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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 Envi ronmental Assessment Office of Research and Development U.S. Environmental Protection Agency Cincinnati, OH 45268Commonly Used Abbreviations

BMD Benchmark Dose IRIS Integrated Risk Information System

IUR inhalation unit risk

LOAEL lowest-observed-adverse-effect level

LOAEL ADJ LOAEL adjusted to c ontinuous exposure duration

LOAEL HEC LOAEL adjusted for dosimetric di fferences across species to a human

NOAEL no-observed-adverse-effect level

NOAEL ADJ NOAEL adjusted to cont inuous exposure duration

NOAEL HEC 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 inhala tion reference concentration p-RfD provisional oral reference dose

RfC inhalation refe rence concentration

RfD oral reference dose UF uncertainty factor

UF A animal to human uncertainty factor

UF C composite uncertainty factor

UF D incomplete to complete database uncertainty factor

UF H interhuman uncertainty factor

UF L LOAEL to NOAEL uncertainty factor

UF S 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 assessmen ts, establishing the follow ing three tiers as the new hierarchy:

1) U.S. EPA's Integrated Risk Information System (IRIS) 2) Provisional Peer-Reviewed Toxicity Valu es (PPRTVs) used in U.S. EPA's Superfund Program 3) Other (peer-reviewed) to xicity values, including:

< Minimal Risk Levels produced by the Agency for Toxic Substances and Disease Registry (ATSDR)

< California Environmental Protec tion Agency (CalEPA) values

< EPA Health Effects Assessment Summary Table (HEAST) values

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

such a value is not available in U.S. EPA's IR IS. 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 provisiona l toxicity values receive internal review by two U.S. EPA scientists and external peer revi ew 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 speci fically for the Superfund Program.

Because new information becomes available and scientific methods improve over time, PPRTVs are reviewed on a 5-year basis and updat ed into the active database. Once an IRIS value for a specific chemical becomes availabl e 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 se e 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 Recove ry 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 e xposures and circumstances at the Superfund site or RCRA facility in question. PPRTVs are periodic ally updated; therefore, users should ensure that the values

contained in the PPRTV are cu rrent at the time of use.

It is important to remember that a provi sional 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 en tire PPRTV document and understand the strengths and limitations of the derived provisional va lues. PPRTVs are developed by the U.S. EPA

Office of Research and Deve lopment’s National Center for Environmental Assessment, Superfund Health Risk Technica l Support Center for OSRTI. Other U.S. EPA programs or external parties who may choose of their own initiative to us e 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 th e PPRTVs and their ap propriate 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 Deve lopment’s National Center for Environmental Assessment, Superfund Health Risk Technica l Support Center (513-569-7300), or OSRTI.

INTRODUCTION

Diethylene glycol monobutyl ethe r (DGBE) is a colorless liq uid. Its synonyms include 2-(2-butoxyethoxy)-ethanol, butoxydiglycol, butadi gol, and butyl carbitol. The formula for

DGBE is C 8 H 18 O 3 (Figure 1) with a molecular wei ght 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 IR IS (U.S. EPA, 2008) or in the HEAST (U.S. EPA, 1997) or Drinking Water Standards a nd Health Advisories list (U.S. EPA, 2006).

Subchronic and chronic RfDs of 0.04 and 0.004 mg/kg-day, resp ectively, 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 ra ts (Hobson et al., 1987) and an UF of 1,000 (subchronic RfD) or 10,000 (chr onic 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 doc ument 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 IR IS (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 H ealth (NIOSH, 2005), or the Occupational Safety and Health Administration (OSHA, 2008).

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3 DGBE is not listed in the HEAST cancer tabl e (U.S. EPA, 1997), a nd 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 Res earch 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 He alth Criteria documents (WHO, 2008) for DGBE. CalEPA (2002, 2005a,b) has not derived oral or inhalation recommende d 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 provi sional toxicity values for DGBE. Databases searched included: MEDLINE (including cancer subset), TOXLIN E (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-w eek 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 ev aluated for clinical signs, body weight, food consumption, water consumption, urinalysis va riables, 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 stat istically significant), with effects greatest in high-dose males (1 4% lower body weight than controls by study termination). Red blood cell count (RBC), hemo globin (Hgb), and hematocrit (Hct) were decreased (not statistically significant) relative to controls at all doses, w ith 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 st atistically significant treatment-related adverse effects on serum chemistry, urinalysis, or organ weights. The only eff ect observed at gross necropsy was urine staining of the perineal fur (all doses).FINAL9-28-2009Table 1. Summary Table for Diethylene Glycol Monobutyl EtherSpecies andStudy TypeExposureCritical EffectsNOAELLOAELCommentsReferenceOral StudiesRatShort-term0, 51, 94, 210, 650,970, or1,830 mg/kg-day indrinking water for30 days.Reduced water intake at≥94 mg/kg-day andhistopathological changesin the liver (congestionand slight cloudyswelling) and kidneytubules (cloudy swellingand increased secretion)at≥650 mg/kg-day.Cannot bedeterminedCannot bedeterminedEndpoints included mortality, waterintake, body weight and histology ofthe liver, kidneys, spleen, and testis.Small groups of rats (five/sex/dose).This is a range-finding study that islimited in scope and inadequatelyreported.Smyth, 1940.Smyth and Carpenter,1948.RatSubchronic891, 1,782, or3,564 mg/kg-day bygavage on5 days/week, for6 weeks.(average daily dose:636, 1,273, or2,546 mg/kg-day)Hyperkeratosis of thestomach at≥891 mg/kg-day andhematologic effects(reduced erythrocytecount, hemoglobinconcentration, andMCHC and increasedMCV) at≥1,782 mg/kg-day.891mg/kg-day(averagedaily dose of636mg/kg-day)1,782mg/kg-day(average dailydose of 1,273mg/kg-day)HematologicaleffectsEndpoints included clinical signs, foodconsumption, body weight,hematology, clinical chemistry, organweights, gross pathology, andhistopathology. Human relevance ofgastric hyperkeratosis is questionabledue to bolus exposure and likely directirritant properties of DGBE.Kodak, 1984.RatSubchronic0, 70, 330, or1,630 mg/kg-day(males) or 0, 50, 250,or 1,270 mg/kg-day(females) by gavageon 5 days/week, for13 weeks.(Average daily dose:0, 50, 236, or1,164 mg/kg-day[males]; 0, 36, 179,or 907 mg/kg-day[females].)Decreased total WBC andlymphocyte counts andMCHC in females at≥50 mg/kg-day.Mortality in both sexes at≥250/330 mg/kg-day(some of the deaths weredue to gavage error).N/A50 mg/kg-day(average dailydose of36 mg/kg-day)LymphopeniaEndpoints included clinical signs, foodconsumption, body weight,hematology, clinical chemistry,urinalysis, organ weights, grosspathology, and histopathology. Thehistological examinations did notinclude the stomach.Hobson et al, 1987.4FINAL9-28-2009Table 1. Summary Table for Diethylene Glycol Monobutyl EtherSpecies andStudy Type Exposure Critical Effects NOAEL LOAEL Comments ReferenceRatSubchronic0, 50, 250, or1000 mg/kg-day indrinking water for13 weeks.Decreased RBC count,hemoglobin, andhematocrit at≥250 mg/kg-day. Othereffects only occurred at1,000 mg/kg-day andmainly involved the liver;these included increasesin organ weight andhepatic cytochromeP450s and UGT levels,decreases in serum totalprotein, cholesterol, andserum AST, andhepatocyte hypertrophyand individual hepatocytedegeneration.50 mg/kg-day250mg/kg-dayReduced RBCcount and Hgbin both sexes.Endpoints included clinical signs, foodand water consumption, body weight,hematology, clinical chemistry,urinalysis, functional observationalbattery, sperm analysis, liver metabolicenzymes, organ weights, grosspathology, and histopathology. Studypossibly useful for subchronic andchronic p-RfD derivation.Johnson et al, 2005.RatReproductive0, 250, 500, or1,000 mg/kg-day bygavage in water.Untreated maleswere bred to treatedfemales and viceversa. The maleswere treated for60 days prior tomating, and thefemales were treatedfrom 14 days prior tomating to GD 13 oruntil weaning of theoffspring.Slight reduction in pupweight during the lastweek of lactation in theoffspring of the femalesdosed with 1,000 mg/kgday. No maternal toxicityor reproductive effectswere observed.500mg/kg-day1,000mg/kg-dayDecreased pupweight duringthe last weekof lactation.Endpoints included body weight inboth sexes; numbers of corpora lutea,implants, resorptions, viable embryosand live or dead pups; and clinicalcondition, body weight, and externaldevelopment in pups.Nolen et al, 1985.5FINAL9-28-20096Table 1. Summary Table for Diethylene Glycol Monobutyl EtherSpecies andStudy Type Exposure Critical Effects NOAEL LOAEL Comments ReferenceRatReproductiveandDevelopmental0, 25, 115, or633 mg/kg-day indiet on GD 0–20.Rats were killed onGD 20 for uterineand fetalexaminations orallowed to deliverpups that were reareduntil 10 weeks ofage.Reduced maternal bodyweight gain duringpregnancy at≥25 mg/kgday (the effect wasvariable and not doserelated). No prenatal orpostnatal developmentaltoxicity.633mg/kg-dayMaternal endpoints included clinicalsigns, food consumption, and bodyweight. Developmental endpointsincluded numbers of corpora lutea,implantations, litters, and live fetusesper litter, pre and postimplantationlosses, fetal and placental weights, sexratio, external and oral cavityanomalies, skeletal and internalanomalies, gestation length, numbersof live newborns, and pup body weightand survival.Ema et al., 1988.MouseDevelopmental0 or 500 mg/kg-daygavage in water onGD 7–14.No maternal ordevelopmental toxicity.500mg/kg-dayDevelopmental toxicity screeningassay. Study endpoints includedmaternal body weight, fetal survival,pup perinatal and postnatal survival,and pup body weight. Teratogenicitynot evaluated.Schuler et al., 1984.Inhalation StudyRatSubchronic0, 2, 6, or 18 ppm (0,13, 40, or119 mg/m3) for6 hours/day,5 days/week, for upto 22 exposures in5 weeks.(average dailyconcentration: 0, 2.3,7.1, or 21.3 mg/m3)Liver histopathology(slight hepatocytevacuolization consistentwith fatty change) infemales at≥6 ppm andgross paleness of the liverin 3/10 females at18 ppm.2 ppm(averageconcentrationof 13 mg/m3)6 ppm (averageconcentrationof 40 mg/m3)Endpoints included clinical signs, bodyweight, hematology, clinical chemistry,urinalysis, organ weights, grosspathology, and histopathology.Histological exams werecomprehensive at 0 and 18 ppm butlimited to the liver at 2 and 6 ppm.Study used to derive subchronic andchronic RfCs in previous PPRTVdocument.Gushow et al., 1984.FINAL

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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 drinki ng water at reported doses of 0, 51, 94, 210, 650, 970, or 1,830 mg/kg-day for 30 days (Smyth, 1940; Sm yth and Carpenter, 1948). The study was summarized only briefly. Endpoints included mortality, water intake, body weight, and histology of the liver, kidney, sp leen, 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 a ffected at any dose level. Hist opathological changes occurred at

≥ 650 mg/kg-day that included cloudy swelling and in creased secretion in the kidney tubules and congestion and slight cloudy swelling of the liver. The limited scope and inadequate reporting of this study precludes identifica tion 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 wei ght, measured hematology values (Hgb, Hct, RBC, and total and differential white blood cel l counts [WBC]) and calculated red blood cell indices (mean corpuscular volume [MCV], mean corpuscular hemoglobin [MCH], and mean corpuscular hemoglobin concentrat ion [MCHC]), several serum chemistry indicators of liver damage and kidney function (alanine ami notransferase [ALT or SGPT], aspartate aminotransferase [AST or SGOT], alkaline phosphatase [ALP], lactate dehydrogenase [LDH], blood urea nitrogen [BUN], creatinin e, 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), mi d- (2/10), and low- (1 /10) dose groups. An additional two rats in the high- dose group were killed in mori bund 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 in cluded 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 MC V [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 hi stopathological changes in the kidneys (hyaline droplet degeneration, proteina ceous casts, and hemosiderin in the proximal tubules). The investigators noted that the renal proteinaceous cas ts 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 tr eatment-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 th e controls. The human re levance of the gastric hyperkeratosis is questionable due to the bolus (undiluted gavage) type of exposures and likely direct irritant propertie s of DGBE, as indicated by aspiratio n-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 gr oups 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 we re 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 ra ts/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, hema tology (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 ur ine), 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 fe males starting in Week 4, and high-dose rats of both sexes starting in Week 1. Survivi ng 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 stat istical 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 trac t 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, pulm onary 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 obser vations of foreign body pneumonia and acute pleuritis, indicated that gavage-related aspira tion of DGBE may have occurred, contributi ng to mortality in some of the rats. Study author s reported that over the entire study, respiratory tract le sions 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 a ttributable 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 hi gh-dose rats (i.e., unclear pathogenesis of respiratory lesions) and experi mental observations by the dosi ng technicians also did not completely support a definitive relationship betw een 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 b een 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 DG BE exposure (Hobson et al., 1987) included dose-related decreases in total WBC and lympho cyte 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 d ecrease 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 eith er sex at 13 weeks. However, the researchers reported that at 6 weeks (i.e., before all th e 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; ne vertheless, neutrophil counts were significantly increase d 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 B UN 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 drop lets 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 hemato logical 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 Leukopen ia/Lymphopenia in Female Rats Exposed to DGBE by Gavage for 13 Weeks a

Duration-Adjusted Dose b (mg/kg-day) Endpoint (Mean ± SD) 0 35.7 178.6

Number Examined 7 9 7

WBC count (× 10 3 /mm 3 ) 4.46 ± 0.73 3.03 ± 0.59 c (67.9%) d 2.56 ± 0.31 c (57.4%)

Lymphocyte count (× 10 3 /mm 3 ) 3.55 ± 0.6 2.31 ± 0.42 c (65.1%) 1.85 ± 0.31 c (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 pres ented because this dose group was dropped from further analysis due to high mortality.

Johnson et al. (2005) administ ered 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 w eekly for detailed clinical evaluation, body weight, and food and water consum ption. Ophthalmological examinations and a behavioral functional observational battery (F OB) were administered before initiation of exposure and during the last w eek 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, wa s conducted for all rats. Comprehensive histopathological evalua tions 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 ot her than some urine so iling in the high-dose group, there were no treatment-rela ted 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 consump tion among high-dose rats was 7–8% less than controls throughout the study, and weekly f ood consumption among high-dose rats was

5–11% less than control values. No treatment-re lated effects on body weight or food or water consumption were noted in the mid- and low- dose groups. No treat ment-related effects on urinalysis, ophthalmology or the FOB (sensory eval uation, 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 hemato logical variables relative to controls in mid- (RBC, Hgb) a nd 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). Th ere were no other treatment-re lated effects on hematological variables, including total and di fferential WBC (data not shown).

Table 3. Data Sets for Anemia in Male and Female Rats Exposed to DGBE in Drinking Water for 13 Weeks a

Dose (mg/kg-day) Endpoint 0 50 250 1000

Males

Number Examined 10 10 10 10

RBC Count (× 10 6 /μL) 9.27 ± 0.35 9.13 ± 0.22 (98.5%) c 8.94 ± 0.34 b
(96.4%) 8.53 ± 0.31 b
(92.0%)

Hgb (g/dL) 15.8 ± 0.5 15.7 ± 0.4 (99.3%) 15.3 ± 0.4 b
(96.8%) 14.8 ± 0.4 b
(93.6%)

Females

Number Examined 10 10 10 10

RBC Count (× 10 6 /μL) 8.26 ± 0.22 8.06 ± 0.31 (97.6%) 8.07 ± 0.24 b
(97.7%) 7.54 ± 0.17 b
(91.3%)

Hgb (g/dL) 15.6 ± 0.3 15.2 ± 0.4 (97.4%) 15.2 ± 0.4 b
(97.4%) 14.8 ± 0.3 b
(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 (10 6 / μ L), RBC in females 7.81–8.20 (10 6 / μ L), 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 generall y unremarkable, although there were slight decreases in serum protein, chol esterol, and AST in high-dose males and females (Johnson et al., 2005). Liver metabolic enzymes were slightly in creased relative to controls only for high-dose

animals as follows: EROD 1 (24% both sexes); PROD 2 (38% males, 24%, females); and UGT 3
(17% males, 16% females). Orga n weight results included increased absolute and relative liver and kidney weights in the high- dose groups of both sexes, desp ite 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 treatm ent-related and consis ted of hepatocellular hypertrophy (“very slight” in 6/10 high-dose fema les) and foci of necrotic cells in the centrolobular region. No treatment-related hist opathologic 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 th is 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 re garding 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 (Nol en 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 mati ng to gestation day (GD) 13 or until weaning of the offspring. Half of the females in each grou p 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 treatm ent 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 toxic ity 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 re ported) 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 ra ts 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 include d daily evaluations for clinical signs of t oxicity 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 a nd postimplantation losses,

1 Ethoxyresorufin-O-dealkylase.

2 Pentoxyresorufin-O-dealkylase.

3 UDP-glucuronosyltransferase.

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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 mate rnal 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 e ndpoint is not considered to be an adverse response to DGBE. The lack of a ny prenatal or postnatal developmental effects indicates that 633 mg/kg-day is a NOAEL for de velopmental 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 postn atal survival (numbers of live and dead pups per litter and pup survival to age 2.5 days), and pup body we ights (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, litt er, or pup effects were observe d, indicating that 500 mg/kg-day was a NOAEL for maternal and developmental toxicity.

Inhalation Exposure

Chronic Studies —No pertinent data were located re garding 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 pp m of DGBE (~98.6% pure) (0, 13, 40, or 119 mg/m 3 ) for 6 hours/day, 5 days/week, for up to 22 exposur es in 5 weeks (Gushow et al., 1984). The 119 mg/m 3 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/s ex/group included clin ical 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, ketone s, bilirubin, urobilinogen, occult blood, protein, and specific gravity) after 16–22 e xposures; 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/m 3 groups but limited to the liver in the 13 and 40 mg/m 3 groups because this was the only tissue identified as a possible target at 119 mg/m 3 . Statistically significant ( p < 0.05) decreases in serum glucose levels were observed

in females at ≥ 13 mg/m 3 and males at ≥ 40 mg/m 3 , but the reductions were minimal (9–16% lower than controls). Other effects incl uded slightly increased WBC in males (within the normal range of variation and not consider ed to be toxicological ly significant by the researchers) and hepatic ch anges in both sexes at ≥ 40 mg/m 3 . The hepatic effects included inconsistent changes in relative li ver weight (decreased in males at ≥ 40 mg/m 3 and increased in females at 119 mg/m 3 ) with no effects on absolute liver weight, histological changes (slight hepatocyte vacuolization consistent with fatty change) in females at ≥ 40 mg/m 3 , and gross paleness of the liver in 3/10 females at 119 mg/m 3 . Hepatocytes in the control and 13-mg/m 3
females were vacuolated to a somewhat lesser degr ee (i.e., very slight), in dicating that the effect

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was only minimally increased at ≥ 40 mg/m 3 . No other biologically significant effects were found in clinical chemistry, hematology, and urine chemistry. The low concentration of DGBE (i.e., 13 mg/m 3 ) is identified as a NOAEL based on si gnificant incidences of hepatocyte

vacuolization consistent w ith fatty change that occurred in female rats at ≥ 40 mg/m 3 and is supported by increased relative liver weight at 119 mg/m 3

.

Other Studies

There are several short-term and subchronic toxicity studies (Proctor and Gamble, 1982; Bio/dynamics, Inc., 1989; Auletta et al., 1993), repr oductive toxicity studie s (Bio/dynamics, Inc., 1989; Auletta et al., 1993), and a subchronic neur otoxicity study (Bio-Res earch 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 th e urine) reported by Bio/dynamics, Inc. (1989) and Aulett a et al. (1993) are consistent wi th 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 ex changes (SCEs) in CHO cells (Thompson et al., 1984), and weakly positive fo r forward mutations in mouse lymphoma L5178Y cells (Thompson et al., 1984). In vi vo 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 GLYCO L MONOBUTYL ETHER (DGBE)

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

several studies in rats as su mmarized in Table 1. A range-fin ding 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 effect s at 51 mg/kg-day, reduced water intake at

≥ 94 mg/kg-day, and histopathological ch anges 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 drin king 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 administ ered 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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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 ad ministered 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 identif ied a LOAEL of 50 mg/kg-day (36 mg/kg-day average daily dose) based on le ukopenia and lymphopenia in fema le rats (Hobson et al., 1987). In a subsequent comprehensive drinking water st udy, 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 ba sed on the critical

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

The database for DGBE consistently indicate s that hematological effects are the most sensitive response after exposur e to DGBE; however, inconsiste nt 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 re sponses in the lone surviving female in the highest-dose group. In addition, neither leukopen ia nor lymphopenia was reported in the same strain of rats in a 13-week drinking water study (Johnson et al., 2005) at do se levels as high as 1,000 mg/kg-day, and no similar effects were fou nd 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, consiste nt 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 th e DGBE-induced hematological response.

There are no indications that reproductive or developmental toxi city 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 e ffect 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/dyn amics, 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 dos es of 50, 250, or 1,000 mg/kg-day (Johnson et al., 2005). A dermal study in rats also showed no ef fects on a number of ne urotoxicity 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 fo r 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). Becau se the changes in Hgb levels in females were comparable to those in males (s ee Table 3), a POD for th is endpoint is expected to be at the range close to 328 mg/kg-day base d 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 derivati on 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 differ ences 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 extrapolati on is not needed because the POD is a

BMDL.

• An UF of 3 (10 0.5 ) 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 re productive 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 cons istent with that found from other short-term studies (Kodak, 1984, Hobson et al., 1987). However, a 13-w eek 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. Instea d, 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 st udies (Johnson et al., 2005; Hobson et al., 1987), a lack of a multigeneration reproduction study, a nd 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 differ ences 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 extrapolati on is not needed because the POD is a

BMDL.

• An UF of 10 was applied for using a st udy with a subchronic duration of exposure to approximate chronic exposure.

• An UF of 3 (10 0.5 ) is applied to account for deficiencies in the database. The

database includes critical three subchr onic studies in rats, two developmental studies in rats and mice, and a one-gener ation 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 (Johns on et al., 2005; Hobson et al., 1987), a lack of chroni c oral toxicity studies, and a lack of a multigeneration reproduc tion study. Overall, confidence in the chronic p-RfD is low.

DERIVATION OF PROVISIONAL SUBCHRON IC AND CHRONIC INHALATION RfC

VALUES FOR DIETHYLENE GLYCO L MONOBUTYL ETHER (DGBE)

Information on the inhalation toxi city 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/m 3 for 6 hours/day, 5 days/week (Gushow et al., 1984). Evaluation of clinical condition, body weight, hematology (including RBC fragility), blood a nd urine chemistry, organ weights, gross pathology, and histopathology (inclu ding nasal turbinates and lungs) showed slight hepatocyte

vacuolization consistent with fatty change in female rats at concentrations of ≥ 40 mg/m 3 and increased relative liver weight and gr oss paleness of the liver at 119 mg/m 3 . Although there were inconsistent changes (decreases) in relative liver weight in male rats at ≥ 40 mg/m 3 with no effects on absolute liver weight, the hepatic change s in female rats were considered biologically significant. The potential for he patic effects is supported by findi ngs 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; Sm yth and Carpenter, 1948). Exposure to concentrations higher than 119 mg/m 3 could not be tested because this level was the maximum sustainable vapor conc entration of DGBE.

Subchronic and chronic p-RfCs were derive d 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 in cidence 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 concentrati on (HEC) (see Table 4) were calculated before BMD modeling. An example of the duration adjustme nt calculation is presented below:

NOAEL ADJ = NOAEL mg/m 3 × 6/24 hr × 5/7 d = 13 mg/m 3 × 6/24 hr × 5/7 d = 2.3 mg/m 3

For purposes of deriving a p-RfC based on extr arespiratory effects, DGBE was treated as a Category 3 gas. The HEC was calculated assuming periodicity was attained using the following equation:

NOAEL HEC = NOAEL ADJ × (H b/g ) A /(H b/g ) H

where (H b/g ) A /(H b/g ) H is the ratio of blood:gas (air) partiti on coefficients of the chemical in the test animals and humans. Since a blood:gas pa rtition coefficient is not available for DGBE in humans or rats, a unity value is assumed fo r the ratio (U.S. EPA, 1994b). The BMD modeling

resulted in an estimated BMC 10ADJ of 1.7 mg/m 3 and BMCL 10ADJ of 0.32 mg/m 3 , and the

BMCL 10ADJ is equivalent to BMCL 10HEC

.

Table 4. Data Sets for Hepatic Changes in Female Rats

Exposed to DGBE Through Inhalation for 5 Weeks a

Concentration (mg/m 3 ) Endpoint 0 13 40 119

Females Duration-adjusted concentration (mg/m 3 ) b 0 2.3 7.1 21.3

Human equivalent concentration (mg/m 3 ) c 0 2.3 7.1 21.3

# Examined 3 4 10 10

Slight vacuolization consistent with fatty change 3 4 9 d 10 d

a Gushow et al. (1984)

b The duration adjustment was calculated as follows:

Conc ADJ = 119 mg/m 3 × 6/24 hr × 5/7 d = 21.3 mg/m 3

c HEC = Conc ADJ × (H b/g ) A /(H b/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 BMCL 10HEC of 0.32 mg/m 3 as follows:

Subchronic p-RfC = BMCL 10HEC ÷ UF = 0.32 mg/m 3 ÷ 300

= 0.001 or 1 × 10 -3 mg/m 3

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

• An UF of 3 (10 0.5 ) is applied for interspecies extr apolation 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 defi ciencies in the database. The database includes only one 5-week study; it lacks developmental toxicity studies and a multigeneration reproduction study.

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

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

The principal study was well conducted with re spect 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-gene ration 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 e ndpoints 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 BMCL 10HEC of 0.32 mg/m 3 as follows:

Chronic p-RfC = BMCL 10HEC ÷ UF = 0.32 mg/m 3 ÷ 3,000

= 0.0001 mg/m 3 or 1 × 10 -4 mg/m 3

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

• An UF of 3 (10 0.5 ) is applied for interspecies extr apolation 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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• 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 st udy with a subchronic duration of exposure to approximate chronic exposure.

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

• An UF of 10 is applied to account for defi ciencies 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 sa me reasons as described above for the subchronic RfC, with the additional shortcom ing that chronic studies are al so lacking from the database.

PROVISIONAL CARCINOGENICITY ASSESSMENT FOR

DIETHYLENE GLYCOL MONO BUTYL ETHER (DGBE)

Weight-of-Evidence Descriptor There is no adequate information on the carci nogenicity of DGBE due to a lack of oral, inhalation, or dermal studies in animals longer than 13 weeks in dura tion (Johnson et al., 2005; Bio/dynamics, Inc., 1989; Gushow et al., 1984; Hobson et al., 1987; Koda k, 1984). Genotoxicity testing of DGBE in vitro with bacteria and ma mmalian cells and in vivo with Drosophila and mice have yielded negative result s in all but one study (a weakly positive mutagenic response in mouse lymphoma cells in vitro) (Thomp son 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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Proctor and Gamble Co. 1982. 28-Day subchron ic percutaneous study of DGBE in rabbits. Huntington Research Center . (Cited in SRC, 1992).

Schuler, R.L., B.D. Hardin, R.W. Ni emeier et al. 1984. Results of te sting fifteen glycol ethers in a short-term in vivo reproductive toxicity a ssay. 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. Me llon 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. Furthe r experience with the range finding test in the industrial toxicology laboratory. J. Ind. Hyg. Toxicol. 30(1):63–68.

SRC (Syracuse Research Corporation). 1992. H ealth and Environmental Effects Document on Glycol Ethers. Prepared by Syracuse Research Corporation, Syracuse, NY for the Office of Health and Environmental Assessment, Envi ronmental 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. Envir on. Health Perspect. 57:105–112.

U.S. EPA. 1984. Health Effects Assessment fo r Glycol Ethers. Prepared by the Office of Health and Environmental Assessment, Envi ronmental Criteria and Assessment Office, Cincinnati, OH, for the Office of Emergency and Remedial Response, Washington, D.C.

EPA/540/1-86-052.

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U.S. EPA. 1991. Chemical Assessments and Rela ted Activities (CARA). Office of Health and Environmental Assessment, Washington, DC. April.

U.S. EPA. 1994a. Chemical Assessments and Re lated 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 Dosi metry. Office of Research and Development, Washington, DC.

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

U.S. EPA. 1997. Health Effects Assessment Su mmary Tables. FY-1997 Update. Prepared by the Office of Research and Development, Na tional 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 Gu idance 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/dri nking/standards/dwstandards.pdf.

U.S. EPA. 2008. Integrated Risk Information Sy stem (IRIS). Online. Office of Research and Development, National Center for Envi ronmental Assessment, Washington, DC. http://www.epa.gov/iris/.

WHO (World Health Organization). 2008. On line catalogs for the Environmental Health Criteria Series. Online. http://www.who.in t/ipcs/publications/ehc/e hc_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.

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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 da ta 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 li near 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 acc ount 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 wh ile 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 dropp ed, if appropriate, and the entire procedure is repeated. If no fit is obtained after dropping the highest dose, the next hi ghest dose is dropped, if appropriate, and the procedure is repeated. Dose-dropping conti nues until (1) adequate fit is obtained; (2) there are only contro ls and two dose groups remaining; or (3) it is inappropriate to continue dropping doses due to a lack of statistical significan ce 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 consid ered to be amenable to BMD modeling.

Model Fitting Results for Decreased RBCs in Male Rats (Johnson et al., 2005) Following the above procedure, continuous-var iable models in the EPA Benchmark Dose Software (BMDS) (version 2.1) were fit to the da ta shown in Table 3 (main text) for decreased RBCs in male rats. In the absence of a biolog ically 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 Ta ble A-1, all models provi de 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.

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Table A-1. Model Predictions for Decreased RBCs in Male Rats a

Data Set/Model Variance

p -Value b Means

p -Value b AIC BMD 1SD (mg/kg-day) BMDL 1SD (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

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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 800Mean Response

Dose Hill Model with 0.95 Confidence Level

16:27 07/10 2008 BMD BMDL Hill

1000

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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

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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

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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

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Model Fitting Results for Decreased RBCs in Female Rats (Johnson et al., 2005) Following the above procedure, continuous-var iable 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 leve l, 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 fo r Decreased RBCs in Female Rats a

Data Set/Model Variance

p -Value b Means

p -Value b AIC BMD 1SD (mg/kg-day) BMDL 1SD (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

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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 800Mean Response

Dose Linear Model with 0.95 Confidence Level

16:32 07/10 2008 BMD BMDL Linea r

1000

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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.

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

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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

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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

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Model Fitting Results for Decreased Hem oglobin 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 re sponse 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 m odels provide identical lin ear fit to the data.

Fit to the Hill Model cannot be tested due to insu fficient 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 Rats a

Data Set/Model Variance

p -Value b Means

p -Value b AIC BMD 1SD (mg/kg-day) BMDL 1SD (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

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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 1000Mean Response

Dose

16:41 07/10 2008 BMD BMDL

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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

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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

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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

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Model Fitting Results for Decreased Hemogl obin 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 ade quate fit to the data, ev en 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 D ecreased Hemoglobin in Female Rats a

Data Set/Model Variance

p -Value b Means

p -Value b AIC BMD 1SD (mg/kg-day) BMDL 1SD (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

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Modeling Procedure for Hepatocyte Vacuolization in Females (Gushow et al., 1984) The BMD modeling for dichotomous data wa s conducted with the EPA’s BMD software (BMDS version 2.1). For all the dichotomous data, the incidence data on hepatocyte vacuolization with duration-adju sted 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/m 3 ). Therefore, the BMCL of 0.32 mg/m 3 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 Rats a

Data Set/Model Goodness-of-Fit p -Value b AIC Scaled Residual BMC 10 (mg/m 3 ) BMCL 10 (mg/m 3 )

Gamma c 0.9917 38.18 0.001 2.0 0.32

Logistic 0.8180 36.58 0.367 0.9 0.52

Log-logistic d 0.8523 38.24 0.053 2.1 0.50

Multistage e 0.9889 36.20 -0.119 1.7 0.32

Probit 0.8330 36.54 0.338 0.8 0.55

Log-probit d 0.9398 38.19 0.017 2.1 0.60

Weibull c 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

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0 0.2 0.4 0.6 0.8

1

0 5 10 15 20Fraction 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 fr om the control, and are in units mg/m 3

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0 0.2 0.4 0.6 0.8

1

0 5 10 15 20Fraction Affected

dose

Figure A-5. Fit of Multistage Model to Data Multistage Model with 0.95 Confidence Level

14:53 07/28 2009 BMD BMDL 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 fr om the control, and are in units mg/m 3

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

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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

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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