HEC = ConcADJ × (Hb/g)A/(Hb/g)H d

Statistically significantly different from controls (p < 0.05) 

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Subchronic p-RfC

A subchronic p-RfC is derived by applying an UF of 300 to the BMCL10HEC of 

0.32 mg/m3 as follows: 

 Subchronic p-RfC = BMCL10HEC ÷ UF 

 = 0.32 mg/m3

 ÷ 300 

 = 0.001 or 1 × 10-3 mg/m3

The composite UF of 300 is composed of the following factors: 

• An UF of 3 (100.5) is applied for interspecies extrapolation to account for potential 

pharmacodynamic differences between rats and humans. Converting the rat data 

to human equivalent concentrations by the dosimetric equations, accounts for 

pharmacokinetic differences between rats and humans; thus, an UF of 10 for 

interspecies extrapolation was not used. 

• An UF of 10 for intraspecies differences is used to account for potentially 

susceptible individuals in the absence of information on the variability of 

response in humans. 

• An UF of 10 is applied to account for deficiencies in the database. The database 

includes only one 5-week study; it lacks developmental toxicity studies and a 

multigeneration reproduction study. 

• An UFL is not applied, as a BMCL was used as the POD. 

• An UFS is not applied because the principal study is a subchronic study. 

The principal study was well conducted with respect to scope of examinations, numbers 

of animals, and exposure levels but is given low-to-medium confidence because the duration was 

short (5 weeks). Confidence in the database is low due to the lack of a 90-day inhalation study, 

supporting data in a second species, a two-generation reproduction study, and developmental 

toxicity studies. With regard to the latter, however, it is noteworthy that studies by oral and 

dermal exposure did not identify reproductive and developmental endpoints as sensitive for this 

chemical. Low confidence in the subchronic p-RfC follows. 

Chronic p-RfC

A chronic p-RfC is derived by applying an UF of 3,000 to the BMCL10HEC of 

0.32 mg/m3

 as follows: 

 

 Chronic p-RfC = BMCL10HEC ÷ UF 

 = 0.32 mg/m3

 ÷ 3,000 

 = 0.0001 mg/m3

 or 1 × 10-4 mg/m3 

The composite UF of 3,000 is composed of the following factors: 

• An UF of 3 (100.5) is applied for interspecies extrapolation to account for potential 

pharmacodynamic differences between rats and humans. Converting the rat data 

to human equivalent concentrations by the dosimetric equations accounts for 

pharmacokinetic differences between rats and humans; thus, an UF of 10 for 

interspecies extrapolation was not used. 

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9-28-2009 

• An UF of 10 for intraspecies differences is used to account for potentially 

susceptible individuals in the absence of information on the variability of 

response in humans. 

• An UF of 10 is applied for using a study with a subchronic duration of exposure 

to approximate chronic exposure. 

• An UFL was not applied, as a BMCL was used as the POD. 

• An UF of 10 is applied to account for deficiencies in the database. The database 

includes only one 5-week study; it lacks chronic toxicity studies, developmental 

toxicity studies, and a multigeneration reproduction study. 

Confidence in the chronic RfC is low, for the same reasons as described above for the subchronic 

RfC, with the additional shortcoming that chronic studies are also lacking from the database. 

PROVISIONAL CARCINOGENICITY ASSESSMENT FOR 

DIETHYLENE GLYCOL MONOBUTYL ETHER (DGBE) 

Weight-of-Evidence Descriptor 

There is no adequate information on the carcinogenicity of DGBE due to a lack of oral, 

inhalation, or dermal studies in animals longer than 13 weeks in duration (Johnson et al., 2005; 

Bio/dynamics, Inc., 1989; Gushow et al., 1984; Hobson et al., 1987; Kodak, 1984). Genotoxicity 

testing of DGBE in vitro with bacteria and mammalian cells and in vivo with Drosophila and 

mice have yielded negative results in all but one study (a weakly positive mutagenic response in 

mouse lymphoma cells in vitro) (Thompson et al., 1984; Zeiger et al., 1992; 

Gollapudi et al., 1993). Under current U.S. EPA (2005) cancer guidelines, “there is inadequate 

information to assess the carcinogenic potential” of DGBE. 

Quantitative Estimates of Carcinogenic Risk 

Derivation of quantitative estimates of cancer risk for DGBE is precluded by the lack of 

carcinogenicity data. 

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ATSDR (Agency for Toxic Substances and Disease Registry). 2008. Toxicological Profile 

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Auletta, C.S, R.E. Schroeder, W.J. Krasavage and C.R. Stack. 1993. Toxicology of diethylene 

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Kodak. 1984. Toxicity Studies with Diethylene Glycol Monobutyl Ether. III. Six-Weeks 

Repeated Dose Study. Eastman Kodak Co., Rochester, NY. Submitted to U.S. EPA under 

TSCA Section 4, Fiche No. OTS0512376. 

NIOSH (National Institute for Occupational Safety and Health). 2005. NIOSH Pocket Guide to 

Chemical Hazards. Online. http://www.cdc.gov/niosh/npg/. 

Nolen, G.A., W.B. Gibson, J.H. Benedict et al. 1985. Fertility and teratogenic studies of 

diethylene glycol monobutyl ether in rats and rabbits. Fund. Appl. Toxicol. 5:1,137–1,143. 

NTP (National Toxicology Program). 2005. 11th Report on Carcinogens. Online. 

http://ntp.niehs.nih.gov/index.cfm?objectid=32BA9724-F1F6-975E-7FCE50709CB4C932. 

NTP (National Toxicology Program). 2008. Management Status Report. Online. 

http://ntp.niehs.nih.gov/index.cfm?objectid=78CC7E4C-F1F6-975E-72940974DE301C3F. 

OSHA (Occupational Safety and Health Administration). 2008. OSHA Standard 1910.1000 

Table Z-1. Part Z, Toxic and Hazardous Substances. Online. http://www.osha.gov/pls/

oshaweb/owadisp.show_document?p_table=STANDARDS&p_id=9992. 

Proctor and Gamble Co. 1982. 28-Day subchronic percutaneous study of DGBE in rabbits. 

Huntington Research Center. (Cited in SRC, 1992). 

Schuler, R.L., B.D. Hardin, R.W. Niemeier et al. 1984. Results of testing fifteen glycol ethers in 

a short-term in vivo reproductive toxicity assay. Environ. Health Perspect. 57:141–146. 

Smyth, H.F. 1940. Special report on the toxicity of the glycols and their derivatives. 61. 

Repeated doses of butyl “carbitol” by mouth. Mellon Institute of Industrial Research. Submitted 

to U.S. EPA under TSCA Section 4, Fiche No. OTS0512492. 

Smyth, H.F., Jr. and C.P. Carpenter. 1948. Further experience with the range finding test in the 

industrial toxicology laboratory. J. Ind. Hyg. Toxicol. 30(1):63–68. 

SRC (Syracuse Research Corporation). 1992. Health and Environmental Effects Document on 

Glycol Ethers. Prepared by Syracuse Research Corporation, Syracuse, NY for the Office of 

Health and Environmental Assessment, Environmental Criteria and Assessment Office, 

Cincinnati, OH. SRC TR-92-007. 

Thompson, E.D., W.J. Coppinger, R. Valencia et al. 1984. Mutagenicity testing of diethylene 

glycol monobutyl ether. Environ. Health Perspect. 57:105–112. 

U.S. EPA. 1984. Health Effects Assessment for Glycol Ethers. Prepared by the Office of 

Health and Environmental Assessment, Environmental Criteria and Assessment Office, 

Cincinnati, OH, for the Office of Emergency and Remedial Response, Washington, D.C. 

EPA/540/1-86-052. 

 21
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U.S. EPA. 1991. Chemical Assessments and Related Activities (CARA). Office of Health and 

Environmental Assessment, Washington, DC. April. 

U.S. EPA. 1994a. Chemical Assessments and Related Activities (CARA). Office of Health and 

Environmental Assessment, Washington, DC. December. 

U.S. EPA. 1994b. Methods of Derivation of Inhalation Reference Concentrations and 

Application of Inhalation Dosimetry. Office of Research and Development, Washington, DC. 

EPA/600/8-90/066F. October. 

U.S. EPA. 1997. Health Effects Assessment Summary Tables. FY-1997 Update. Prepared by 

the Office of Research and Development, National Center for Environmental Assessment, 

Cincinnati OH for the Office of Emergency and Remedial Response, Washington, DC. July. 

EPA/540/R-97/036. NTIS PB97-921199. 

U.S. EPA. 2000. Benchmark Dose Technical Guidance Document [external review draft]. 

EPA/630/R-00/001. Online. http://www.epa.gov/iris/backgr-d.htm. 

U.S. EPA. 2005. Guidelines for Carcinogen Risk Assessment. U.S. Environmental 

Protection Agency, Risk Assessment Forum, Washington, DC. EPA/630/P-03/001B. Online. 

http://www.thecre.com/pdf/20050404_cancer.pdf.

U.S. EPA. 2006. 2006 Edition of the Drinking Water Standards and Health Advisories. Office 

of Water, Washington, DC. EPA 822-R-06-013. Washington, DC. Online. 

http://www.epa.gov/waterscience/drinking/standards/dwstandards.pdf. 

U.S. EPA. 2008. Integrated Risk Information System (IRIS). Online. Office of Research and 

Development, National Center for Environmental Assessment, Washington, DC. 

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

WHO (World Health Organization). 2008. Online catalogs for the Environmental Health 

Criteria Series. Online. http://www.who.int/ipcs/publications/ehc/ehc_alphabetical/en/

index.html. 

Zeiger, E., B. Anderson, S. Haworth et al. 1992. Salmonella mutagenicity tests: V. Results 

from the testing of 311 chemicals. Environ. Mol. Mutagen. 19:2–141. 

 22
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APPENDIX A. DETAILS OF BENCHMARK DOSE MODELING 

FOR SUBCHRONIC/CHRONIC p-RfD 

Model Fitting Procedure for Continuous Data 

The model fitting procedure for continuous data is as follows. The simplest model 

(linear) is first applied to the data while assuming constant variance. If the data are consistent 

with the assumption of constant variance (p ≥ 0.1), then the fit of the linear model to the means is 

evaluated, and the polynomial, power, and Hill models are fit to the data while assuming 

constant variance. Among the models providing adequate fit to the means (p ≥ 0.1), the one with 

the lowest Akaike Information Criterion (AIC) for the fitted model is selected for BMD 

derivation. If the test for constant variance is negative, the linear model is run again while 

modeling the variance as a power function of the mean to account for nonhomogenous variance. 

If the nonhomogenous variance model provides an adequate fit (p ≥ 0.1) to the variance data, 

then the fit of the linear model to the means is evaluated, and the polynomial, power, and Hill 

models are fit to the data and evaluated while the variance model is applied. Among those 

providing adequate fit to the means (p ≥ 0.1), the one with the lowest AIC for the fitted model is 

selected for BMD derivation. If the test for constant variance is negative and the 

nonhomogenous variance model does not provide an adequate fit to the variance data, then the 

data set is considered unsuitable for modeling. If after these attempts, no model provides an 

adequate fit to the data, the highest dose is dropped, if appropriate, and the entire procedure is 

repeated. If no fit is obtained after dropping the highest dose, the next highest dose is dropped, if 

appropriate, and the procedure is repeated. Dose-dropping continues until (1) adequate fit is 

obtained; (2) there are only controls and two dose groups remaining; or (3) it is inappropriate to 

continue dropping doses due to a lack of statistical significance or biologically important 

differences between controls and the remaining treatment groups. If no fit is obtained following 

application of this procedure, then the data set is not considered to be amenable to BMD 

modeling. 

Model Fitting Results for Decreased RBCs in Male Rats (Johnson et al., 2005) 

Following the above procedure, continuous-variable models in the EPA Benchmark Dose 

Software (BMDS) (version 2.1) were fit to the data shown in Table 3 (main text) for decreased 

RBCs in male rats. In the absence of a biologically relevant response level, the benchmark 

response (BMR) was chosen to be 1 standard deviation (SD) from the control mean, as 

recommended by U.S. EPA (2000). As shown in Table A-1, all models provide adequate fit to 

the data. The Hill model (Figure A-1) estimated a significantly lower BMDL than other models; 

therefore, the BMDL for this endpoint is 81 mg/kg-day. 

 23
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Table A-1. Model Predictions for Decreased RBCs in Male Ratsa

Data Set/Model 

Variance 

p-Valueb

Means 

p-Valueb AIC 

BMD1SD 

(mg/kg-day)

BMDL1SD

(mg/kg-day) 

Linear (constant variance)c 0.4943 0.454 -50.5176 440.732 328.449 

2-Degree Polynomial (constant variance)c 0.4943 0.454 -50.5176 440.732 328.449 

3-Degree Polynomial (constant variance)c 0.4943 0.454 -50.5176 440.732 328.449 

Power (constant variance)d 0.4943 0.454 -50.5176 440.732 328.449 

Hill (constant variance)d 0.4943 0.6312 -49.8664 222.31 81.408 a

Johnson et al. (2005)

b

Values <0.10 fail to meet conventional goodness-of-fit criteria 

c

Coefficients restricted to be negative

d

Power restricted to ≥1 

 24
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9.6

Figure A-1. Fit of Hill Model to Data on Decreased RBCs 

in Male Rats (Johnson et al., 2005) 

 ==================================================================== 

 Hill Model. (Version: 2.14; Date: 06/26/2008) 

 Input Data File: C:\USEPA\BMDS21\Data\hilRBCmaleHill.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS21\Data\hilRBCmaleHill.plt 

 Fri Aug 28 16:22:41 2009 

 ==================================================================== 

 BMDS Model Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the response function is: 

 Y[dose] = intercept + v*dose^n/(k^n + dose^n) 

 Dependent variable = Mean 

 Independent variable = Dose 

 rho is set to 0 

 Power parameter restricted to be greater than 1 

 A constant variance model is fit 

8.2

8.4

8.6

8.8

9

9.2

9.4

0 200 400 600 800

Mean Response

Dose

Hill Model with 0.95 Confidence Level

16:27 07/10 2008

BMDL BMD

 

Hill

1000

 25
FINAL 

9-28-2009 

 Total number of dose groups = 4 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial Parameter Values 

 alpha = 0.09565 

 rho = 0 Specified 

 intercept = 9.27 

 v = -0.74 

 n = 0.812681 

 k = 323.171 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -rho -n 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 alpha intercept v k 

 alpha 1 1.8e-009 2.2e-008 -9.9e-009 

 intercept 1.8e-009 1 0.33 -0.57 

 v 2.2e-008 0.33 1 -0.94 

 k -9.9e-009 -0.57 -0.94 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 alpha 0.0865823 0.0193604 0.0486366 

0.124528 

 intercept 9.24677 0.080516 9.08896 

9.40457 

 v -1.20681 0.578497 -2.34065 -

0.0729792 

 n 1 NA 

 k 689.459 784.134 -847.415 

2226.33 

NA - Indicates that this parameter has hit a bound 

 implied by some inequality constraint and thus 

 has no standard error. 

 Table of Data and Estimated Values of Interest 

 26
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9-28-2009 

 Dose N Obs Mean Est Mean Obs Std Dev Est Std Dev Scaled Res. 

------ --- -------- -------- ----------- ----------- ---------- 

 0 10 9.27 9.25 0.35 0.294 0.25 

 50 10 9.13 9.17 0.22 0.294 -0.378 

 250 10 8.94 8.93 0.34 0.294 0.155 

 1000 10 8.53 8.53 0.31 0.294 -0.0263 

 Model Descriptions for likelihoods calculated 

 Model A1: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma^2 

 Model A2: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma(i)^2 

 Model A3: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma^2 

 Model A3 uses any fixed variance parameters that 

 were specified by the user 

 Model R: Yi = Mu + e(i) 

 Var{e(i)} = Sigma^2 

 Likelihoods of Interest 

 Model Log(likelihood) # Param's AIC 

 A1 29.048402 5 -48.096804 

 A2 30.246635 8 -44.493271 

 A3 29.048402 5 -48.096804 

 fitted 28.933200 4 -49.866400 

 R 16.206159 2 -28.412317 

 Explanation of Tests 

 Test 1: Do responses and/or variances differ among Dose levels? 

 (A2 vs. R) 

 Test 2: Are Variances Homogeneous? (A1 vs A2) 

 Test 3: Are variances adequately modeled? (A2 vs. A3) 

 Test 4: Does the Model for the Mean Fit? (A3 vs. fitted) 

 (Note: When rho=0 the results of Test 3 and Test 2 will be the same.) 

 Tests of Interest 

 Test -2*log(Likelihood Ratio) Test df p-value 

 Test 1 28.081 6 <.0001 

 Test 2 2.39647 3 0.4943 

 Test 3 2.39647 3 0.4943 

 Test 4 0.230404 1 0.6312 

The p-value for Test 1 is less than .05. There appears to be a 

difference between response and/or variances among the dose levels 

It seems appropriate to model the data 

The p-value for Test 2 is greater than .1. A homogeneous variance 

 27
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9-28-2009 

model appears to be appropriate here 

The p-value for Test 3 is greater than .1. The modeled variance appears 

 to be appropriate here 

The p-value for Test 4 is greater than .1. The model chosen seems 

to adequately describe the data 

 

 Benchmark Dose Computation 

Specified effect = 1 

Risk Type = Estimated standard deviations from the control mean 

Confidence level = 0.95 

 BMD = 222.31 

 BMDL = 81.4084 

 28
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9-28-2009 

Model Fitting Results for Decreased RBCs in Female Rats (Johnson et al., 2005) 

Following the above procedure, continuous-variable models in the EPA BMDS (version 

2.1) were fit to the data shown in Table 2 (main text) for decreased RBCs in female rats. In the 

absence of a biologically relevant response level, the BMR was chosen to be 1 SD from the 

control mean, as recommended by U.S. EPA (2000). As shown in Table A-2, the linear and 

power models (linear model shown in Figure A-2) provide identical fit to the data and yielded 

AIC values that were significantly lower than the other models fitted. Thus, the estimated 

BMDL for this endpoint is 280 mg/kg-day. 

Table A-2. Model Predictions for Decreased RBCs in Female Ratsa

Data Set/Model 

Variance 

p-Valueb

Means 

p-Valueb AIC 

BMD1SD 

(mg/kg-day)

BMDL1SD

(mg/kg-day)

Linear (constant variance)c 0.3052 0.2239 -69.286 364.222 279.607 

2-Degree Polynomial (constant variance)c 0.3052 0.0862 -67.3357 408.969 280.868

3-Degree Polynomial (constant variance)c 0.3052 0.08802 -67.3694 429.854 281.283

Power (constant variance)d 0.3052 0.2239 -69.286 364.876 280.266 

Hill (constant variance)d 0.3052 0.08352 -67.2847 364.222 279.604 a

Johnson et al. (2005)

b

Values <0.10 fail to meet conventional goodness-of-fit criteria 

c

Coefficients restricted to be negative

d

Power restricted to ≥1 

 29
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9-28-2009 

Figure A-2. Fit of Linear Model to Data on Decreased RBCs in Female Rats 

(Johnson et al., 2005) 

BMDs and BMDLs indicated are associated with a change of 1 SD from the control and are in units mg/kg-day 

 ==================================================================== 

 Hill Model. (Version: 2.14; Date: 06/26/2008) 

 Input Data File: C:\USEPA\BMDS21\Data\hilRBCmaleHill.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS21\Data\hilRBCmaleHill.plt 

 Fri Aug 28 16:57:54 2009 

 ==================================================================== 

 BMDS Model Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the response function is: 

 Y[dose] = intercept + v*dose^n/(k^n + dose^n) 

 Dependent variable = Mean 

 Independent variable = Dose 

 rho is set to 0 

 Power parameter restricted to be greater than 1 

 A constant variance model is fit 

 Total number of dose groups = 4 

7.4

7.6

7.8

8

8.2

8.4

0 200 400 600 800

Mean Response

Dose

Linear Model with 0.95 Confidence Level

16:32 07/10 2008

BMDL BMD

 

Linear

1000

 30
FINAL 

9-28-2009 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial Parameter Values 

 alpha = 0.05775 

 rho = 0 Specified 

 intercept = 8.26 

 v = -0.72 

 n = 0.970124 

 k = 490.566 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -rho -n 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 alpha intercept v k 

 alpha 1 0.00028 0.0044 -0.0044 

 intercept 0.00028 1 0.061 -0.065 

 v 0.0044 0.061 1 -1 

 k -0.0044 -0.065 -1 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 alpha 0.056016 0.0125257 0.0314661 

0.0805658 

 intercept 8.19343 0.048292 8.09878 

8.28808 

 v -240.216 6001.98 -12003.9 

11523.5 

 n 1 NA 

 k 369305 9.25312e+006 -1.77665e+007 

1.85051e+007 

NA - Indicates that this parameter has hit a bound 

 implied by some inequality constraint and thus 

 has no standard error. 

 Table of Data and Estimated Values of Interest 

 Dose N Obs Mean Est Mean Obs Std Dev Est Std Dev Scaled Res. 

------ --- -------- -------- ----------- ----------- ---------- 

 31
FINAL 

9-28-2009 

 0 10 8.26 8.19 0.22 0.237 0.889 

 50 10 8.06 8.16 0.31 0.237 -1.35 

 250 10 8.07 8.03 0.24 0.237 0.522 

 1000 10 7.54 7.54 0.17 0.237 -0.0632 

 Model Descriptions for likelihoods calculated 

 Model A1: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma^2 

 Model A2: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma(i)^2 

 Model A3: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma^2 

 Model A3 uses any fixed variance parameters that 

 were specified by the user 

 Model R: Yi = Mu + e(i) 

 Var{e(i)} = Sigma^2 

 Likelihoods of Interest 

 Model Log(likelihood) # Param's AIC 

 A1 39.139849 5 -68.279698 

 A2 40.951049 8 -65.902099 

 A3 39.139849 5 -68.279698 

 fitted 37.642370 4 -67.284740 

 R 21.815106 2 -39.630212 

 Explanation of Tests 

 Test 1: Do responses and/or variances differ among Dose levels? 

 (A2 vs. R) 

 Test 2: Are Variances Homogeneous? (A1 vs A2) 

 Test 3: Are variances adequately modeled? (A2 vs. A3) 

 Test 4: Does the Model for the Mean Fit? (A3 vs. fitted) 

 (Note: When rho=0 the results of Test 3 and Test 2 will be the same.) 

 Tests of Interest 

 Test -2*log(Likelihood Ratio) Test df p-value 

 Test 1 38.2719 6 <.0001 

 Test 2 3.6224 3 0.3052 

 Test 3 3.6224 3 0.3052 

 Test 4 2.99496 1 0.08352 

The p-value for Test 1 is less than .05. There appears to be a 

difference between response and/or variances among the dose levels 

It seems appropriate to model the data 

The p-value for Test 2 is greater than .1. A homogeneous variance 

model appears to be appropriate here 

 32
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9-28-2009 

The p-value for Test 3 is greater than .1. The modeled variance appears 

 to be appropriate here 

The p-value for Test 4 is less than .1. You may want to try a different 

model 

 

 Benchmark Dose Computation 

Specified effect = 1 

Risk Type = Estimated standard deviations from the control mean 

Confidence level = 0.95 

 BMD = 364.222 

 BMDL = 279.607 

 33
FINAL 

9-28-2009 

Model Fitting Results for Decreased Hemoglobin Concentration in Male Rats 

(Johnson et al., 2005) 

Following the above procedure, continuous-variable models in the EPA BMDS 

(version 1.4.1c) were fit to the data shown in Table 3 (main text) for decreased hemoglobin 

concentration in male rats. In the absence of a biologically relevant response level, the BMR 

was chosen to be 1 SD from the control mean, as recommended by U.S. EPA (2000). As shown 

in Table A-3, the linear, polynomial, and power models provide identical linear fit to the data. 

Fit to the Hill Model cannot be tested due to insufficient degrees of freedom. Fit to the linear 

model is illustrated in Figure A-3. For this endpoint, the estimated BMDL is 328 mg/kg-day. 

Table A-3. Model Predictions for Decreased Hemoglobin in Male Ratsa

Data Set/Model 

Variance 

p-Valueb

Means 

p-Valueb AIC 

BMD1SD 

(mg/kg-day)

BMDL1SD

(mg/kg-day)

Linear (constant variance)c 0.8494 0.2992 -23.84147 439.943 328.015 

2-Degree Polynomial (constant variance)c 0.8494 0.2992 -23.84147 439.943 328.015 

3-Degree Polynomial (constant variance)c 0.8494 0.2992 -23.84147 439.943 328.015 

Power (constant variance)d 0.8494 0.2992 -23.84147 439.943 328.015 

Hill (constant variance)d Fit cannot be tested for this model due to inadequate degrees of 

freedom a

Johnson et al. (2005)

b

Values <0.10 fail to meet conventional goodness-of-fit criteria 

c

Coefficients restricted to be negative

d

Power restricted to ≥1 

 34
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9-28-2009 

Linear Model with 0.95 Confidence Level

Figure A-3. Fit of Linear Model to Data on Decreased 

Hemoglobin Concentration in Male Rats (Johnson et al., 2005) 

BMDs and BMDLs indicated are associated with a change of 1 SD from the control and are in units mg/kg-day 

 ==================================================================== 

 Polynomial Model. (Version: 2.13; Date: 04/08/2008) 

 Input Data File: C:\USEPA\BMDS21\Data\linRBCmaleLinear.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS21\Data\linRBCmaleLinear.plt 

 Fri Aug 28 17:11:28 2009 

 ==================================================================== 

 BMDS Model Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the response function is: 

 Y[dose] = beta_0 + beta_1*dose + beta_2*dose^2 + ... 

 Dependent variable = Mean 

 Independent variable = Dose 

 rho is set to 0 

14.4

14.6

14.8

15

15.2

15.4

15.6

15.8

16

Linear 16.2

0 200 400 600 800 1000

Mean Response

Dose

16:41 07/10 2008

BMDL BMD

 

 35
FINAL 

9-28-2009 

 The polynomial coefficients are restricted to be negative 

 A constant variance model is fit 

 Total number of dose groups = 4 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial Parameter Values 

 alpha = 0.1825 

 rho = 0 Specified 

 beta_0 = 15.7086 

 beta_1 = -0.000949416 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -rho 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 alpha beta_0 beta_1 

 alpha 1 -1.3e-009 1.7e-009 

 beta_0 -1.3e-009 1 -0.63 

 beta_1 1.7e-009 -0.63 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 alpha 0.174464 0.0390113 0.0980032 

0.250925 

 beta_0 15.7086 0.0850278 15.5419 

15.8752 

 beta_1 -0.000949416 0.000164784 -0.00127239 -

0.000626445 

 Table of Data and Estimated Values of Interest 

 Dose N Obs Mean Est Mean Obs Std Dev Est Std Dev Scaled Res. 

------ --- -------- -------- ----------- ----------- ---------- 

 0 10 15.8 15.7 0.5 0.418 0.692 

 50 10 15.7 15.7 0.4 0.418 0.295 

 250 10 15.3 15.5 0.4 0.418 -1.3 

 1000 10 14.8 14.8 0.4 0.418 0.309 

 36
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9-28-2009 

 Model Descriptions for likelihoods calculated 

 Model A1: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma^2 

 Model A2: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma(i)^2 

 Model A3: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma^2 

 Model A3 uses any fixed variance parameters that 

 were specified by the user 

 Model R: Yi = Mu + e(i) 

 Var{e(i)} = Sigma^2 

 Likelihoods of Interest 

 Model Log(likelihood) # Param's AIC 

 A1 16.127312 5 -22.254625 

 A2 16.527404 8 -17.054808 

 A3 16.127312 5 -22.254625 

 fitted 14.920736 3 -23.841473 

 R 2.835616 2 -1.671231 

 Explanation of Tests 

 Test 1: Do responses and/or variances differ among Dose levels? 

 (A2 vs. R) 

 Test 2: Are Variances Homogeneous? (A1 vs A2) 

 Test 3: Are variances adequately modeled? (A2 vs. A3) 

 Test 4: Does the Model for the Mean Fit? (A3 vs. fitted) 

 (Note: When rho=0 the results of Test 3 and Test 2 will be the same.) 

 Tests of Interest 

 Test -2*log(Likelihood Ratio) Test df p-value 

 Test 1 27.3836 6 0.0001227 

 Test 2 0.800183 3 0.8494 

 Test 3 0.800183 3 0.8494 

 Test 4 2.41315 2 0.2992 

The p-value for Test 1 is less than .05. There appears to be a 

difference between response and/or variances among the dose levels 

It seems appropriate to model the data 

The p-value for Test 2 is greater than .1. A homogeneous variance 

model appears to be appropriate here 

The p-value for Test 3 is greater than .1. The modeled variance appears 

 to be appropriate here 

The p-value for Test 4 is greater than .1. The model chosen seems 

to adequately describe the data 

 

 37
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9-28-2009 

 Benchmark Dose Computation 

Specified effect = 1 

Risk Type = Estimated standard deviations from the control mean 

Confidence level = 0.95 

 BMD = 439.943 

 BMDL = 328.015 

 38
FINAL 

9-28-2009 

Model Fitting Results for Decreased Hemoglobin Concentration in Female Rats 

(Johnson et al., 2005) 

Following the above procedure, continuous-variable models in the EPA BMDS 

(version 1.4.1c) were fit to the data shown in Table 3 (main text) for decreased hemoglobin 

concentration in female rats. In the absence of a biologically relevant response level, the BMR 

was chosen to be 1 SD from the control mean, as recommended by U.S. EPA (2000). As shown 

in Table A-4, none of the models provides an adequate fit to the data, even with the highest dose 

dropped. Dropping the highest dose reduces the data set to controls and two dose groups; 

therefore, it is not possible to drop further dose groups to attempt to identify a fit. Therefore, this 

data set is not amenable to BMD modeling. 

Table A-4. Model Predictions for Decreased Hemoglobin in Female Ratsa

Data Set/Model 

Variance 

p-Valueb

Means 

p-Valueb AIC 

BMD1SD 

(mg/kg-day)

BMDL1SD

(mg/kg-day)

All Doses

Linear (constant variance)c 0.652 0.05345 -35.534 579.683 408.287 

2-Degree Polynomial (constant variance)c 0.652 0.05345 -35.534 579.683 408.287 

3-Degree Polynomial (constant variance)c 0.652 0.05345 -35.534 579.683 408.287 

Power (constant variance)d 0.652 0.05345 -35.534 579.683 408.287 

Hill (constant variance)d 0.652 0.03974 -35.163 127.36 11.921

Highest Dose Dropped

Linear (constant variance)c 0.5991 0.03223 -22.281 331.251 170.263

2-Degree Polynomial (constant variance)c 0.5991 0.03223 -22.281 331.251 170.263

Power (constant variance)d 0.5991 0.03223 -22.281 331.251 170.263

Hill (constant variance)d # observations < # parameters: cannot run this model a

Johnson et al. (2005)

b

Values <0.10 fail to meet conventional goodness-of-fit criteria 

c

Coefficients restricted to be negative

d

Power restricted to ≥1 

 39
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9-28-2009 

Modeling Procedure for Hepatocyte Vacuolization in Females (Gushow et al., 1984) 

The BMD modeling for dichotomous data was conducted with the EPA’s BMD software 

(BMDS version 2.1). For all the dichotomous data, the incidence data on hepatocyte 

vacuolization with duration-adjusted concentrations (see Table 4 of main text) were modeled 

with eight dichotomous models (i.e., Gamma, Multistage, Logistic, Log-logistic, Probit, Logprobit, Weibull, and Quantal-linear models) with a BMR of 10% extra risk. An adequate model 

fit was judged based on the goodness-of-fit p-value (p > 0.1), scaled residual at the range of 

BMR, and visual inspection of the model fit. As shown in Table A-5 and Figures A-4 and A-5, 

all the models provided adequate goodness-of-fit p-values (≥0.1), and estimated BMCLs from 

these models are sufficiently close (ranging from 0.32 to 0.60 mg/m3

). Therefore, the BMCL of 

0.32 mg/m3

 estimated from the Multistage model (Figure A-5), with the lowest AIC from these 

five models, was considered to be an appropriate estimate for this endpoint. 

Table A-5. Model Predictions for Increased Hepatocyte Vacuolization in Female Ratsa

Data Set/Model Goodness-of-Fit p-Valueb AIC Scaled Residual 

BMC10 

(mg/m3) 

BMCL10

(mg/m3

) 

Gammac 0.9917 38.18 0.001 2.0 0.32 

Logistic 0.8180 36.58 0.367 0.9 0.52 

Log-logisticd 0.8523 38.24 0.053 2.1 0.50 

Multistagee 0.9889 36.20 -0.119 1.7 0.32 

Probit 0.8330 36.54 0.338 0.8 0.55 

Log-probitd 0.9398 38.19 0.017 2.1 0.60 

Weibullc 0.9999 38.18 0.000 1.9 0.32 

Quantal-linear 0.5132 37.60 0.360 0.48 0.27 a

Gushow et al. (1984)

b

Values <0.10 fail to meet conventional goodness-of-fit criteria 

c

Power restricted to ≥1 

d

Slope restricted to ≥1 e

Betas restricted to ≥0 

 40
FINAL 

9-28-2009 

 0

 0.2

 0.4

 0.6

 0.8

 1

 0 5 10 15 20

Fraction Affected

dose

Figure A-4. Fit of Quantal-Linear Model to Data 

Quantal Linear Model with 0.95 Confidence Level

13:56 07/28 2009

BMDL BMD

 

Quantal Linear

Dose

on Increased Hepatocyte Vacuolization in Female Rats (Gushow et al., 1984) 

BMCs and BMCLs indicated are associated with a change of 10% extra risk from the control, and are in units mg/m3

 41
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9-28-2009 

 0

 0.2

 0.4

 0.6

 0.8

 1

 0 5 10 15 20

Fraction Affected

dose

Figure A-5. Fit of Multistage Model to Data 

Multistage Model with 0.95 Confidence Level

14:53 07/28 2009

BMDL BMD

 

Multistage

Dose

on Increased Hepatocyte Vacuolization in Female Rats (Gushow et al., 1984) 

BMCs and BMCLs indicated are associated with a change of 10% extra risk from the control, and are in units mg/m3

 ==================================================================== 

 Multistage Model. (Version: 3.0; Date: 05/16/2008) 

 Input Data File: C:\USEPA\BMDS21\Data\mstHepaticmulti.(d) 

 Gnuplot Plotting File: C:\USEPA\BMDS21\Data\mstHepaticmulti.plt 

 Fri Aug 28 17:38:06 2009 

 ==================================================================== 

 BMDS Model Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP( 

 -beta1*dose^1-beta2*dose^2)] 

 The parameter betas are restricted to be positive 

 Dependent variable = Response 

 42
FINAL 

9-28-2009 

 Independent variable = Dose 

 Total number of observations = 4 

 Total number of records with missing values = 0 

 Total number of parameters in model = 3 

 Total number of specified parameters = 0 

 Degree of polynomial = 2 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 Default Initial Parameter Values 

 Background = 0 

 Beta(1) = 0 

 Beta(2) = 2.27362e+017 

 Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Beta(1) 

 have been estimated at a boundary point, or have been specified by 

the user, 

 and do not appear in the correlation matrix ) 

 Background Beta(2) 

Background 1 -0.41 

 Beta(2) -0.41 1 

 Parameter Estimates 

 95.0% Wald Confidence 

Interval 

 Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. 

Limit 

 Background 0.288586 * * * 

 Beta(1) 0 * * * 

 Beta(2) 0.0381152 * * * 

* - Indicates that this value is not calculated. 

 Analysis of Deviance Table 

 Model Log(likelihood) # Param's Deviance Test d.f. P-value 

 Full model -16.0896 4 

 Fitted model -16.1007 2 0.0222818 2 0.9889 

 Reduced model -25.8979 1 19.6166 3 0.0002038 

 AIC: 36.2015 

 Goodness of Fit 

 43
FINAL 

9-28-2009 

 Scaled 

 Dose Est._Prob. Expected Observed Size Residual 

 ------------------------------------------------------------------------ 

 0.0000 0.2886 2.886 3.000 10 0.080 

 2.3000 0.4185 4.185 4.000 10 -0.119 

 7.1000 0.8958 8.958 9.000 10 0.043 

 21.3000 1.0000 10.000 10.000 10 0.000 

 Chi^2 = 0.02 d.f. = 2 P-value = 0.9889 

 Benchmark Dose Computation 

Specified effect = 0.1 

Risk Type = Extra risk 

Confidence level = 0.95 

 BMD = 1.66261 

 BMDL = 0.318188 

 44